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L-2985 Luxembourg OF THE EUROPEAN COMMUNITIES OFFICE FOR OFFICIAL PUBLICATIONS 9 ISBN 92-894-1735-8 7 8 9 2 8 9 4 1 7 3

5

8 KH-40-01-448-EN-N 14 http://europa.eu.int/comm/environment/pubs/home.htm See our publications catalogue at: European Commission and in urban A great deal of additional information on the European Union is available on the Internet. It can be accessed through the Europa server (http://europa.eu.int).

Cataloguing data can be found at the end of this publication.

Luxembourg: Office for Official Publications of the European Communities, 2001

ISBN 92-894-1735-8

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1 Executive Summary

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

Pollutants in Urban Waste Water and 2 Executive Summary

Table of Contents

EXECUTIVE SUMMARY 5

1. INTRODUCTION 9

1.1 INTRODUCTION TO POLLUTANTS IN URBAN (UWW) AND SEWAGE SLUDGE (SS) 1.2 OBJECTIVES AND GOALS

2. POTENTIALLY TOXIC ELEMENTS, SOURCES, PATHWAYS, AND FATE THROUGH URBAN 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 2.2 INFLUENCE OF VARIOUS TREATMENT PROCESSES ON THE FATE OF POTENTIALLY TOXIC ELEMENTS THROUGH WASTEWATER TREATMENT SYSTEMS (WWTS) AND (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

Pollutants in Urban Waste Water and Sewage Sludge 3 Executive Summary

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 IN URBAN ENVIRONMENT (B) SUSTAINABLE URBAN DRAINAGE (C) 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 AND SEWAGE SLUDGE (G) USE OF POLYELECTROLYTES; THE ACRYLAMIDE MONOMER IN WATER TREATMENT (H) CASE STUDY: (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

Pollutants in Urban Waste Water and Sewage Sludge 4 Executive Summary

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 major objective of the new Water Framework Directive (2000/60/EC) is the long-term progressive reduction of contaminant discharges to the aquatic environment in urban wastewater (UWW). Sewage sludge is also a product of wastewater treatment and the Urban Waste Water Treatment Directive (91/271/EEC) aims to encourage the use of sludge whenever appropriate. Potentially toxic elements and hydrophobic organic contaminants largely transfer to the sewage sludge during waste water treatment with potential implications for the use of sludge although some may be emitted with the water.

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

Platinum and Pd are components of vehicle catalytic converters and emissions occur as the autocatalyst deteriorates. Catalytic converters are the main source of these metals emitted to 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 sewage sludge in a similar way to other potentially toxic elements. The Pt content in sludge is typically in the range 0.1 – 0.3 mg kg-1 (ds) and the background value for is 1 mg kg-1. PGMs are inactive and immobile in soil.

In contrast to potentially toxic elements, inputs of the main persistent organic pollutants of concern, including: PAHs, PCBs and PCDD/Fs, to UWW are principally from atmospheric onto paved surfaces and runoff. Combustion from traffic and commercial sources accounts for the major PAH release to the environment, although inputs from preparation sources also represent an important and often under-estimated contribution of certain PAH congeners. PCDD/Fs are released during waste and also by coal combustion. Soil acts as a long-term repository for these contaminant types and remobilisation by volatilisation from soil is an important mechanism responsible for and redistributing them in the environment. For example, the industrial use of PCBs was phased out in Europe during the 1980s-1990s, but 90 % of the contemporary emissions of PCBs are volatilised from soil. Since emission controls are already in place for the main point sources and PAHs, PCDD/Fs or PCBs enter UWW principally from diffuse atmospheric deposition and environmental cycling, there is probably little scope, from source control, to further reduce inputs and concentrations of these persistent organic substances in UWW or sewage sludge.

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

5 Executive Summary

Potentially toxic element contamination of urban wastewater and sewage sludge is usually attributed to discharges from major commercial premises. However, significant progress has occurred in eliminating these sources and this is reflected in the significant reductions in potentially toxic element concentration in sewage sludge and surface reported in all European countries where temporal data on sludge and have been collected. However, potentially toxic element concentrations remain higher in sludge from large urban wastewater treatment (WWTP) compared with small WWTP and they are also greater in from industrial catchments compared with rural locations. These patterns in sludge content suggest that commercial sources may still contribute significantly to the total metal load entering UWW. Indeed, recent regional surveys of metal emissions from commercial premises confirm that further reductions in most elements could be achieved from this sector. The primary targets for source control include health establishments, small manufacturing industries (particularly metal and vehicle related activities) and hotel/catering enterprises, as 30 % of medical centres and 20 % of the other types of activity could be discharging significant amounts of potentially toxic elements in UWW. is a specific case where compulsory use of dental amalgam separators, and substituting Hg with alternative thermoreactive materials in thermometers, may be effective in reducing discharges 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 industrial premises. Faecal matter typically contains 250 mg Zn kg-1, 70 mg Cu kg-1, 5 mg Ni kg-1, 2 mg Cd kg-1 and 10 mg Pb kg-1 (ds). The other principal sources of metals in domestic wastewater are body care products, pharmaceuticals, cleaning products and liquid . is the main source of Cu in hard water areas, contributing >50 % of the Cu load and Pb inputs equivalent to 25 % of the total load of this element have been reported in districts with extensive networks of Pb pipework for water conveyance. Adjusting water hardness in order to reduce metal solubilisation from plumbing is technically feasible, but is likely to be impractical at the regional scale necessary to significantly reduce metal concentrations in UWW and sludge and may be unpopular with consumers in hard water areas. The gradual replacement of Pb water pipes can be achieved during building renewal and renovation programmes.

Reductions in domestic discharges of metals may be possible through increased public awareness of appropriate liquid waste disposal practices and the provision of accessible liquid waste disposal facilities. It may be impractical to eliminate the use of metals in body care products when they are an important active ingredient, but advice and labelling could be improved to minimise excessive use. may be a contaminant present in phosphatic and removing phosphate from detergent formulations can reduce associated potential discharges of Cd from domestic sources.

Detergent residues (e.g.nonyl , NP), surfactants (e.g. linear alkyl sulphonates, LAS), plasticising agents (e.g. di-(2-ethylhexyl)phthalate, DEHP) and polyacrylamide compounds, added to sludge to aid dewatering, are quantitatively amongst the most abundant organic contaminants present in UWW and/or sewage sludge. Dewatering agents based on polyacrylamide may contain traces of the potentially toxic acrylamide monomer, but this is rapidly degraded and polyacrylamide itself is biologically inactive. Detergent residues and DEHP are primarily of domestic origin and they are effectively degraded during aerobic wastewater treatment and are not considered to represent a potential environmental problem from the of treated to surface waters. is the principal method employed for stabilising sewage sludge, but NP accumulates during anaerobic digestion, DEHP is not removed by this conventional process and, although a significant amount of LAS is biodegraded, residues of this substance remain because of the large concentrations initially present in raw sludge. The inability to degrade detergent residues anaerobically and the large concentrations present in sludge and UWW have prompted ecolabelling initiatives in a number of European countries to influence consumer choice away from detergents containing these surfactants to

Pollutants in Urban Waste Water and Sewage Sludge 6 Executive Summary alternative products. This has been successful when supported by extensive public awareness campaigning. For example, the market share for ecolabelled detergents in increased to 95 % and the consumption of LAS has decreased to a similar extent. Surfactant residues and plasticisers degrade quickly when added to aerobic . The oestrogenic activity of NP is however, a principal concern and measures are proposed to eliminate the discharge of this substance to UWW.

Natural and synthetic oestrogens are degraded in WWT, but trace amounts remain and represent the main source of oestrogenic activity in treated effluents. Further work is necessary to link these substances to oestrogenic responses in aquatic , but it may be necessary 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 are identified as being potentially resistant to wastewater and sewage sludge treatment and the most significant of these are brominated diphenyl ethers (PBDEs) and chlorinated paraffins. Further research is warranted, in particular to assess the persistence and potential environmental significance of these compounds. Synthetic nitro musks are used in perfumed products and traces may be present in UWW and sludge. Little is known about the environmental fate of these compounds, but effects on human health from this route seem unlikely given that the main exposure route is through direct contact.

The degree of removal and of pharmaceutical compounds during WWT varies considerably, although many common rapidly biodegrade. They are soluble and transfer to sludge is only of minor concern. Significant amounts of prescribed drugs are excreted from the body and controlling these inputs from the general population would be impractical. However, the disposal of unused drugs into UWW should be reviewed and alternative methods of disposal should be encouraged. The potential significance of pharmaceuticals in the environment should be assessed in context of the major inputs and presence arising from widespread veterinary administration of drugs to livestock and farm waste disposal to .

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 to UWW or sewage sludge and the subsequent implications for the environment. Specified criteria regarding and biodegradation could be set for compounds that exhibit a propensity to enter the WWTS and restrictions could be enforced regarding production and use if these were not met. These decisions would need to balanced against the potential benefits to health derived from the administration of pharmaceutical drugs.

Strategies aimed at controlling pollutant discharges can only focus on those sources that can be identified and quantified. Published mass balance calculations indicate there is a high degree 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 load entering 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 variable inputs of metals in urban runoff and the underestimation of discharges from commercial premises 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 these chemicals to the environment. Current developments also suggest that Zn, Cu and LAS may be the most limiting constituents in sludge if the proposed maximum permissible concentrations for these substances in soil (Zn and Cu) and sludge (LAS) are carried through 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

Pollutants in Urban Waste Water and Sewage Sludge 7 Executive Summary minimise their to UWW as far as is practicable and to ensure there is a consistent link and approach to defining the environmental quality standards for sludge with those for sustainable water use and contaminant discharge reduction.

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

· To reduce the uncertainty in quantifying contaminant discharges to UWW by identifying and surveying specific sources to determine the potential for controlling inputs particularly from small commercial sources and medical establishments;

· To establish the extent and variability of contaminant entry into UWW by catchment investigations in relation to frequency and changes in sludge quality;

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

· To determine the extent of volatilisation-deposition cycling of persistent organic pollutants in the environment, identifying the processes controlling the extent and magnitude of diffuse inputs of these substances to UWW and to provide long-term predictions of changes in release patterns and the consequences for UWW and sludge;

· To develop a consistent statistical and reporting protocol for national chemical composition data presented in surveys of sewage sludge quality.

Pollutants in Urban Waste Water and Sewage Sludge 8 1. Introduction

1. INTRODUCTION

The primary objective of this study is to determine the sources of in urban wastewater (UWW) treated in wastewater treatment systems (WWTS). This includes the pollutants introduced into the UWW collecting system with run-off rainwater, from domestic and small commercial sources. The pollutant contents in urban wastewater and sewage sludge has been evaluated by review of the existing literature, in order that measures may be 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), (Cr III and Cr VI), copper (Cu), mercury (Hg), nickel (Ni), (Pb) and zinc (Zn),

· organic pollutants including PAHs, PCBs, DEHP, LAS, NPE, dioxins (PCDD) and furans (PCDF). Over 6,000 organic compounds have been detected in raw water sources most of which are due to human activities. While some of these are highly persistent, others are easily biodegradable in WWTS.

Other pollutants of interest are the metalloids, 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 main categories: · Domestic, · Light industrial (connected to the WWTS) and commercial, · Urban runoff (which also encompass lithospheric and atmospheric sources).

Pollutants in Urban Waste Water and Sewage Sludge 9 1. Introduction

Atmosphere wet & dry deposition deposition

Lithosphere

INDUSTRY

products

DOMESTIC Wastes Product wastes RUNOFF

UWW COLLECTING COMBINED UWW STORM UWW Wastes SYSTEMS COLLECTING SYSTEMS COLLECTING SYSTEMS

WASTEWATER TREATMENT WORKS

Receiving Water

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

A summary of the various inputs, outputs and pathways followed by water and associated contaminants from both natural and anthropogenic sources encountered in urban environments 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 point sources. Large point sources are easily quantifiable and result from specific activities in the area that are connected to UWW collecting systems. The contribution from small point sources, such as households and small businesses, is much more difficult to identify and quantify, compared to point sources which are usually regulated. UWW is also vulnerable to illegal discharges of pollutants.

Diffuse sources, such as atmospheric deposition and road runoff have also been characterised and this study will attempt to present an overview of the available information in 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 diffuse sources, can link observed water quality trends to changes in socio-economic activities.

Pollutants in Urban Waste Water and Sewage Sludge 10 1. Introduction

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

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

A review of the sources and pathways of potentially toxic element pollutants in urban wastewater 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 in domestic, commercial, and urban run-off wastewater, which end up in the UWW collecting system. · To make a qualitative and quantitative assessment of the pollutants in urban wastewater and runoff rainwater on the basis of the available data in the literature. · To evaluate the percentage of inorganic and organic pollutants concentrated in sewage sludge and the percentage of pollutants released in the environment with the treated effluents. · To review wastewater and sewage sludge treatment processes and possible measures to prevent pollution at source. The most important practices to treat wastewater and sewage sludge in Europe will be closely examined. · Based on an overall assessment of the existing data from various sources, to identify further research directions in those areas with insufficient data.

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

Pollutants in Urban Waste Water and Sewage Sludge 11 2. Potentially Toxic Elements

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

The aim is to reduce inputs of pollutants entering the wastewater system to background levels because this represents the minimum potential extent of contamination that can be achieved. Potentially toxic elements are of concern because of their potential for long-term accumulation in soils and .

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

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

Pollutants in Urban Waste Water and Sewage Sludge 12 2. Potentially Toxic Elements

2.1. Sources and pathways of potentially toxic elements in UWW

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

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

Element Domestic Commercial Wastewater [mg.l-1] Wastewater [mg.l-1]

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

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

Pollutant Country Domestic Commercial Urban Not Reference Wastewater Wastewater Runoff Identified Cd 20 61 3 16 ADEME, 1995 Norway 40 SFT report 97/28 UK 30 29 41 WRc, 1994 Cu France 62 3 6 29 ADEME, 1995 Norway 30 SFT report 97/28 UK 75 21 4 WRc, 1994 Cr France 2 35 2 61 ADEME, 1995 Norway 20 SFT report 97/28 UK 18 60 22 WRc, 1994 Hg France 4 58 1 37 ADEME, 1995 Pb France 26 2 29 43 ADEME, 1995 Norway 80 SFT report 97/28 UK 43 24 33 WRc, 1994 Ni France 17 27 9 47 ADEME, 1995 Norway 10 SFT report 97/28 UK 50 34 16 WRc, 1994 Zn France 28 5 10 57 ADEME, 1995 Norway 50 SFT report 97/28 UK 49 35 16 WRc, 1994

Pollutants in Urban Waste Water and Sewage Sludge 13 2. Potentially Toxic Elements

Cd distribution Cu distribution

Domestic Domestic Storm events Storm events Commercial Commercial Non Identified Non Identified

Cr distribution Hg distribution

Domestic Domestic Storm events Storm events Commercial Commercial Non Identified Non Identified

Pb distribution Ni distribution

Domestic Domestic Storm events Storm events Commercial Commercial Non Identified Non Identified

Zn distribution

Domestic Storm events Commercial Non Identified

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

Pollutants in Urban Waste Water and Sewage Sludge 14 2. Potentially Toxic Elements

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

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

In the , 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 of potentially toxic elements in the UWW system is expected to decrease, mainly due to a decrease in runoff and industrial sources, while the potentially toxic elements share in WWTS loads from households was expected to increase. As industrial sources of potentially toxic 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. Their studies showed a significant decrease (50-90%) in the potentially toxic element content of sewage 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 current regulations require monitoring of the influent, in order to set up a baseline database from which limits may then be devised.

Pollutants in Urban Waste Water and Sewage Sludge 15 2. Potentially Toxic Elements

2.1.1 Domestic sources

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

A study by RIVM (Dutch Institute of and the Environment) in the Netherlands [SPEED, 1993], quantified the waterborne emissions of potentially toxic elements from household sources, dentistry and utility buildings in the urban environment. Table 2.3 shows the 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].

Potentially Gross waterborne emissions* tonnes.y-1 to surface toxic element water (1993) Household Dentistry Utility buildings sources Copper 94 0.6 27 Zinc 118 - 26 Lead 13 - 3.1 Cadmium 0.7 - 0.2 Nickel 7.3 - 0.9 Chromium 2.9 - 0.3 Mercury 0.3 2.3 0.01 * 96 % of the waterborne emissions are expected to go to the UWW collecting systems, with 4% going directly to surface waters.

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

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

Copper: comes mainly from corrosion and of plumbing, fungicides (cuprous chloride), pigments, preservatives, larvicides (copper acetoarsenite) and antifouling paints.

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 dental amalgams. Also, mercury can still be found as an additive in old paints for water proofing and marine antifouling (mercuric arsenate), in old (mercuric chloride in fungicides, insecticides), in wood preservatives (mercuric chloride), in embalming fluids (mercuric chloride), in germicidal soaps and antibacterial products (mercuric chloride and mercuric cyanide), as mercury-silver-tin alloys and for "silver mirrors".

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

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

Pollutants in Urban Waste Water and Sewage Sludge 16 2. Potentially Toxic Elements for uses in glazes), also in "crystal glass". Lead has also been found in wines, possibly from the lead-tin capsules used on bottles and from old wine processing installations.

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

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

Arsenic and Selenium: are among the potentially toxic metalloids found in urban wastewaters. 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 from natural background sources and from household products such as products, medicines, garden products, wood preservatives, old paints and pigments. Selenium comes from food products and food supplements, shampoos and other cosmetics, old paints and pigments. Arsenic is present mainly as DMAA (dimethylarsinic acid) and as As (III) (arsenite) in urban effluents and sewage sludge [Carbonell-Barrachina et.al., 2000].

Pollutants in Urban Waste Water and Sewage Sludge 17 2. Potentially Toxic Elements

Household products

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

Product Zinc Copper Cadmium Nickel (µg g-1) (µg g-1) (µg g-1) (µg g-1) Washing Powders a 37.9 1.4 74.3 <0.5 ‘Big Box’ b 35.9 <0.5 136.0 <0.5 c 3.3 <0.5 6.6 <0.5 Washing Powders ‘Ultra’ a <0.1 <0.5 24.0 <0.5 b 2.3 1.40 10.6 <0.5 c 1.0 1.38 11.8 <0.5 Fabric Conditioners a 0.1 <0.5 9.4 0.6 b <0.1 <0.5 9.0 <0.5 c 0.1 <0.5 10.7 <0.5 Hair Conditioners a <0.1 <0.5 16.8 <0.5 b 1.0 1.4 17.2 <0.5 c 1.7 <0.5 8.6 <0.5 d 0.5 1.4 68.0 1.0 Cleaners a 0.3 2.8 26.0 <0.5 b <0.1 <0.5 17.8 <0.5 Shampoo (medicated) a 4900 1.4 17.4 <0.5 Washing Up Liquid a 0.2 1.1 11.0 0.8 Bubble bath a 0.2 <0.5 13.6 <0.5 b <0.1 1.4 10.4

As can be seen from Table 2.4 there is a great deal of variability between products and also between 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 be explained 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 on source [Hutton et al reported in WRc report 1994]. Reducing the amount of phosphate in washing powders, or choosing phosphate ores with low Cd concentration could lead to a reduction in Cd in wastewater from diffuse sources. In Sweden the amount of cadmium 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 substantially reduced due to the switch to phosphate-free detergents [SPEED, 1993]. The 'Ultra' washing powders, usually phosphate-free, have smaller potentially toxic element contents than the traditional powders, and are designed to be used in smaller quantities. A shift to these newer products 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 thus raise the zinc inputs to the UWW collecting system. In 1991 these shampoos were estimated to represent 26% of the market [*BLA Group 1991- reported in Comber and Gunn 1996 and WRc 1994]. Cosmetics are not included here but they may also contain high levels of zinc and 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 toxic elements in domestic wastewater came from cosmetic products, medicines, cleaning products and liquid wastes (including paint), which were directly discharged from the household sink.

Pollutants in Urban Waste Water and Sewage Sludge 18 2. Potentially Toxic Elements

Table 2.5 provides a general picture of some of the potentially toxic elements in various domestic 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 in this 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 more information 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 Zn Amalgam fillings Ö and thermometers Cleaning products Ö Ö Cosmetics, Ö Ö Ö Ö Ö Ö Ö shampoos Ö Fire extinguishers Ö Ö Ö Ö Ö Inks Ö Ö Ö Ö Ö Medicines and Ö Ö Ö Ö Ö Ointments Health supplements Ö Ö Ö Ö Ö Food products Ö Ö Ö Ö Ö and lubricants Ö Ö Ö Ö Paints and Ö Ö Ö Ö Ö Ö Ö Ö Ö Ö pigments Photographic Ö Ö Ö (hobby) Polish Ö Ö Ö Pesticides and Ö Ö Ö Ö Ö gardening products

Washing powders Ö Ö Ö Wood- Ö Ö Ö preservatives Other sources Faeces and Ö Ö Ö Ö Ö Ö Ö Ö Ö Tap Water Ö Ö Ö Ö Ö Water treatment Ö Ö Ö Ö Ö and heating systems

Domestic activities

The main domestic sources of potentially toxic elements in wastewater were estimated by WRc [1994] to be (in order of importance): cadmium: faeces > bath water > laundry > tap water > kitchen chromium: laundry > kitchen > faeces > bath water > tap water copper: faeces > plumbing >tap water > laundry > kitchen lead: plumbing > bath water > tap water > laundry > faeces > kitchen nickel: faeces > bath water > laundry > tap water > kitchen zinc: faeces > plumbing > tap water > laundry > kitchen.

Pollutants in Urban Waste Water and Sewage Sludge 19 2. Potentially Toxic Elements

Estimates of the mean potentially toxic element inputs to UWW collecting systems from domestic 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 major contributors to the overall level of potentially toxic elements reaching the WWTS. Most of the zinc 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 rather than from plumbing.

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

Activity (study in a hard Load (µg.person-1.day-1) water catchment area) Zn Cu Pb Cd Ni Cr Washing Machine Input Water 662 6859 36.0 0.6 27 4 Washing 4452 977 515 11 52 238 Dishwashing Input Water 39 69 2.9 0.03 2 0.3 (machine) Washing 42 8 6 1.3 2 10 Dishwashing Input Water 591 6125 32 0.5 24 3.7 (hand) Washing 1010 <20 46 7.8 138 136.7 Bathing Input Water 1140 10651 46 1.0 40 5.9 Bathing 1095 67 45 13.1 9 7.4 Input Water 2531 8082 63 2.0 77 8.5 Faeces 11400 2104 121 48.0 284 51.5 Miscellaneous Input Water 1453 6951 62.7 1.2 54 7.0

Predicted Total 24416 41894 978 86 710 464.2 Measured mean from Catchment 15314 46772 1237 71 925 686 housing estates Population 50 000 Predicted load to UWW 1.2 2.1 0.05 0.01 0.04 0.02 collecting systems from domestic sources (kg/day) Measured mean total load to 2.6 3.3 0.3 N/A 0.1 0.15 the WWTS kg per day % of potentially toxic 46.0 64.2 16.9 N/A 25.7 15.3 elements from domestic sources in the UK

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

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

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

Pollutants in Urban Waste Water and Sewage Sludge 20 2. Potentially Toxic Elements

Table 2.7 Potentially toxic elements entering wastewater, breakdown by source [WRc, 1994] Cu Zn Cd Ni Pb Cr Percentages of total load Break down of domestic sources as percentage of total metal entering WWTS Plumbing Input Bathing 12.4 1.2 0.3 1.6 4.4 0.1 (% of total entering WWTS) Toilet 8.9 2.1 0.5 4.3 8.1 0.2 Washing 8.0 0.7 0.2 0.9 3.2 0.1 Machine Miscellaneous 8.0 1.2 0.3 2.4 6.1 0.2 Dishwashing 7.1 0.6 0.2 0.8 2.7 0.1 Activities Faeces 20.6 28.0 20.0 23.6 3.1 2.1 (% of total metal entering Washing 9.53 10.8 4.6 4.4 18.6 9.3 WWTS) Machine Bathing 0.2 2.6 0.9 0.8 1.1 0.3 Dishwashing 2.6 4.6 11.2 1.1 5.1

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

Concentrations are expected to vary with the intake of metals in the diet, and and may also be influenced by the increasing prevalence of supplementation of food, for example with zinc, iron, selenium, and manganese. One study in France (ADEME 1997) found the following concentrations of potentially toxic elements in faeces (as dry matter): Zn: 250mg kg-1, Cu: 68mg kg-1, Pb: 11mg kg-1, Ni: 4.7mg kg-1, and Cd: 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 mg kg-1 in the UK) is typical of the amount originating naturally in faeces from the normal trace amounts of this element ingested in food (typically 18.8 mg d-1). Other differences in reported concentrations 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 the distribution-plumbing-heating networks contribute major inputs of Pb, Cu and Zn. Lead concentration for instance can vary between 14 µg.l-1 at the household input and 150 µg.l-1 at the output.

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

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

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

Pollutants in Urban Waste Water and Sewage Sludge 21 2. Potentially Toxic Elements supplies [Comber et al 1996]. The 'first draw' (initial flow of water in the morning) has higher amounts of copper and lead compared to subsequent draws [Isaac et.al., 1997]. The Cu content was found to be between 73.7 and 1430 µg l-1, and Pb content between 8.3 and 22.3 µg l-1, much greater than in the average effluent from households. Water treatment would be recommended for certain water domestic uses, such as boilers and heating systems, in order to reduce the metal corrosion.

The type of housing was also found to be important by the WRc report [1994]. Table 2.8 gives 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 levels from the "modern estate" can be explained by the newer plumbing system. In many countries copper is the major element used in plumbing. In the UK it has been estimated that leaching from copper plumbing accounts for over 80% of the copper entering domestic wastewater [Comber et al 1996]. The higher lead level found in the newer housing did not correlate with similar studies comparing old and new housing and could not be explained satisfactorily; as in general older houses in the UK contain more lead plumbing. Lead from solders in the piping system may also be an important source. In other regions of the EU steel and zinc galvanised iron are used widely. This may explain why zinc in sludge is proportionally greater than Cu in other member states compared with the UK.

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

Potentially toxic Zn Cu Ni Cd Pb Cr Hg 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 domestic sources 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 recycling schemes · how much of the products are discharged into wastewater.

Pollutants in Urban Waste Water and Sewage Sludge 22 2. Potentially Toxic Elements

2.1.2 COMMERCIAL SOURCES

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

Cadmium could originate from laundrettes, small electroplating and coating shops, manufacture, 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 coloured pigments.

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

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

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

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

Nickel is used in the production of alloys, electroplating, catalysts and nickel-cadmium batteries. 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 occurs in 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 towards electrolytic production of zinc which, in contrast to thermally produced zinc, has virtually no cadmium contamination. This means that cadmium pollution to UWW due to the corrosion of galvanised steel will in time become negligible. [SPEED 1993].

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

Pollutants in Urban Waste Water and Sewage Sludge 23 2. Potentially Toxic Elements

Silver could potentially be emitted from photographic and printing shops, from jewellery manufacturers 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 Zn in urban wastewater and in the coastal environment [Castro, et.al., 1996], with large concentrations 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 in different 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 Reference (mg l-1) Cd Germany All sectors 3-1,250 Wilderer et al 1997 Greece Petroleum industries 300-400 NTUA, 1985 Cu Germany All sectors 37-26,000 Wilderer et al 1997 Greece Metal and electrical industries 5,000-10,000 NTUA, 1985 Italy Artisanal galvanic shops 20,500 EBAV, 1996 Goldsmiths and jewellery 700-1,900 shops (max.13,300) Cr Germany All sectors <10-20,100 Wilderer et al 1997 Greece Metal and electrical industries 500-13,000 NTUA, 1985 Tanneries 100-7,000,000 Italy Artisanal galvanic shops 16,000 EBAV, 1996 Pb Germany All sectors <50-13,400 Wilderer et al 1997 Greece Metal and electrical industries 500 NTUA, 1985 Italy and photoceramics 6,000 EBAV, 1996 shops Ni Germany All sectors <10-7,300 Wilderer et al 1997 Greece Metal and electrical industries 500-14,500 NTUA, 1985 Italy Artisanal galvanic shops 19,700 EBAV, 1996 Zn Germany All sectors 30-133,000 Wilderer et al 1997 Greece Metal and electrical industries 60-2,830 NTUA, 1985 Italy Goldsmiths and jewellery 1,000 (max. 7,000) EBAV, 1996 shops As Spain Paper mills 3.4 Navarro, et.al 1993

In 1999, a project carried out by Anjou Recherche [LIFE, 1999], attempted to classify all commercial sources of wastewater pollution based on a matrix of 73 main pollutants including many potentially toxic elements and certain organic pollutants. The UWW collecting 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 were recorded, of which health and social action (33.5%), manufacturing industry (20%), hotels and restaurants (17.8%), and collection services (10%) appeared most often as potential polluters. It was found on average, that between a third and a half of the activities emitted pollutants. In the Louviers area, it was found that 53 urban businesses and institutions could potentially emit cadmium, 168 chromium, 147 copper, 35 nickel, 167 mercury, 50 lead, and 63 zinc. This suggests that more can be done to reduce trade effluent discharges.

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

Pollutants in Urban Waste Water and Sewage Sludge 24 2. Potentially Toxic Elements

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

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

Motor industry - vehicle washing

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

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

A study by the Norwegian Pollution Control Authority [SFT, 1999] examining potentially toxic element pollutants in Norway found that out of six petrol stations investigated, only one had an oil separator/ trap that worked effectively. Although designed to reduce the contamination of urban wastewater, oil separating devices are generally ineffective at reducing 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 Henrikdal WWTS, Stockholm, prompted an investigation to identify potential sources (Table 2.10). It was concluded that the high mercury content of sludge was attributable to dental practices and the use of mercury in dental amalgams. Amalgam separators were ineffective at retaining 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 and thermometers, has led to a reduction of mercury input into UWW [Ulmgren, 2000a and Ulmgren, 2000b]. Mercury recycling schemes have also proved to be successful, and could be extended to other countries and activities.

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

Pollutants in Urban Waste Water and Sewage Sludge 25 2. Potentially Toxic Elements

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

Source Comments Heavy Industry no Ruled out as a source of mercury to UWW and SS as these were not connected to the UWW collecting system Small and no Ruled out as they were operated in such a way that no Medium contamination of the wastewater was likely Enterprises (18 Companies) Storm Drains yes 20% of the total mercury load came from storm drains. This was largely traced to the deposition of emitted from crematoria which are estimated to be about 50 kg of mercury a year Household yes About 15% of the mercury entered the waste water system through Wastewater the use of mercury thermometers in the home, and also from small amounts of mercury in food and amalgam fillings in teeth Dentists and yes High mercury content of sludge was largely attributable to dental Dental practices and the use of mercury in dental amalgams Technicians Hospitals yes Samples indicated that hospitals emit 10% of the mercury loading Old Sewage yes Investigations in the last few years have found many sources of Pipes mercury in old pipes

Similar findings have been reported in other countries. A WRc report [1994] established that in the UK mercury emissions are much higher from commercial, rather than domestic sources, 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 amalgam fillings corrosion. In the Louviers area of France, the analysis of wastewater and sewage sludge 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 hospital activities, 4% from veterinary activities, and 1% from ambulance activities. Thus, in this instance, targeting medical/dental practices may help reduce pollution from mercury [LIFE, 1999].

Other sources of mercury In 1993, the amount of mercury entering France was 209 tonnes, the amount leaving France was 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 recycled Zinc and lead metallurgy 18 4.8 Thermometers 9 5.6 Dentistry amalgams 9 11.4 Batteries 6.8 7.4 Laboratories 0.9 0.0 Fluorescent tubes 0.8 12.5 Barometers 0.4 10.0 High intensity lamps 0.2 10.0 Chlorine production 77 ?

Table 2.11 highlights the amount of mercury waste produced by zinc/lead metallurgy and chlorine production. In Galicia, north-western Spain, high mercury levels in UWW and sewage 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 and the lagoons of Aveira is linked with chlor-alkali production [Lucas et.al, 1986 and Pereira

Pollutants in Urban Waste Water and Sewage Sludge 26 2. Potentially Toxic Elements et.al, 1998]. There appears to be potential for improved control and recycling of mercury waste associated with these activities.

Sources of chromium

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

Table 2.12 Chromium emissions to water in Finland [adapted from Mukherjee, 1998] Source Category Emissions to water (not exclusively UWW) tonnes per annum % of total contribution Chemical Industry 14.3 58.1 Paint Manufacture 0.01 <0.1 Electroplating 0.1 0.4 Ferro-chrome and Stainless Steel 4.6 18.7 Leather Processing 5.5 22.5 Total 24.6 100.0

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

All wet-textile processing in Finland discharges its wastewater, containing chromium and other metals, to the WWTS. The textile companies studied [Kalliala, 2000] produced between 50 and 500 litres of wastewater per kg of textile produced. Wastewater analyses were carried out at six major Finnish textile companies (two of these include analysis for potentially 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 60 Copper 0.4 80 Zinc - 20

There is a very high variation in the process emissions between these two plants. Company 2 was noted to use cellulose blends while company 4 was noted to have mainly polyester and polyester blends.

Pollutants in Urban Waste Water and Sewage Sludge 27 2. Potentially Toxic Elements

Sources of lead

Data from ADEME [1995] showed that worldwide consumption of lead is around 5.4 million tonnes per year. In a Swedish study [Palm, Östlund, 1996] in the Stockholm area the total amount of lead used in products such as those listed previously, was estimated at between 44,000 and 47,000 tonnes per annum. Clearly the potential for lead entering UWW from these sources will vary greatly. The largest amount of lead that finds its way to the WWTS is likely to be contributed by piping. Estimates for the amount of lead used are 8,000 tonnes in lead jointed water pipes used inside buildings, followed by 2,000 tonnes used in lead jointed water 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 water pipelines are predominantly plastic (85% PVC and PEH), with 11% cast iron; no lead is used for pipes conveying water. The wastewater pipes for the UWW collecting system are 57% concrete and 41% plastic.

Pollutants in Urban Waste Water and Sewage Sludge 28 2. Potentially Toxic Elements

2.1.3 URBAN RUNOFF

Runoff to UWW collecting systems and waterways has been intensely studied due to its potentially high loading of potentially toxic elements [WRc, 1994]. Atmospheric inputs to the urban runoff depend on the nature of surrounding industries, on the proximity of major emission sources such as smelters and coal fired power stations and the direction of the prevailing wind. Potentially toxic element loads can be five fold greater in runoff near commercial activities, than in residential areas far from industrial emitters. Roof runoff and building runoff also contribute to the total runoff loading and may be a source of considerable amounts of potentially toxic elements such as zinc, lead, copper and cadmium. Road and roof runoff sources are particularly important during storm events, which will allow flushing of potentially toxic elements and other pollutants from surfaces. Furthermore, it is important to note 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 on traffic, material and age of roofs and other surfaces, and meteorological and environmental conditions.

Although a number of studies had focused on the effects of urban land use in the quantification 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 provides concentrations for a number of potentially toxic elements in urban runoff, as a summary of various investigations from 1975 to 1978. It is important to note that the measured concentrations differ considerably. The main sources of pollution in urban precipitation runoff can be summarised as follows [based on Mitchell, 1985]:

· Road and vehicle related pollution · Degradation of roofing materials · Construction · , vegetation and associated human activities · 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, Mance and Harman, Mattraw and Pollutant 1975 1978 Sherwood, 1977 Mean Max Mean Max Mean Max Lead 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 precipitation runoff is subject to variation, dependent on catchment characteristics, time of year and measurement procedures. Urban precipitation runoff pollution has received far less attention than other forms of urban pollution (i.e. atmospheric). Table 2.15, gives an indication of the geographic and temporal variability involved.

1 Focusing on pollutant loads in rainfall runoff

Pollutants in Urban Waste Water and Sewage Sludge 29 2. Potentially Toxic Elements

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

Site Cd Cu Pb Zn Reference Urban (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 pollution in urban precipitation in Paris. In this recent study of individual events, bulk samples were collected within four gutter pipes (roof runoff), three yard-drainage pipes (yard runoff), six (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 in respect to potentially toxic elements pollution load [after Garnaud et al., 1999]

A comparison of samples from consecutive phases of precipitation runoff (precipitation to roof runoff to urban runoff to catchment outlet) indicated that the potentially toxic element concentration 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 to this contamination, including corrosion or emission from vehicles and commercial activities. Bulk metal concentrations were similar within all urban runoff samples except for zinc and lead, which were particularly concentrated in roof runoff samples, due to their contamination by corrosion of roof materials. At all sites it appeared that bulk metal concentrations could be ranked as: Cd << Cu < Pb << Zn.

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

Pollutants in Urban Waste Water and Sewage Sludge 30 2. Potentially Toxic Elements

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

Pollutant Precipitation Run-off from Run-off from [µg/l] roofs [µg l-1] streets [µg l-1] Pb 5 104 311 Cd 1 1 6.4 Zn 5 24 603 Cu 1.5 35 108 Ni 5

A study carried out by Rougemaille (1994) analysed the wastewaters of the treatment plant in Achères in the Paris region and found that lead concentrations varied between 0.05 and 0.5mg l -1 and that the average concentration was 0.1mg l-1. These wastewaters come from four different urban areas. The study showed that lead concentrations were 3 times larger during wet weather than during dry weather, hence proving the importance of the runoff sources.

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

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

PAH Pb Cu Cd Zn (ng l-1) ( g l-1) ( g l-1) ( g l-1) ( g l-1) Mean <96 58 45 1 356 Median <74 43 33 0.74 254 Range <11-474 14-188 11-146 0.2-4.2 104-1544 SD 76 44 27 0.86 288

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

Pollutants in Urban Waste Water and Sewage Sludge 31 2. Potentially Toxic Elements

A Road and vehicle contribution

Roads are a major source of pollution in urban environments and contribute to wastewater pollution both directly and indirectly (airborne pollutants generation). Sartor and Boyd [USEPA, 1972] determined the major constituents of road related runoff to be inorganic matter, but the total mass of inorganic matter present seemed to increase as the antecedent dry period (ADP) increased. Sources of the organic and inorganic fraction of road-produced pollutants 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, tyre wear and road surface wear. Motor fuel combustion was the largest source of lead to runoff but it is on the decrease due to the gradual phasing out of leaded fuel in the EU. Other metals emitted from exhausts are zinc, chromium and more recently tin from the replacement anti-knock compounds in petrol. The presence of Zn and Cd in road surface sediments can also be explained by the addition of Zn dithiophosphate in the manufacturing of lubricating oil, Cd being present as an impurity of the original Zn. Brake lining wear contributes copper, nickel, chromium and lead to runoff. Tyre abrasion contributes to the load of zinc, lead, chromium and nickel due to the and metal oxides constituents. Cadmium in car tyres is attributed to zinc-diethylcarbonate, which is used during the vulcanisation 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 from vehicle 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

Pollutants in Urban Waste Water and Sewage Sludge 32 2. Potentially Toxic Elements

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

Sources Solids Pb Cu Cd Zn Vehicles Tyre wear 314 0.002 0.0006 0.0008 3.22 Brake 100 0.390 14.2 0.0003 2.17 linings Petrol - 13.0 - - - Road Safety - 0.002 0.0002 0.0002 0.95 fence De-icing 130 0.015 0.002 0.0007 0.002 agent 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 that a large proportion of the pollutants released do not end up in the runoff waters but are probably emitted into the . The Zn content in urban runoff remains highly variable depending on the use of safety barriers made from galvanized steel or from an alternative material. All parameters are affected by traffic and by variables connected to the meteorological and environmental conditions of the sites concerned, which make the comparison highly uncertain (Montrejaud-Vignoles et al., 1996).

The following mechanisms summarise the way the pollutants are transported (pathways) from the catchment over the 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; Sylvester and DeWalle, 1972] - Removal by air-dispersal involving transfer of the surface contaminants to the atmosphere either as dry particles or dissolved in .

Note that the first two mechanisms above result in pollutants being transferred ultimately into receiving waters or sewage sludge, while the last mechanism removes pollutants prior to their ultimate disposal. The general conclusion, however, is that the majority of road contamination is associated. In particular, the prime transport mechanisms and 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 .

The transportation of surface particulates is sporadic in nature [Mitchell, 1985]. Metal levels tend to fall after periods of rain, whilst elevated concentrations have been recorded after prolonged dry periods. Also introduced in the movement patterns are local storage and residence effects due to the intrinsic configurations and micro-topography of the road surface [Harrop, 1983]. The fact that contaminants move through this sequence is well known, however, the relationship between the contaminants and the various mechanisms involved 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, as discussed earlier) may account for this underestimation.

Pollutants in Urban Waste Water and Sewage Sludge 33 2. Potentially Toxic Elements from atmosphere to receiving water body, clearly demonstrates a metal mobility evolution (Garnaud et al., 1999). Extremely labile (i.e. hydrosoluble or exchangeable), within dry atmospheric deposits Cd, Pb and Zn, become bio-available within street runoff and stable within UWW collecting system or sediment. In conclusion particulate metal mobility may be classified as: Cu<

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

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

Approximate concentration of potentially toxic elements Inorganic -1 particle size associated with particles (mg kg ) fraction (µm) Zinc Lead Copper Cadmium <40 900 920 240 24000 40-63 275 100 100 5000 63-80 300 100 125 5000 80-125 350 150 175 5000 125-250 400 200 200 5000 250-500 450 175 300 3000 500-1000 240 225 30 3000

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

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

Pollutant mean Test catchments concentrations Pleidelsheim Obereisesheim Ulm / West (mg/l) Cd 0.0059 0.0059 0.0028 Cr 0.0096 0.02 0.0052 Cu 0.097 0.117 0.058 Fe 3.42 5.16 2.18 Pb 0.202 0.245 0.163 Zn 0.36 0.62 0.32

Pollutants in Urban Waste Water and Sewage Sludge 34 2. Potentially Toxic Elements

Vehicle emissions: Prior to the lead ban in fuel3, research had identified exhaust lead emissions to account for more than 90% of atmospheric lead pollution. Although the main source of lead in the average urban atmosphere is lead from fuel additives, it appears that only 5% of the lead can be traced in runoff water. The greatest part may, therefore, disperse in the atmosphere or settle on the soil by the roadside (Hewitt and Rashed, 1990). The actual rate and form of lead emission is crucially dependent on driving conditions. High engines speeds and rapid acceleration can increase emission levels due to reactivation of particles deposited on the exhaust system. Total lead emissions can double if the engine and/or weather is cold. The deposition from petrol to the road surface was estimated to be 0.049g Pb.vehicle-1. km-1. In the 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 to legislation. In Austria for example, the lead emissions between 1985 and 1995 have been reduced by approximately 88 % [UBA, 1999]. Figure 2.4 is indicative of this period of policy change (between 1972 and 1992). Note the high correlation between lead in petrol and lead despite the large rise in traffic flow experienced in the same period.

Figure 2.4 Reduction in lead concentrations in air and petrol between 1972-1992 in 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 in detail in Section 6, case studies.

3 Amid mounting concern that the lead dispersed in the environment was causing damage to humans and the environment, a series of regulations and directives have been adopted in Europe since the 1980s, in order to phase out the use of leaded petrol. Lead content of petrol has been reduced from 0.4g l-1 to 0.15gl-1, following the Lead In Petrol Directive 85/210/EEC. Further regulations on air quality have now come into force limiting the levels of lead in air, with an attainment date of 2005. Some countries, namely: Austria, , 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 newly independent states and the Russian Federation, permitted lead levels are 0.15-0.37 g l-1. Various strategies for reduction of lead in petrol have been made in these countries, though it is expected that some will have difficulty in achieving these targets. 4 Platinum, palladium in vehicle exhaust catalysts and rhodium in three-way catalysts

Pollutants in Urban Waste Water and Sewage Sludge 35 2. Potentially Toxic Elements

Vehicle degradation:

Tyre wear releases lead, zinc and hydrocarbons, either in particulate form or in larger pieces as a result of tyre failure. The deposition of Zn on the road surface from tyres has been determined to be 0.03g vehicle km-1 (Mitchell, 1985). Metal particles, especially copper and nickel, are released by wear of clutch and brake linings. Additionally, the presence of Ni and Cr in storm runoff can result from the degradation of car bumpers and window sealings where both metals are used in the manufacture of . Copper is a common constituent of piping and other components of the engine and chassis work. The contribution to road surface material of vehicle tyres and brake linings from different road types has been summarised in Table 2.22.

Table 2.22 Input from tyre wear and brake lining degradation for each road category [after Muschack, 1990] Type Total Individual elements of street Abrasion (g ha of road-1 annum-1) (kg ha of road-1) Pb Cr Cu Ni Zn Tyre Brake Tyre Brake Tyre Brake Tyre Brake Tyre Brake Tyre Brake wear lining wear lining wear lining wear lining wear lining wear lining Residential way 137 6.8 60 7.3 10 13.7 12 210 10 51 35 0.9 Residential street 62 8.5 76 9.0 13 17.0 17 260 12 63 43 1.7 Distributor 72 12.4 112 16.6 19 24.9 26 381 18 93 63 2.4 road Main distributor 109 19.1 172 20.4 29 38.3 39 586 27 143 96 2.6 road Main road 127 30.3 266 32.2 44 60.5 60 926 42 226 149 2.1 Dual 120 43.4 382 46.3 64 86.9 72 1329 61 324 214 5.8 carriageway 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) in an experimental site in concluded that brake pad and tyre wear, caused by rapid vehicle 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 was 100 µm, but have settling times of around 24 hours under conditions. The explanation offered to this unexpected behaviour is the presence of potentially toxic elements. One of the most important findings of this study (which took into account both Australian 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 on whether or not the precipitation runoff should be treated may not be acceptable. Drapper et al. (2000) also stated that precipitation intensity and preceding dry days could be a significant factor influencing actual pollutant concentrations. It can be argued that a more suitable criterion for treatment need assessment would be the environmental significance of the receiving waters.

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

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

Pollutants in Urban Waste Water and Sewage Sludge 36 2. Potentially Toxic Elements 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 accumulated contaminants in roadside soils. In soils exposed to high Na concentrations with a subsequent supply of lower water, as in snowmelt periods and storm flow events, there is a risk of colloid dispersion and mobilization [Norrstrom and Jacks, 1998]. Soil column leaching experiments with NaCl and low-electrolyte water have provided evidence for the mobilisation of organic colloids and Fe-oxides, suggesting that potentially toxic elements may reach the via colloid-assisted transport [Amrhein et al., 1992, 1993]. Moreover, the use of road-salt may result in the increased mobilisation of potentially toxic 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-caking agent to de-icing agents, and compounds containing phosphorus may also be added as rust inhibitors. Novotny et al. (1998) argue that although ferrocyanide is non-toxic in its original form, its instability under predominant natural surface waters conditions, results in decomposition to free cyanide which is toxic (free cyanide« HCN(aq)+CN-(aq)). The initial cyanide form is only stable within the pH range 8 to 14 and zero to –600mV potential (Eh). The rate of decomposition is estimated around 10.2µg l-1 h-1 in salt water. Meeusen et al. (1992) estimated similar rates of decomposition.

Road surface degradation is likely to release various substances: bitumen and aromatic hydrocarbons, tar and emulsifiers, carbonate and metals depending on the road construction techniques and materials used. The following table provides some indication of the concentrations of potentially toxic elements released from road surface abrasion, classified by type of road encountered in urban areas.

Pollutants in Urban Waste Water and Sewage Sludge 37 2. Potentially Toxic Elements

Table 2.23 Emissions from abrasion of urban streets surface material [after Muschack, 1990] Type Total abrasion Individual elements of street (kg ha of road-1 (g ha of road-1annum-1) annum-1) Pb Cr Cu Ni Zn Residential way 1734 177 619 88 2030 285 Residential street 2148 219 767 110 2513 352 Distributor road 3152 322 1125 161 3688 517 Main distributor 4850 495 1731 247 5674 795 road 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, they can be one of the most serious sources of contaminants in urban areas. They can range from minor losses of fuel to major losses from fractured tanker vehicles. The resulting impact on wastewater treatment plants or directly to water receptors is hard to estimate due to the random nature and the unpredictability to both the extent of the spill and its position relative to the precipitation runoff system.

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

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

Parameter Run-off rain water- mean value Water for firefighting- [mg l-1] [mg l-1] Cd 0.0059 0.057 Cr 0.0096 0.053 Cu 0.097 0.203 Fe 3.42 4.0 Pb 0.202 0.439 Zn 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 incidents registered 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 1979 Road incidents reported 125 177 175 136 212 104 Road incidents where pollution 88 63 64 42 58 36 substantiated

Pollutants in Urban Waste Water and Sewage Sludge 38 2. Potentially Toxic Elements

B Roof runoff

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

However, even in the case of normal rooftops (non-metal dominated) there is a concentration 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. The runoff rate of zinc was proven to be considerably lower than its corrosion rate, varying between 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 are strongly 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 be explained by the use of lead in window frames and rooftops (the case of slate roofs in Figure 2.5) and the use of lead sheeting, particularly in the UK. The lead emission factor from lead flashing and roofing in Denmark is estimated at 5.10-4 kg kg-1year-1 (Jøergensen and Willems, 1987]. The presence of cadmium can be explained by the fact that cadmium is a minor 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 respectively in the urban areas [ADEME, 1997]. A study carried out in Nancy in 1996 analysed wet weather runoff and determined the main metal-contributing sources. For zinc, roof runoff was the largest source as it contributed 72%, and road runoff only contributed 12%. For lead, the main 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 represent one storm event.

Relationships between rooftop material and runoff pollutant load can be observed in Figure 2.5 and Table 2.27

Pollutants in Urban Waste Water and Sewage Sludge 39 2. Potentially Toxic Elements

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

Tile 1 Interlocking Zinc Clay Tiles

Tile 2 Flat clay tiles Copper (70%) + zinc sheets Zinc Zinc Sheets Zinc

Slate Slate Zinc

Table 2.28 provides some indicative ranges of potentially toxic element pollution concentrations from roof runoff. For comparison purposes the ranges of concentrations of the same pollutants from other sources in the same study, are also included. The median value 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 runoff Min. 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 indicated median values of 87%), whereas the distribution between particle and dissolved pollution fluctuates for the rest of the potentially toxic elements. Although solids are the main vector of pollution in street and yard runoff, literature agrees that in the case of roof runoff the dissolved phase is of primary importance [Gromaire – Mertz et al., 1999; Förster, 1996; Förster, 1999; Wallinder et al., 2000]. The very high concentrations of dissolved potentially toxic elements in roof runoff make its hazardous and highly dependent on roof materials. Low settling velocities for roof runoff particles in conjunction with high percentage of dissolved pollutants make settling alone an insufficient technique for treatment. Uncontrolled local infiltration practices (infiltration trenches etc) may lead to serious local and present a threat to groundwater quality.

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

Pollutants in Urban Waste Water and Sewage Sludge 40 2. Potentially Toxic Elements

· 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, there seems to be enough residual material from historic lead use to cause high lead concentrations 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 construction practices in the US can be considered different from those in EU, the conclusions of the Davis and Burns (1998) study should be taken into account due to the variability of the structures investigated. In many cases, high lead concentrations were found. Lead concentrations (100 ml over 1600 cm) from washes of 169 different structures followed the order (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 on paint age and condition. Lead levels from washes of older paints were much higher than from 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 were found to liberate more particulate lead than lower intensities. It can be concluded that old surface paints can contribute high masses of lead into a watershed, targeting these structures 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 is systematically in contact with precipitation, much of which ends up in the UWW collection systems. Recent work (PCA, 1992) identified As, Be, Cd, Cr, Hg, Ni, Pb, Sb, Se and Th in concrete in detectable concentrations. Hillier et al. (1999) discuss the importance of the concentrations of these pollutants in leachate from Portland cement. They concluded that leaching of well-cured Portland cement produces undetectable concentrations of toxic metals (such as the ones outlined in 80/778/EEC for water fit for human consumption). In poorly cured (1 day) Portland cement there were detectable concentrations of vanadium (reaching concentrations of 61.7 ppb). However, even this cannot be considered very significant, as the leaching was restricted to the surface of the samples. Furthermore, the water-to-cement ratio has no significant impact on the leaching potential of the cement. This study 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 from other sources. This was shown in Figure 2.3, which shows the initial potentially toxic element loading of precipitation and the subsequent additional loading from roof, pavement and road surfaces. Both traffic density and location of industry have a strong influence on the deposition 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.

Pollutants in Urban Waste Water and Sewage Sludge 41 2. Potentially Toxic Elements

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 Extreme values 1988 [mg l-1] [mg l-1] [mg l-1] Zn 0.05-0.15 0.02-1.9 0.0945 Cu 0.007-0.2 <0.06 0.0355 Pb 0.03-0.11 <0.24 0.0121 Cd 0.001-0.003 <0.13 0.0014 Mn 0.05 <0.1 - Cr 0.002 <0.08 -

Currently, vehicle traffic, steel and glass production, and combustion processes represent the main sources of lead in the atmosphere. Data by ADEME studies have shown that atmospheric fallout contributes 87-536 g ha-1year-1 of lead, and is particularly high in urban areas. 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 cadmium sources are from combustion processes (vehicle traffic, waste combustion in incineration plants). Combustion processes and chlorine and steel production represent the main sources of the atmospheric mercury emissions in the Central Region. Table 2.30 shows the reduction in Cd, Hg and Pb emissions in Austria between 1985 and 1995 as a result of legislative controls.

Table 2.30 Emissions and predicted yearly emissions of potentially toxic elements in Austria, 1985-2010 [UBA, 1999]. PTEs 1985 1990 1995 Prediction Prediction [t y-1] [t y-1] [t y-1] 2005 2010 [t y-1] [t y-1] Cadmium 4.80 3.10 1.80 1.20 1.30 Mercury 4.30 2.70 1.50 1.20 1.10 Lead 320.00 202.00 39.00 24.00 21.00

The presence of impermeable surfaces in urban areas prevents natural processes in the soil zone from removing the pollution. The actual quality of the rain can vary dependent on pollution sources other than urban traffic. In coastal areas the increased concentrations of sodium and chloride in precipitation will increase the susceptibility of vehicles to corrosion. A road drainage study in Sweden noted precipitation pH of between 3.8 and 4.9 [Morrison et al. 1988].

Atmospheric sources contribute significantly to the mass of contaminant available on an impervious urban surface for transport by surface runoff. This deposition may occur during a storm event or as dust fall during dry periods. Cattell and White (1989) [as reported in Ball et al. 2000] in their study in Sydney reported that the geometric mean of the total phosphorous concentration is 29 µg l-1. If the mean annual precipitation for Sydney of approximately 1600 mm 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 0.7kg ha-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 period between storm events. For dry deposition, removal by stormwater runoff is not the sole mechanism responsible for depletion of the contaminant store, which is developed on the catchment surface. As well as removal by storm events, removal also can occur through street 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 contaminant and to move it from the surface of the urban catchment [Ball et al., 2000].

Pollutants in Urban Waste Water and Sewage Sludge 42 2. Potentially Toxic Elements

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 Mean Pb 1.4 7.4 3.7 Zn 0.7 3.9 1.8 Cu 0.3 1.8 0.9 Cr 0.02 0.61 0.14 P 0.23 1.9 0.8

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

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

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

E Sewer cleaning Sewer cleaning is carried out for a number of reasons, but principally to deal with blockages or 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 location and 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 particular locality.

The frequency with which sewers are cleaned will depend on the maintenance strategy in operation. If reactive maintenance is used, problems are dealt with on a corrective basis as they arise (i.e. after failure). This approach will always be required to a certain extent, as problems and emergencies occur from time-to-time in every system. However, its frequency cannot be predicted. In planned maintenance, potential problems are dealt with prior to failure. Unlike reactive maintenance, planned maintenance is proactive and has the objective of reducing the frequency or risk of failure (Butler & Davies, 2000). Planned maintenance differs from routine maintenance (operations at standard intervals, regardless of need), and involves identifying elements that require maintenance and then determining the optimum frequency of attention. Frequency of cleaning will thus vary from place to place but is usually less than once per year even in heavily affected sewers.

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

Pollutants in Urban Waste Water and Sewage Sludge 43 2. Potentially Toxic Elements

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

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

The removal of metals during secondary wastewater treatment is dependent upon the uptake of metals by the microbial biomass and the separation of the biomass during secondary sedimentation. Several mechanisms control metal removal during biological including:

· physical trapping of precipitated metals in the sludge floc; · binding of soluble metal to bacterial extracellular polymers;

In general; the patterns in metal removal from settled sewage by secondary treatment are similar to those recorded for primary sedimentation. However, the general survey of removal efficiencies listed in Table 2.32 suggests that secondary treatment (by the process) is more efficient at removing Cr than the primary stage. Operational experience and metal removals measured by experimental pilot plant systems provide guidance on the overall likely removal and transfers to sludge of potentially toxic elements from raw sewage during conventional primary and secondary wastewater treatment. This shows that approximately 70 – 75 % of the Zn, Cu, Cd, Cr, Hg, Se, As and Mo in raw sewage is removed and transferred to the sludge (Blake, 1979) and concentrations of these elements in the final effluent would be expected to decrease by the same amount compared with the influent to the works. Lead may achieve a removal of 80 %, whereas the smallest overall reductions are obtained for Ni and approximately 40 % of this metal may be transferred to the sludge.

The majority of potentially toxic elements in raw sewage are partitioned during wastewater treatment into the sewage sludge or the treated effluent. However, atmospheric volatilisation of Hg as methylmercury, formed by aerobic methylation biotransformation processes, is also suggested as a possible mechanism contributing to the removal of this element during secondary wastewater treatment by the activated sludge system (Yamada et al., 1959). Whilst some of the Hg removal observed in activated sludge may be attributed to bacterially mediated volatilisation, it is unlikely that this is a major route of Hg loss because of the significant quantities of Hg recovered in surplus activated sludge (Lester, 1981).

Pollutants in Urban Waste Water and Sewage Sludge 44 2. Potentially Toxic Elements

Table 2.32 PTE removals and transfer to sewage sludge during conventional urban wastewater treatment [Lester, 1981] PTE Removal (%) Primary(1) Secondary(2) Primary + Primary + secondary secondary(3) Zn 50 56 78 70 Cu 52 57 79 75 Ni 24 26 44 40 Cd 40 40 64 75 Pb 56 60 70 80 Cr 40 64 78 75 Hg 55 55 80 70 Se 70 As 70 Mo 70 (1)Mean removal (n = 5) from raw sewage and transfer to sludge during primary sedimentation (2)Mean removal (n = 9) in activated sludge from settled sewage, (3)Blake (1979)

Predicting the fate of potentially toxic elements during WWT and sludge treatment Chemical contaminants present in UWW are associated principally with the organic and mineral fractions in sewage. The microbial biomass in secondary treatment is also effective at scavenging potentially toxic elements from the settled sewage transferring metals to the activated sludge. About 60 - 80 % of the wastewater load of potentially toxic elements such as Cu and Pb is transferred into sludge (Table 2.32) and the sludge contains approximately 1000 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 be determined 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 ds Ws = 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 of 70 % (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 raw sewage effluent, based on analytical data for the metal content in sludge and the dry solids in 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 is estimated to be:

Pollutants in Urban Waste Water and Sewage Sludge 45 2. Potentially Toxic Elements

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 is described by Monteith et al. (1993) using a computer-based mathematical model. The mechanistic model, TOXCHEM (Bell et al., 1989), includes both steady state and dynamic models to simulate operation of a conventional activated sludge plant, incorporating grit removal, primary clarification, aeration and secondary clarification unit processes. Biodegradation, surface volatilisation, air stripping, precipitation and sorption removal mechanisms are determined for trace metallic and organic contaminants in the model’s database. A sensitivity analysis feature allows the user to investigate the impact of design and operating conditions and the contaminant’s physical/chemical properties on removal by unit process and by mechanism.

Inputs to the TOXCHEM model include influent wastewater flow rate, influent metals concentrations, dimensions of individual process units, plant operation conditions (e.g. mixed liquor suspended solids concentration), and raw wastewater, primary effluent, and secondary effluent suspended solids concentrations. The model then predicts metal concentrations in primary sludge, return activated sludge, surplus activated sludge, and secondary settler effluent, based on mass balance calculations and removals through precipitation and sorption. The steady-state model accounts for removal by sorption of soluble 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 metal concentrations in the soluble and solid phases:

Ct = Cs + Cx Cx = Cp + KpXCs

Where:

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

The values of Kp and metal solubilities are experimentally defined. Solubility was determined from dosing studies in which metal concentrations were measured in filtered and unfiltered wastewater 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 concentration KP.1 = primary sorption coefficient Xo = influent volatile suspended solids (VSS) concentration.

Pollutants in Urban Waste Water and Sewage Sludge 46 2. Potentially Toxic Elements

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

Where FRm = fractional removal of metal in settler, and FRsolid = fractional removal of solids in settler.

Fractional removal of the solids is input by the user based on site-specific experience. The output 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 contents collected from a full-scale wastewater treatment plant (Table 2.33). The predictions were compared to observed values in the final effluent using linear regression analysis and the r2 statistic. Copper exhibited the highest correlation between predicted and observed effluent concentrations of the elements tested and the r2 value for this metal was 0.96. The model predicted that 70 % of this element would be transferred to the sewage sludge. Zinc also gave relatively good agreement between predicted and observed effluent concentrations and the 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 was 0.41 in this case. The model underestimated the removal of Ni by the wastewater treatment plant and the predicted value was 26 % compared with the observed removal of 37 %. Lead and Cr have low solubilities and in both cases, the TOXCHEM model predicted much larger removals of approximately 70 % for these metals than was observed in practice at the sewage treatment works. However, the model predictions for Pb and Cr were comparable with operational experience at other wastewater treatment plant (Table 2.32). The slopes of the regressions lines fell within the range 0.58 – 1.96 bracketing the ideal slope of 1.0. The model provided good representations of the concentrations of Cu, Ni and Zn in the final effluent, but was less satisfactory for Pb or Cr. Cadmium was not subject to the verification exercise 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 soluble metals, and the most poorly removed, while Pb and Cr were the least soluble. The majority of metals, with the exception of Ni, were maintained in the mixed liquor, and the mass collected in effluent was small in comparison to the total influent mass. At low metal concentrations (0.02 – 0.1 mg l-1), the soluble fraction is controlled by sorption to solids. At larger influent metal concentrations, precipitation is the main process controlling soluble metal fractions. In consequence of this differential fate of metals in WWTS, nickel may need tighter regulation at source to reduce the amount in the effluent.

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

Pollutants in Urban Waste Water and Sewage Sludge 47 2. Potentially Toxic Elements

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

PTE Mean concentration (mg l-1) r2 for Removal (%) Influent Observed Predicted predicted Observed Predicted effluent effluent vs. observed Zn 0.120 0.051 0.057 0.61 58 53 Cu 0.120 0.041 0.034 0.96 66 72 Ni 0.027 0.017 0.020 0.41 37 26 Pb 0.069 0.051 0.022 0.66 26 68 Cr 0.026 0.016 0.007 0.63 39 73

Effect of stabilization processes on metal concentrations in sewage sludge Potentially toxic element concentrations in sewage sludge are influenced significantly by the type of stabilization process operated at a WWTS. For example, the destruction of volatile solids by microbial decomposition of putrescible in sludge by mesophilic anaerobic digestion or aerobic composting processes increase the metal content in direct relation to the extent of volatile solids removal by microbial decomposition.

Unstabilised, co-settled primary + activated sludge typically contains 75 % of volatile matter on a dry solids basis. During anaerobic digestion, 40 % of the volatile matter is destroyed reducing the volatile solids concentration in digested sludge by about 50 % compared to the undigested product. Potentially toxic elements are conserved and are retained in the sludge during the digestion process. As a consequence of the microbial decomposition of organic matter, metal concentrations increase in direct proportion to the loss of solids and are typically 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 or woodchips, to increase porosity and reduce moisture content to values that will support aerobic microbial decomposition processes. Straw is often used for this purpose and may be mixed with sludge cake at 25 % ds at a rate of 5 – 10 % on a fresh weight basis (Smith and Hall, 1991). The addition of bulking agent to sludge reduces the initial metal concentration. However, metal concentrations in the composting material increase proportionally in linear relation to the extent of volatile matter destruction by microbial activity. In trials (Smith and Hall, 1991), for example, the volatile solids concentration of composting 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 was increased above the value in the original sludge (Figure 2.7).

Pollutants in Urban Waste Water and Sewage Sludge 48 2. Potentially Toxic Elements

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

75

70 Volatile Solids 65 5.0% Content (%) 7.0% 60 8.5%

55

50 0 20 40 60 80 100 120 Time (days)

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

This further emphasizes the major and significant improvements in sludge quality, in terms of potentially toxic element content, that have been achieved in the past 20 years through improved industrial practices and the successful and effective implementation of trade effluent controls.

Pollutants in Urban Waste Water and Sewage Sludge 49 2. Potentially Toxic Elements

1300 1200 1100

ds) 1000 -1 900 800 700 600 Cu (mg kg 500 50 55 60 65 70 75

1600 1400

ds) 1200 -1 1000 800 600 400

Zn (mg kg 200 0 50 55 60 65 70 75 Volatile Solids Content (%) Figure 2.7 (a) Cu and (b) Zn concentrations in sewage sludge-straw compost in relation to volatile solids content (Smith and Hall, 1991)

Pollutants in Urban Waste Water and Sewage Sludge 50 2. Potentially Toxic Elements

(a) Zinc (b) Cadmium 2000 2

1800 1.8

ds) 1.6 ds) -1

-1 1600 1.4 1400 1.2 mg kg 1200 1

1000 0.8 0.6

Total Zn (mg kg 800 0.4 600 0.2 Total Cd (Log 400 0 1978 1983 1988 1993 1998 1978 1983 1988 1993 1998 Year Year

Figure 2.8 Reduction in (a) zinc and (b) cadmium concentrations (untransformed and log10 transformed data, respectively) in sewage sludge from Nottingham STW, UK during the period 1978 – 1999

Pollutants in Urban Waste Water and Sewage Sludge 51 2. Potentially Toxic Elements

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

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

Table 2.34 Concentrations of potentially toxic elements found in influents and effluents of WWTS

PTE Country Domestic Urban WWTS Reference Wastewater Runoff ( g l-1) ( g l-1) Influent Effluent ( g l-1) ( g l-1) Cd Austria 30 (<20-60) 70 (<20- Hohenblum et al., 170) 2000 France 6-85 ADEME, 1995 0.2-4.2 Legret,1999 Germany 0.5 80* 0.4 0.1 Raach et al., 1999 <5 Wilderer et al 1997 Greece <1 <1 Greek report, 2000 Italy: < 5 Braguglia, et.al., 2000 Central n.a Northern Sweden* 200-800 100 Adamsson et al 1998 (domestic) Cu Austria 120 (90- 120 (<70- Hohenblum et al., 130) 190) 2000 Germany 150 Wilderer et al 1997 Greece <100 <100 Greek report, 2000 Sweden* 350000- 50000 Adamsson et al 1998 800000 (domestic) Italy: 20-900 Braguglia, et.al., 2000 Central 2-25 Northern France 0.011- Legret,1999 0.146 Cr Austria 7100 3900 Hohenblum et al., (6200-7900) (<900- 2000 5600) Germany 30 Wilderer et al 1997 Italy: < 20 Braguglia, et.al., 2000 Central 0.5-32 Northern Sweden* 4000-14000 3000 Adamsson et al 1998 (domestic)

(Table 2.34 continued) Hg Austria <10 <10 Hohenblum et al., 2000 France 1-8 ADEME, 1995 Germany 0.4 1* 0.6 0.1 Raach et al., 1999 Greece <1-3 <1-9 Greek report, 2000 Italy: < 1 Braguglia, et.al., 2000 Central n.a Northern Pb Austria 30 (<20-60) 70 (<20- Hohenblum et al., 2000 160)

Pollutants in Urban Waste Water and Sewage Sludge 52 2. Potentially Toxic Elements

France 14-188 51-630 ADEME, 1995 Legret,1999 Germany 100 110* 81 31 Raach et al., 1999 100 Wilderer et al 1997 Greece <1-5000 <1-17 Greek report, 2000 Italy: < 20 Braguglia, et.al., 2000 Central 0.3-9 Northern Spain 10-64 Cabrera, et.al,1995 Sweden* 4000-23000 2000 Adamsson et al 1998 (domesti c) Ni Austria <40-170 240 Hohenblum et al., 2000 (<40- 620) Germany 40 Wilderer et al 1997 Italy: < 50 Braguglia, et.al., 2000 Central 0.5-95 Northern Sweden* 3000-10000 3000- Adamsson* et al 1998 5000 (domesti c) Zn Austria 1000 2500 Hohenblum et al., 2000 (<20-3700) (20- 5000) France 104-1544 Legret,1999 Germany 100-1000 Wilderer et al 1997 Greece 450-3200 <20-900 Greek report, 2000 Italy: 100-900 Braguglia, et.al., 2000 Central 12-185 Northern *Sweden 150000- 50000 Adamsson et al 1998 1300000 (domesti c) As Spain 2.2 Navarro, et.al. 1993 Se Spain: Diaz, et.al., 1996 Se (IV) 0.4 Se (VI) 0.3 Fe France 60-999000 ADEME, 1995 *The data from Sweden are from a very small aquaculture wastewater plant, which only serves a few houses and may not be typical of other WWTS.

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

Table 2.35 Organotin compounds in Zürich WWTS

Unit MBT DBT TBT Raw wastewater ng l-1 140 to 560 130 to 1,030 60 to 220 Digested 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 two German 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.

Pollutants in Urban Waste Water and Sewage Sludge 53 2. Potentially Toxic Elements

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

Flow Element Raach et al.,1999 DAF, 1995 [g inhabitant-1 a-1)] [g inhabitant-1 a-1)] Influent Zn 69 75 Cu 19 28 Pb 14 32 Cd 0.1 0.5 Effluent Zn 26 39 Cu 6.3 6.9 Pb 5.1 5.0 Cd 0.03 0.07

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

Table 2.37 Emissions of potentially toxic elements from WWTS in Sweden and Norway 1998

Element Sweden Sweden Norway 1992 1998 1998 (kg.annum-1) Cadmium 325 137 150 Chromium 5420 3308 3000 Copper 14060 15377 15000 Mercury 530* 304 300 Nickel 8165 7603 8000 Lead 2960 1464 1500 Zinc 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 significant reductions in the concentrations of Cd, Cu, Hg, Ni and Zn over the period 1980 to 1997. For lead, the initial improvements were reversed by drought in the period 1990-1997 resulting in a slight, though statistically significant rise in Pb concentrations. However; Pb had fallen prior to this and was still found to have a statistically significant decrease overall in the period 1980-1996 (Table 2.38).

Pollutants in Urban Waste Water and Sewage Sludge 54 2. Potentially Toxic Elements

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

PTE Thames Thames % Elbe Rhine Humber Severn 1986/7 1995 reduction 1983 1984 1984 1988 µg.l-1 µg.l-1 µg.l-1 µg.l-1 µg.l-1 µg.l-1 Cd 0.43 0.32 24.1 0.10 0.30 0.31 0.25 Cu 31.30 10.70 65.8 2.00 5.50 2.17 5.00 Hg 0.24 0.09 63.0 - - - - Ni 17.30 6.30 63.4 3.25 3.20 - 4.50 Pb 16.30 9.90 39.4 1.90 - 1.24 <0.03 Zn 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 50 percent over the years 1986-1995, with lower reductions being experienced for lead and cadmium. This complements evidence for sludge quality that overall PTE emissions have markedly improved.

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

A study carried out in the Rhine region in France [Commission Internationale pour la Protection du Rhin, 1999] examined all pollutant sources entering the river over a period of 10 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 are important contributors, particularly storm runoff for mercury, lead, and copper.

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

Pollutants in Urban Waste Water and Sewage Sludge 55 2. Potentially Toxic Elements

Percentage contribution of for the different sources in the Rhine 100%

90%

80% Diffuse

70%

60%

50% Communal

40%

30% Industrial 20%

10%

0% Hg Cd Cu Zn Pb Cr Ni

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

The communal sources of wastewater into the river Rhine (in France), include wastewater from the UWW collecting system, hence showing the widespread presence of these pollutants (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 total diffuse metal contribution, particularly for cadmium and zinc.

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

Diffuse sources Hg Cd Cr Cu Zn Pb Ni kg.annum-1 Erosion 18 53 5144 3547 10642 4789 5853 Surface runoff 1 10 474 229 1013 67 62 Drainage 7 286 429 2145 28600 2145 1430 Atmospheric deposition 8 75 189 1131 9425 1508 566 Separating system 8 57 475 1900 7600 1520 855 Storm discharge 32 126 630 3780 17640 3780 1890 Untreated wastewater 16 62 310 1860 8680 1860 930 Houses not connected to 2 7 36 216 1008 216 108 UWW collecting systems Point sources Industry 74 242 5548 11190 41100 3120 14300 Communal sources 40 200 1400 12470 30000 2700 3500

Pollutants in Urban Waste Water and Sewage Sludge 56 2. Potentially Toxic Elements

A comparison of potentially toxic element concentrations in sewage sludge applied to farmland in different countries within the EU (Table 2.40) indicates there is some variation apparent in the metal contents of sludges used in agriculture. For example the reported average 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 to sewer in these countries from industrial, domestic and diffuse inputs or the adoption of different sludge treatment practices and regulatory procedures influencing metal content. However, these issues are difficult to reconcile given the policies on preventing industrial discharges of Cd followed in both countries and that both states practice extensive sludge stabilisation treatment. A possible explanation may be related to the statistical characteristics of metal concentration data and how data on metal contents in sludge are reported. For example, CEC (2000a) does not state whether arithmetic averages or weighted averages are given for metal contents. The UK figures are weighted according to works size and provide a conservative estimate of sludge metal content because sludge from large treatment 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 large works (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 of higher industrial inputs of metals to the large works, although it may also be explained by the greater interception of atmospheric deposition of metals by paved areas in urban centres that are served by the largest sewage treatment works. The majority of sludge recycled to land in the UK is produced by 55 large works (160,000 t ds y-1) whereas approximately 840 small works generate sludge (45,000 t ds y-1) for agricultural use. Therefore, an arithmetic mean would indicate a significantly lower concentration was apparent for sludge because all works would have equal weighting. Indeed, the median concentrations recorded for UK sludge are comparable to the values reported for Germany.

Table 2.40 Comparison of potentially toxic element concentrations (mg kg-1) in sewage sludge applied to agricultural land in Germany and the in 1996 (CEC, 2000)

Potentially toxic Germany United Kingdom United Kingdom element (average) (weighted average) (median) Zn 776 792 559 Cu 305 568 373 Ni 24 57 20 Cd 1.45 3.3 1.6 Pb 57 221 99 Cr 40 157 24 Hg 1.35 2.4 1.5

Pollutants in Urban Waste Water and Sewage Sludge 57 2. Potentially Toxic Elements

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/s Austria 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) 1.6 1996 (6) (b) 1.4 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) 9.93 13.5 0.8 15.3 1999 (3) EU 4.0 1992 (9) 2.2 1994-98 USA 38.1 8.95 1988 (13) 25 1992 (2)

Limits Agricultural Sewage Sludge Soils 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-Chromium CHROMIUM 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) 885.8 1996 (6) (b) 43.8 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 1992 (9) 74 1994-98 USA 589 150 1988 (13) 178 1992 (2)

Limits Agricultural Sewage Sludge Soils EU 30 –100 1000 600* (8) (proposed) USEPA 1200 (14)

Pollutants in Urban Waste Water and Sewage Sludge 58 2. Potentially Toxic Elements

Table 2.41c Survey of potentially toxic elements in sewage sludge: values in mg.kg-1 DS-Copper COPPER 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) 302 1996 (6) (b) 103 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 1992 (9) 365 1994-98 USA 639 444 1988 (13) 616 1992 (2)

Limits Agricultural Sewage Soils Sludge 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 Survey Austria 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 2.7 1992 (9) 2.0 1994-98 USA 3.24 2.3 1988 (13) 2.3 1992 (2)

Limits Agricultural Sewage Soils Sludge EU 1-1.5 10 0.5* (4) WHO 5 (15) USEPA 17 (14)

Pollutants in Urban Waste Water and Sewage Sludge 59 2. Potentially Toxic Elements

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 Survey Austria 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) 67 1996 (6) (b) 23.6 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 1992 (9) 33 1994-98 USA 90.6 46.5 1988 (13) 71 1992 (2)

Limits Agricultural Sewage Sludge Soils 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 Survey Austria 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) 283 1996 (6) (b) 140.6 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 204 152 1988 (13) 170 1992 (2)

Limits Agricultural Sewage Sludge Soils EU 50-300 750-70* (4) WHO 150 (5) USEPA 300 (14)

Pollutants in Urban Waste Water and Sewage Sludge 60 2. Potentially Toxic Elements

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 Survey Austria 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) 2752 1996 (6) (b) 1236 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 1992 (9) 817 1994-98 USA 1490 970 1988 (13) 1285 1992 (2)

Limits Agricultural Sewage Sludge Soils 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 Survey Italy 1.1 1.8 1998/99 (10) UK 2.5 1996/97 (7) USA 11.0 6.7 1988 (13) 4.9 1992 (2)

Limits Agricultural Sewage Sludge Soils 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 Survey UK 1.6 1996/97 (7) USA 6.14 4.5 1988 (13) 6.0 1992 (2)

Limits Agricultural Sewage Soils Sludge WHO 140 (5) USEPA 36 (14)

Pollutants in Urban Waste Water and Sewage Sludge 61 2. Potentially Toxic Elements

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 Survey USA 48.2 852 1988 (15)

Limits Agricultural Sewage Sludge Soils 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 >7 respectively (Commission of the European Communities (2000) working document on Sludge: 3rd draft. ENV.E.3.LM, 27 April, Brussels EU 1992 is for B, DK, F, D, EL, IRL, I, L, NL, P, ESP, UK EU 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 agriculture USA is for all sludge: detection limit set at minimum level Greece: (a) values are specific to Athens WWTS; (b) values are average of two rural WWTS.

References

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

Pollutants in Urban Waste Water and Sewage Sludge 62 2. Potentially Toxic Elements

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

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

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

63 3. Organic Pollutants

3. Organic Pollutants: sources, pathways, and fate through urban wastewater 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 may enter UWW. Paxéus (1996a) identified over 137 organic compounds in the influent of the municipal wastewater plants in Stockholm. The physical and chemical properties of some of these organic pollutants are outlined in Appendix B. The main categories of organic pollutants detailed in this report are:

Polycyclic Aromatic Hydrocarbons: Polycyclic aromatic hydrocarbons (PAHs) arise from incomplete combustion or pyrolysis of organic substances such as wood, carbon or mineral oil. 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 road runoff; and also from incomplete combustion processes in urban .

The most frequent anthropogenic sources of PAHs are: house fires, heat and power stations, vehicle traffic, waste incineration and industrial plants (cement works, metal smelting, aluminium production). fires represent natural sources. PAHs concentrate in sewage 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 for wood and paper, flame protective substances, carrier substances for insecticides and in transformers and condensers. The EU1996 PCB Disposal Directive 96/59/EC requires the phasing out of all PCBs by 2010 or by 1999 under international agreement by the States. Existing transformers and other electrical equipment which contain 50-500 mg.kg-1 PCB may be retained in service until the end of their useful life. · The other main source of PCBs in the environment are combustion processes, from waste incineration plants, burning and to other incomplete combustion processes. PCBs are adsorbed by solids and therefore they accumulate in sewage sludge. The highly substituted (high chlorine content) PCBs are the main representatives potentially present in sewage sludge, while they amount to just 35% of the total technical PCBs. Recycling of PCBs in the environment is very important and remediating historical pollution would be necessary 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 in PVC and about 10% in laquers and paints. It is common to use DEHP as antifoaming agent in paper production, as an emulsifier for cosmetics, in perfumes and pesticides, they aid in the production of different synthetic materials such as dielectric in condensers, and substitute for substances such as PCBs and oil. DEHP specific emissions from various 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

64 3. Organic Pollutants

DEHP is found regularly in municipal wastewater and, because of its lipophilic properties, it concentrates 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 as follows [Mahnke, 1997, Horstmann, 1995]: · Chemical reactions or chemical reaction processes: Dioxins arise as unwanted by-products from the production or use of many organo-chlorine compounds, such as chlorine bleaching of cellulose in paper production and chlorine alkali electrolysis. In these cases the formation mechanism can be explained by substitution, condensation or cyclisation reactions. · Combustion processes or thermal processes: Dioxins arise by thermal processes and are released into the atmosphere. The dioxin formation results from a de-novo synthesis. 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 involving chlorine. This explains the ubiquitous dioxins occurrence in the environment. Anthropic production of dioxins has predominated since the introduction of organochlorine compounds in industrial applications (1920). With the improvement of the catalysts in waste incineration plants and other measures for reducing the dioxins emission, the fraction of anthropic dioxins has been declining since 1970. Dioxins can be formed and released into the atmosphere also by natural events, e.g. forest fires. Dioxins can also be generated by the biochemical transformation of precursor compounds (for example during degradation of chlorophenols).

Dioxins speciation in household wastewater and laundry wastewater is similar to those in the sediments 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. Dioxins were also detected in shower water, and in urban run-off from various human activities [Horstmann 1993, 1995]. These results suggest that the importance of household wastewater 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 long chain fatty acids and their methyl and ethyl esters, originating from faeces, soaps and food oils. Being relatively hydrophobic these compounds are attached to particles, the concentration of fatty acids and esters in the unfiltered influent is more than 500 µg/l. Other organic pollutants from domestic origin are the sterols from animal and faeces and indol from faeces. Caffeine is also found from discharges from coffee processing.

65 3. Organic Pollutants

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

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

Solvents both chlorinated and non-chlorinated (, ethers, ketones) are present in a large range of products such as car shampoos and degreasing products, household cleaners and degreasing agents from vehicle maintenance and production. Chlorinated , such as and trichloroethane, are in increasingly wide use: the amounts consumed in France per year are 24,000 and 28,000 tonnes, respectively. The principal sources of diffuse pollution from chlorinated solvents are due to artisanal activities such as metal finishing activities and dry cleaners. Nevertheless, domestic sources from and other agents are not negligible. Pollution by metal cleaning activities is usually considered as diffuse discharges as they are usually from small firms with only few employees. Garages consumed around 15,000 tonnes of solvents in 1988, about 60% of which is lost to the atmosphere and the rest as waste. Of the 6,000 tonnes, of waste some will be discharged into the UWW collecting system [Agences de l'Eau, 1993]. Metal finishing used 50,000 tonnes of solvents in 1991 and their aqueous wastes are discharged into UWW collecting systems, although these are usually in low levels. Dry cleaning consumed around 19,500 tonnes of solvents in 1988 and it has been determined that 0.3x10-3 kg of solvent/100 kg of clothes cleaned ended up in wastewater.

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

Pesticides and are also a common component of the urban wastewaters and they result from road and rail weed treatment, and from gardens, parks and urban woodland areas. They include the triazine group, the phenyl urea group (e.g. chlorotoluron, isoproturon and 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 tonnes of human drugs were prescribed in 1995 in Germany [Ternes, 1998]. Pharmaceuticals in the Urban Environment are discussed in Case Study (d), Section 6.

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

66 3. Organic Pollutants

3.1.1 Domestic and Commercial Sources

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

Table 3.1 Principal sources of organic micropollutants in urban wastewater treatment works [ADEME, 1995] +++ very likely, ++ likely, + less likely present POLLUTANT ORIGIN Domestic Storm Commercial usage runoff effluent

Aliphatic Fuel ++ ++ ++ hydrocarbons Monocyclic aromatic Solvents, phenols + + ++ hydrocarbons PAHs By-products of petrol + + + transformation and insecticides Halogens Solvents, plastics, ++ + ++ chlorination Chlorophenols and Solvents, pesticides + + ++ Chlorobenzenes Chlorinated PAHs PCB, hydraulic fluids (+) + ++ Pesticides + + ++ Phthalate esters Plastifier + + ++ Detergents ++ + ++ Nitrosamines Industrial by-products 0 + ++ (rubber)

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

A. PAHs and PCBs

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

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

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

67 3. Organic Pollutants

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

PAHs WWTS HST GRYAAB SSW µg/l µg/l µg/l Naphthalene, dimethyl 1 0.5

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

68 3. Organic Pollutants

B. DEHP

The Danish Ministry of the Environment and Energy [Danish Report, 1999] have estimated the 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% was discharged to surface and groundwater [Brüggermann, 1988].

The vast majority of phthalate emissions to the environment occur, not during the manufacture, but during the use of the finished products. While in some cases this is a commercial setting (such as vehicle washing, which will be examined subsequently), there are also major sources in the domestic environment. Mattson et al. (1991) mentioned previously regarding domestic sources of PAHs, estimated the household contribution of phthalates and adipates to the Gothenburg sewage works as 70% of the total load (this figure emphasises the ubiquity of compounds and difficulty of control). Two major sources of domestic releases to wastewater (shown in bold in Table 3.3) are floor and wall coverings and 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 air Emission to air Release to during production during use wastewater (t y-1) (t y-1) during use (t y-1) Cars 1000 - 0.1-1 2-10 Floor and Wall 2000 - 0.2 1-5 Coverings Textiles with 5-15 - - 2-13 PVC prints

C Dioxins and furans (PCDD/Fs)

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

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

Table 3.4 shows the dioxin emissions for the years 1994-1998 in Austria. There was little or no change in the dioxin emissions in Austria over this period, but slight reductions, were achieved in some sectors. The main reason for the emission reduction in 1998 is due to the air hold ordinance, which limited dioxin emissions from waste combustion as well as from steam-boiler plants.

69 3. Organic Pollutants

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

1994 1995 1996 1997 1998 Issuer groups Dioxin emissions (tonnes per annum) Small consumer (household, trade, 16,820 18,160 18,400 16,780 16,260 administration) Industry (burning 3,470 3,730 3,880 3,980 3,910 and processes) Industry 8,170 8,900 7,990 8,550 8,380 processes Waste handling 180 180 180 180 180 and landfills Total 28,640 30,970 30,450 29,500 28,740

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

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

Type of sewage Range of Mean value sludge concentrations (pg.g-1 DW as I-TEQ) (pg.g-1 DW as I-TEQ) Fresh [1999] 7 to 160 55 Archived [1979-1987] 29 to 8,300 620

E. Other organic compounds

Adsorbable organo-halogen compounds (AOX) resulting from bleach products and from chlorine 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 average AOX concentration in sewage was reported as 37 µg.l-1.

Sterols were reported in sewage sludge and around discharge wastewater points in Portugal, in Faro, Tavira and Olhao [Mudge et al., 1997, 1998 and 1999]. Concentrations ranged between 0.1 to 27.8 µg.g-1 sterols of dry weight of sludge. Hospital wastewater may contain 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 urban wastewater is vehicle washing, which consists of two distinct phases:

o Actual cleaning, involving the removal of oily dirt, which, on a quantitative basis would be expected to be similar to the type of oily dirt (asphalt and vehicle exhaust particles) which is in road runoff. However, this would also involve the use of degreasing solvents and surfactants which can enter the wastewater treatment process. o Vehicle Treatment, involves the use of protective treatments, often coatings using different types of wax against corrosion, dust and dirt.

70 3. Organic Pollutants

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

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

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

71 3. Organic Pollutants

Table 3.6 Concentration of organic pollutants in car wash effluents in mg l -1 [after Paxeus, 1996]

Conventional LV HV parameters Mean Median Range Mean Median Range Total oil 291 242 10-1750 550 460 65-1200 COD 1263 1180 120-4200 4600 4500 1700- 7500 Aliphatic hydrocarbons C8-C16 29 22 1-139 103.86 76.72 41-220 C17-C30 0.6 0.4 <0.001 1.84 1.87 0.9-3.0 Aromatic hydrocarbons Benzene 0.01 0.01 <0.01-0.2 0.02 0.02 0.02-0.03 Toluene 0.08 0.05 <0.01-0.6 0.10 0.08 0.03-0.2 Naphthalene 0.17 0.13 <0.001- 1.1 0.75 0.3-3 0.7 Biphenyl 0.015 0.005 <0.001- 0.12 0.11 0.04-0.2 0.1 Dibenzofuran 0.001 0.002 <0.001- 0.011 0.011 0.009- 0.03 0.012 Phenathrene 0.005

It is not known if this area is representative of the Scandinavian region as a whole in terms of the car washing input. However, it does seem that car washing is also an important source of pollutants in Norway [SFT, 1998a, 1998b]. In Norway 41 businesses were reported on as sources of hazardous organic pollutants, PAHs, phthalates (DBP, BBP, DEHP), nonylphenols (nonylphenol, nonylphenol mono- and di-ethoxylates). The studies found the highest pollutant loads in the effluents from motor vehicle workshops to urban wastewater came 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 in significant differences in wastewater quality: · Water-based formulations (microemulsions) containing 10-30% hydrocarbons but increased surfactants (10-30%); · Petroleum-based degreasing formulations containing 95-99% of hydrocarbons and 3% surfactants.

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

72 3. Organic Pollutants

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 extractable organics are dominated by petroleum hydrocarbons, which arise from motor oil and tyres from road surfaces. Organic pollutant sources have not received the extent of research attention that potentially toxic element pollution has. For example, in the case of PAHs which are combustion by-products and enter wastewater principally through atmospheric deposition and urban runoff, the sources can be stationary (industrial sources, power and heat generation, residential heating, incineration and open fires) and mobile (petrol and diesel engine automobile) [Sharma et al.,1994]. Different PAH species are associated with each one of these sources.

A. Road and vehicle related pollution

The main sources of road and vehicle related metals pollution have been outlined in Section 2.1.3. 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 Accidents Petrol Tar and bitumen Petrol (PAHs and MTBE) Oil Oil Oil Grease Grease Grease Antifreeze Solvents Solvents Hydraulic fluid PAHs Asphalt PCBs Pesticides and herbicides

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

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

Pollutant mean Test catchments -1 concentrations (mg.l ) Pleidelsheim Obereisesheim Ulm / West PAH 2.61 2.97 2.51

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

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

73 3. Organic Pollutants

Montague, 1994]. The introduction of the catalyst technology for motor vehicles lowered the emissions 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 of tyre failure. A tyre loses about 10 to 20 per cent of its weight in a lifetime. Annually it is estimated 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 an important parameter of organic pollution load in urban runoff. Cary et al. [1989], stated that plasticisers, 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 the greater concentrations than the other seven plasticisers combined. The main sources of plasticisers are traffic grime and dirt, associated with the degradation of plastic components of the vehicles.

B. Roof Runoff

Regarding roof runoff as an interface between atmospheric boundary layer and the runoff receiving system, Förster (1993) investigated the role of roofs as source and sink of organic pollutants. The trace organics analysed included PAH, chlorinated hydrocarbons and nitro phenols. The research indicated that the insecticide HCH was primarily introduced to the roof 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 colder times of the year where fossil fuel heating systems constitutes additional source for this pollutant. The concentration profiles for a number of PAHs are illustrated in Figures 3.1 and 3.2 below.

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

74 3. Organic Pollutants

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 be about ten-fold higher than for tar roofs. There is a difference in the pattern of distribution for PAH concentration at different precipitation flow rates. For tar roofs PAH concentration is highest at the lower and higher precipitation flows and lower at intermediate events, whereas for 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 the roof material itself as well as on the characteristics of the precipitation event.

A number of hydrocarbons are present in urban rainfall runoff, particularly those associated with motor vehicles, such as petrol, fuel oils and lubricants. In an unmodified form these liquids are insoluble in, and lighter than, water. Typically, 70-75% of hydrocarbon oils show a strong attachment to suspended solids [Luker and Montague, 1994]. PAHs have an even greater affinity. In contrast, Methyl-tertiary-butyl-ether (MTBEs) the new additive to unleaded fuel 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 thus adversely affect surface waters. Most hydrocarbons eventually degrade by a combination of microbial and oxidative processes; degradation though is slow, so the increase in oxygen demand in watercourses and wastewater is likely to be marginal and not a principal environmental impact.

C. Urban vegetation control practices

Herbicides and pesticides are used in road maintenance operations to control weeds and pests on the roadsides and verges. The triazine group of herbicides, including atrazine and simazine, has been used extensively for roadside weed clearance and is more soluble and mobile than their organo-chlorine predecessors. Combined levels of atrazine and simazine above 1µg l-1 are not uncommon in watercourses near highways (Ellis, 1991). Collins and Ridgeway (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 integral part of the control of unwanted vegetation by local and municipal authorities, rail and airport operators. The main herbicides used in the UK are of the triazine group, the phenyl urea group (e.g. chlorotoluron, isoproturon and diuron), the phenoxy acid group (e.g. Mecoprop and 2,4-D) and glyphosate (Revitt et al., 1999). Of the phenyl urea compounds, only diuron

75 3. Organic Pollutants has been widely used in the urban environment and in 1989 this accounted for 13% of the total 550 tonnes of active ingredient used in the UK (Department of the Environment, 1991). The comparable use of triazines was 39% but following the introduction of restrictions for the non-agricultural use of these herbicides in 1992, many users converted to 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 by rainfall characteristics, the time interval between herbicide application, the precipitation event and the properties of the herbicide. However, the full range of factors that influence herbicide release from sites of application and the mechanisms governing the transport to, and fate of herbicides in the aquatic environment are not fully understood [Davies et al., 1995; Heather and 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, were consistently above the drinking water limit of 0.1 mg l-1 recommended for simazine and diuron; the mean concentrations of which reached 0.34 and 0.45 µg l-1, respectively. In France [Farrugia et al., 1999], the average application rates for pesticides on the most consuming urban land uses are reported as 900 g ha-1 for roads and streets, 4000 g ha-1 for cemeteries and 500 to 800 g ha-1 for parks and sport yards. Householders may also use large amounts of herbicides and other pesticides but information on the quantities applied is not available in published literature. However, there was considerable variation in the extent of water contamination with herbicides between catchments. Farrugia et al, (1999), reported the average concentration of diurons in water receiving urban runoff was 5 µg l-1, and attributed this entirely to use in urban situations.

It is to be noted that the hydrological characteristics of hard urban surfaces provide the ideal conditions 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 herbicides in urban areas (although in low use), to be an important potential source of contamination of waste water.

D. Wet and dry deposition

The main repository of PCBs, PAHs and PCCD/Fs is soil. Volatilisation from soil, then further atmospheric transport and deposition of PAHs, PCBs and PCDD/Fs is considered to be one of 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 (surface and wastewater, sediment, soil, sewage sludge, compost, plants, street dusts and ambient air) between 1989 and 1998 [Gans, et.al., 1999]. Only 10 % of samples were above the detection limit for PAHs of between 2.6 and 20.3 ng l-1 and these were all taken during winter and , suggesting that PAH originates from the emissions of heating systems during the cold period.

Once released (by the sources mentioned in the previous paragraphs), airborne PAHs are transported by the prevailing meteorology before being removed from the atmosphere through various scavenging mechanisms. As with other airborne pollutants the major mechanisms 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 compounds is better understood than particulate PAH removal [Ligocki et al., 1985]. The extent of in-

76 3. Organic Pollutants cloud or below cloud scavenging, collection efficiency of falling precipitation, solubility and size particles has been examined in the literature [McVeety, 1986 as reported in Sharma et al., 1994].

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

77 3. Organic Pollutants

3.2 INFLUENCE OF VARIOUS TREATMENT PROCESSES ON THE FATE OF ORGANIC POLLUTANTS THROUGH WASTEWATER TREATMENT AND SEWAGE SLUDGE TREATMENT

3.2.1 PARTITIONING OF ORGANIC POLLUTANTS IN WASTEWATER TREATMENT PROCESSES.

The general effect of wastewater treatment processes is to concentrate the organic pollutants in the sewage sludge and the extent of this removal depends on the properties of the organic species. The overall result of this process is to discharge a treated wastewater relatively free of organic and inorganic contaminants and a sewage sludge that contains most of the organic contamination present in the feed wastewater. The main complication of this general study arises from the large number of possible organic species that could be present in the feed 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 to degrade. Indeed, some common organic pollutants such as LAS, are specifically added to detergents because they are aerobically biodegradable. A considerable body of literature exists on this aspect and a variety of oxidants have been proposed. The main aim of this type of work has concentrated on reducing the organic pollutant content in sewage sludge prior to land disposal. Advanced oxidation processes might be used in tertiary treatment especially if the final effluent is to be used for drinking water. However, use of these processes; regardless of the power of the oxidant, cannot be expected, a priori, to degrade all types of organic pollutants within a reasonably short time scale. Indeed, the presence of organics 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 inlet screens used to remove large objects such as wood plastics and paper. However, in reality wastewater conditioning starts in the sewer, in large conurbations the wastewater can have quite a significant residence time in a sewer. However, it is suggested that dilution of sewage with runoff water is likely to have an adverse effect on the efficiencies of the downstream treatment processes (Dorussen et al., 1997).

Primary treatment is installed to enable sedimentation of the feed wastewater. This process is used to settle, retain and concentrate most of the particulate material to the bottom of the tank as primary sludge. The process is affected by temperature and the solids content of the supernatant or primary overflow is significantly higher if the temperature is low, as it is in winter. Though simple, primary sedimentation is a widespread process in Europe, although not practised in all WWTS. In some cases primary sedimentation is not installed and in other plants 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) with air in the presence of aerobic and other micro-organisms, which convert the organic matter to carbon dioxide and water to a variable extent. There are two types of plant commonly used for this process: bio filters and activated sludge. Most WWTS use primary and secondary processes. However some plants may have tertiary treatment which, can involve coagulation, flocculation and rapid gravity filtration.

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

78 3. Organic Pollutants loaded on to the magnetic flocs at a low pH and washed off in a high pH medium (Booker et al., 1996).

There is some concern about the use of iron coagulants, which is of direct relevance to this study. Some iron reagents used in wastewater treatment are made as a by-product of titanium oxide production. The titanium ore contains traces of vanadium and uranium. Two other tertiary methods often cited are activated carbon and membrane filtration. Both however are rather expensive. Activated carbon is a very efficient means of removal of organic pollutants and the technique is widely used in small domestic plants used to polish drinking water. Membrane filtration is also very effective in removing particulate material from water. However, the membranes are expensive and fouling can occur.

In order to estimate organic and inorganic pollutant removal in wastewater treatment processes models are required to simulate them. In such models physical properties of the pollutants 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 their transformation throughout the different treatment methods.

It is clear that the regular screening of priority organic pollutants on a day-to-day basis would be complex and uneconomical. It has been suggested that determination of adsorbable organic 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). These substances are sorbed from the water on charcoal, which is subsequently pyrolysed. The hydroxyhalides produced are sorbed and analysed by titration. Another general test mentioned in the literature (Ono et al., 1996) is the bacterial umu-test, which measures damage caused by organic pollutants on DNA.

3.2.3 Properties of Organic Pollutants Octanol-water partition coefficient and solubility The octanol- water partition coefficient is the ratio of a compound’s concentration in octanol to that in water at equilibrium. Concentration of compound in octanol Kow = Concentration of compound in water

-3 7 Kow is dimensionless and values vary over the range of at least 10 to 10 and are usually expressed logarithmically. Large Kow values are characteristic of large hydrophobic molecules which tend to be associated with solid organic matter while smaller hydrophilic molecules have low Kow values. Octanol-water partition coefficients can be measured directly by using conventional “shake flask” methods (Leo and Hansch, 1971). This experimental approach is restricted to compounds of low-to-medium hydrophobicity, since for compounds with high hydrophobicity, the concentration in the aqueous phase is too low to be measured accurately.

Kow can also be correlated with various environmental parameters, such as solubility. By definition, the partition coefficient expresses the concentration ratio at equilibrium of an organic chemical partitioned between an organic liquid and water. This partitioning is, in essence, equivalent to partitioning the organic chemical between itself and water. One would expect that a correlation would exist between the partition coefficient and solubility. Lyman et al. (1990) presented the following correlation between solubility based on 156 compounds: 1 0.978 log = 1.339log Kow Sw -1 where Sw is the solubility expressed in mol l . This correlation was obtained empirically and the correlation coefficient was found to be 0.874.

79 3. Organic Pollutants

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 partition coefficient. It is defined as the ratio between the concentration of the organic compound on organic carbon (mg.g-1) and its concentration in water (mg.l-1), at equilibrium.

Concentration of compound on organic carbon Koc = Concentration of compound in water The likelihood of the leaching of a compound through soil or adsorption onto soil organic carbon can be assessed from Koc values. Generally, organic compounds with high Koc values will tend to adsorb onto organic carbon whilst compounds with low values have a greater tendency to be leached. Koc values can be estimated from the octanol-water coefficient or water 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, chlorinated hydrocarbons, 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 that 5 their solubilities are very low but the Koc values are very high in the region of 10 indicating that the sorbtion would be very favourable. From the Koc values and the weight fraction of organic carbon species present in the feed (f) an estimate of the removal of organic species can be made. The amount left in the supernatant water as a percentage left (L) is given by: æ 1 ö L = 100 ç ÷ è K oc f ø -3 -5 Thus if f = 10 and Koc = 10 , the percentage left would be 1%. There is limited data available 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 treatment plant.

3.2.4 Modelling

Understanding the processes involved in wastewater treatment is likely to provide a basis for understanding the pathways and partitioning of pollutants in these processes. The way to do this is to develop models of the processes and to simulate plants using computers. An example of such a comprehensive model has been published (Gabaldon et al., 1998). The model 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 in open and closed sewers (Olsen et al., 1998). Another study measures the removal of COD and proteins within a sewer (Raunkjaer et al., 1995) and found that there were quite noticeable losses in a sewer. In another study the sewer pipe was considered to consist of a sediment 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 wastewater plant. In the secondary section aeration for stripping, biodegradation and sorption on to a PAC (Powdered Activated Carbon) were considered. PCB, PCDD/F or PAH were not included but the methodology presented in this paper could be applicable to the study of the fate of these high molecular weight pollutants in secondary treatment. Work has been done on modelling -beds (Shandalor et al., 1997). This predicts the drop of solids loading in the water as it trickles down the bed. On the more specific case of organic

80 3. Organic Pollutants pollutant removal, a detailed paper has been published on the removal of volatile organic contaminants in a wastewater plant (Melcer et al., 1994). However, again no specific mention of PAH or similar organics was made.

3.2.5 Organic Degradation in Wastewater Among the organic pollutants being studied in this report, LAS is somewhat unusual as it is added 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 in WWTS is quite effective some 16% of the feed LAS is taken out in the sewage sludge (Field et al., 1995). In this sorbed form it is more difficult to degrade. Some studies of LAS in river sediments (Tabor et al., 1996) show that this compound is sorbed on to the solids and only slowly biodegradable. Thus there would be an amount of non degraded LAS in the solid residue.

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 were conducted in laboratory rigs and the data reported as half-life analogous to radioactive decay. The mean half-life quoted is given in Table 3.09.

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

Substance Half life in water (years) PCDD 2.6 PCDF 5.0 PCB 9.3

This study seems to indicate that these organics will not be degraded in a WWTS. These half-lives are considerably longer than the residence time in a sewage treatment plant or sewer. 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 that presents evidence that some of these compounds can be degraded in periods of 12-80 hours (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 and nutrients were added. In a practical case where low temperature and few nutrients are present, 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 treatment plant. Research in Greece by Samara et.al. [1995] and Manoli et al. [1999], shows that the lower-molecular mass PAHs are removed effectively in Thessaloniki's WWTS, whereas the higher molecular mass PAHs are resistant to the biological treatment. The heavy molecular mass PAHs are partially removed by adsorption, whereas the lower molecular mass PAHs are removed by volatilisation and/or biodegradation.

Work on oestrogenic compounds, analysing 17b-oestradiol equivalent concentrations, found that the load of oestrogenic activity in the wastewater was reduced by about 90% in the sewage plant. Less than 3% of the oestrogenic activities was found in the sludge (Korner et al 2000).

81 3. Organic Pollutants

3.2.6 Removal of Organics Coagulants such as aluminium and ferric are used in water treatment to remove particulate matter. However, soluble organics may also be removed by coagulation by mechanisms such as specific adsorption to floc particles and co-precipitation (Semmens and Ocanas, 1977). Sridhan and Lee (1972) studied the removal of phenol, citric acid and glycine from waters by co-precipitation with iron. Though these results were reasonable, excessive concentrations of coagulant (300-1500 mg.l-1) were required. Other workers made similar findings. Semmens and Ocanas (1977) examined the removal of dihyroxybenzoic acid (DHBA) and resorcinol from distilled water by coagulation with ferric sulphate. Results indicated that the extent of organic removal increased as coagulant dosage increased. Maximum percentage removals were 35% for DHBA and 8% for resorcinol. Semmens and Ayers (1985) examined the effectiveness of alum and ferric sulphate in removing octanoic acid, salicylic acid, phenol and benzoic acid from water and water samples free of organic matter. These compounds were generally poorly removed by coagulation and in most cases the extent of removal did not depend strongly on coagulant dosage. Removals ranged between 3-20%. Salicylic acid was most efficiently removed and benzoic acid was most poorly removed. Generally, better removal of the organic compounds occurred when natural organics were not present.

The general consensus of the work done to date indicates that the use of coagulants for removing organics is feasible. However it is impractical as the excessive addition of coagulants 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 evidence that they will sorb organic matter by binding with them. Activated carbon is widely used as a means of removing organic compounds from water. The presence of humic acid reduces the rate of organics uptake (Kilduff et al., 1988). The capacity of activated carbon for trichloroethylene (Summers et al., 1989, Wilmanki and Breeman, 1990), trichlorophenol (Najm et al., 1996) and decreased in the presence of humic substances. Other sorption media such as organoclays (Dentel et al., 1998, Zhoa and Vance, 1998) and an organic polymer resin (Frimmel et al., 1999) are not so badly affected by the presence of humic substances. Ying et al., (1988) studied the effects of iron precipitation on the removal of natural organic compounds like tannic acid and humic acid, and toxic organic compounds like chlorendic acid (HET), polychlorobiphenyls (PCBs) and organochlorine pesticides. Freshly formed ferric hydroxide flocs were very effective in removing humic acid and tannic acid and it was found that the presence of humic acid enhanced significantly the removals of PCBs and many of the organochlorine pesticides by ferrous and ferric hydroxide precipitates. Removals were achieved by a combined mechanism of complexation, adsorption and co-precipitation. This evidence suggests that humic substances are capable of sorbing organic material.

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

3.2.7 Conclusions – removal of organics in wastewater The practical issue of the removal of organics in wastewater treatment is not well documented in the literature. Modelling work reviewed here, has shown that the work has concentrated on the removal and degradation of organic matter of biological origin and that synthetic organic pollutants have been largely neglected. Clearly modelling work for pollutants should be promoted.

82 3. Organic Pollutants

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

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

With more rapid analysis techniques in place, there would be the opportunity to make process changes to reduce the amounts of organics present in the final effluents. It is felt that advanced oxidative techniques such as the use of would not be applicable in the present context as the organics have a very small concentration in solution and have a low reactivity. One interesting possibility is in situ treatment in sewers such as adding activated carbon to contaminated . As humic substances are efficient scavengers for organic pollutants, humic acid derived from composting food waste could be added in a tertiary stage to strip organics from the final effluent. It is a matter of policy, to see if such ideas should be promoted further but initial work could start before the rapid analysis methods had been agreed.

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

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

Volatilisation and thermal degradation Many sludge organics are lipophilic compounds that adsorb to the sludge matrix and this mechanism limits the potential losses in the aqueous phase in the final effluent. A proportion of the volatile organics in raw sludge including: benzene, and the dichlorobenzenes may 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 to describe the passive loss of organic compounds to the atmosphere from the surface of open tanks such as . The majority of volatilisation, however, occurs through air stripping in aerated 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 of volatilisation losses of specific organic compounds during sewage treatment can be predicted based on Henry’s constant (Hc) and Kow (Rogers, 1996):

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

83 3. Organic Pollutants

However, more recent studies (Melcer et al., 1992) suggest that the stripping of volatiles may not be as significant as was initially thought and biodegradation during secondary biological wastewater treatment may be the main mechanism of loss of the potentially volatile compound types (Table 3.10). For example, Melcer et al. (1992) reported that biodegradation processes removed ³ 90 % of the dichoromethane, 1,1,1-trichoromethane, trichloroethylene, toluene and xylene from a municipal wastewater. Volatilisation was only a significant mechanism of removal for 1,4-dichlorobenzene (20 %) and tetrachloroethylene (60 %). The fate and behaviour of volatile organic compounds in wastewater treatment plant have been modelled numerically by the TOXCHEM computer-based model that incorporates four removal mechanisms including: volatilisation, stripping, biodegradation and sorption on to solids (Melcer et al., 1992).

Table 3.10 Observed and predicted (TOXCHEM) removals of volatile organic contaminants during wastewater treatment by stripping and biodegradation (Melcer et al., 1992) Compound Air stripping (%) Biotransformation (%) Observed Predicted Observed Predicted 2.6 3.2 92.4 91.9 7.4 7.8 73.6 71.9 1,1,1-Trichloroethane 10.5 6.0 79.7 89.1 Trichloroethylene 10.7 3.1 82.7 91.3 Tetrachloroethylene 58.7 64.2 15.8 0.0 1,4-Dichlorobenzene 19.1 17.2 54.7 54.8 Toluene 1.2 0.4 98.6 98.3 p- 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 can enhance the loss of volatile compounds. Mono- and two-ringed aromatic compounds (benzene, toluene, xylene, naphthalene, dichlorobenzene etc) may be partially lost under these conditions (Wild and Jones, 1989). Other more persistent hydrophobic compounds, eg lesser chlorinated PCBs, and the three-ringed PAHs, may also be susceptible to volatilisation. Thermal drying is being introduced as an enhanced treatment process to produce sanitised for unrestricted use and for improved handling and bulk reduction. This process is potentially the most effective at removing volatile substances from sludge because the solids are exposed to high temperatures (400 oC) and the sludge is dried to >90 % ds. Thermal degradation may also be an important mechanism for the removal of organic contaminants from sewage sludge during heat treatment (Wild and Jones, 1989). Volatile organic compounds in sewage sludge are not regarded as a potential risk to human health or the environment when sludge is used in agriculture (Wilson et al., 1994).

Destruction by sludge stabilisation processes Mesophilic anaerobic digestion is the principal sludge stabilisation process adopted in most European countries, where approximately 50 % of sludge production is treated by this methods. Volatile compounds are generally lost to the atmosphere or transferred to the supernatant during digestion, whereas PAHs and phthalate acid esters are conserved (Bridle and Webber, 1982). Many organic contaminants are biodegraded under anaerobic conditions and this is enhanced by increasing retention time and digestion temperature. Five characteristic behaviour patterns (Figure 3.3) of decay are observed for organic contaminants in anaerobic digestion systems based on net gas (total CH4 + CO2) production (Battersby and Wilson, 1989): · Easily degradable (eg glycol, diethylene glycol, triethylene glycol, sodium stearate, );

84 3. Organic Pollutants

· 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, hexylene glycol, neopentyl glycol, n-undecane, n-hexadecane, 2,4-D, , 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- and 2,5-dinitrophenol, 4-nonylphenol, sodium dodecylbenzene sulfonate, sodium 4- octylbenzene sulphonate, 2,4,5-T, butyltin trichloride, dibutyltin dichloride, tributyltin chloride). Degradation is generally aided by carboxyl and hydroxyl groups, whereas chloro or nitro groups tend to inhibit anaerobic biodegradation and gas production.

Figure 3.3 Typical patterns of net gas production (CH4 + CO2) from organic chemicals incubated 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).

100 1 2

}3 Time 4 Inhibition Net gas production (% theoretical)

-100 5

Biodegradation during anaerobic digestion may virtually eliminate certain organic contaminants from sewage sludge, but in general the destruction achieved is typically in the range of 15 – 35 % (WRc, 1994). Aromatic surfactants including linear alkyl benzene sulphonates (LAS) and 4-nonylphenol polyethoxylates (NPnEO) occur in sludge in large concentrations. These compounds are not fully degraded during sewage treatment and there is significant accumulation in digested sludge. For example, mass balance calculations suggest that approximately 80 % of LAS is biodegraded during the activated sludge process and 15-20 % is transferred to the raw sludge (Brunner et al., 1988). Approximately 20 % of the LAS in raw sludge may be destroyed by mesophilic anaerobic digestion sludge. The compounds, nonylphenol monoethoxylate (NP1EO) and nonylphenol diethoxylate (NP2EO) are formed during sewage treatment from the microbial degradation of NPnEO. These

85 3. Organic Pollutants metabolites are relatively lipophilic and accumulate in the sludge and are also discharged with the treated sewage effluent. One of the most important consequences of anaerobic digestion, however, is the production of nonylphenol (NP), which accumulates in digested sludge. Approximately 50 % of the NPnEO in raw sewage is transformed to NP in digested sewage sludge. The loadings of LAS and NP to soil in sewage sludge used on farmland are significantly larger than for most of the other organic contaminants present in sludge and there is concern about their potential environmental effects. This is particularly the case for NP in sludge due to its potential oestrogenic activity (UKWIR, 1997). However, in the aerobic soil environment, these compounds provide substrates for microbial activity and are rapidly degraded so there is minimal potential risk to the environment or transfer to the human foodchain. For example, LAS has a short half-life in soil in the range 7 – 22 days in temperate 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 of compounds present in sewage because of their extensive use as plasticising agents. The phthalates are also suspected as being potential environmental oestrogens (UKWIR, 1997). Shelton et al. (1984) reported the complete degradation of the lower molecular weight phthalate esters, and of butyl benzyl phthalate, within 7 days in laboratory scale anaerobic digesters operated at 35 oC. Therefore, these phthalate compounds should generally be removed by most municipal anaerobic digesters at the normal mean retention times operated in practice (>12 days). The extent and rate of biodegradation during anaerobic digestion is apparently related to the size of the alkyl side chain and compounds with larger C-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 microbial mineralisation and are generally not removed by conventional anaerobic stabilisation processes. 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 blending dewatered sludge at approximately 25 % ds with a bulking agent, such as straw or wood chips, to increase porosity of the sludge to facilitate microbial activitiy. The biodegradation of relatively persistent organic compounds such has been reported for composted sludge (Wild and Jones, 1989). For example, PAHs may be partially degraded by composting sludge and average removals of 13 % and 50 % have been measured for benzo(a)pyrene and anthracene, respectively, although phenanthrene persisted unchanged in laboratory composting trials (Martens, 1982; Racke and Frink, 1989).

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

86 3. Organic Pollutants

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

For the main list of organic pollutants considered in this report there is little available data of the 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 the Gothenburg wastewater treatment plant. The high cost of testing explains the lack of data on dioxins in urban wastewater.

Most of these compounds were reduced to below the limit of detection during the treatment process. Some of the organic compounds, such as caffeine were reduced from a level of 37µg.l-1 to 4µg.l-1. Some of the phosphorus containing compounds were not reduced during the treatment process (although the influents and effluents were quite low at 1µg.l-1). The overall toxicity of the influent and the effluent were also measured and found to have decreased by approximately 50% during the treatment process.

450 400 350 German limit = 100 ng TEQ kg-1 ds 300 250 200 150 100 ds) -1 50 Dioxin concentration (ng TEQ kg 0 1944 1949 1953 1956 1958 1960 1998 Year

Figure 3.4 Dioxin content of archived samples of sewage sludge form Mogden WWTS, UK It can be seen (Figure 3.4) that there has been a significant reduction in the concentration of dioxins 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 primary sources of these substances. In 1984, McIntyre and Lester (1984) measured median and 99th percentile concentrations for PCBs in sludge (444 samples from UK sewage treatment works) 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 sewage treatment 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 declined markedly in response to the ban on industrial production, use and discharge of these substances. Similar trends are apparent in Germany (Table 3.11). In effect, this means that the chemical composition of sewage sludge is already subject to stringent, albeit indirect, controls that have been effective in minimising industrial sources and inputs of persistent organic contaminants.

87 3. Organic Pollutants

Table 3.11 Mean concentrations of organic contaminants in German sewage sludge in 1988/89 relative to data collected until 1996 (Leschber, 1997)

Contaminant 1988/89 1991/96 Adsorbable organo-halogens mg kg-1 ds 250-350 140-280 Polychlorinated biphenyls(1) mg kg-1 ds <0.1 0.01-0.04 Polycyclic aromatic hydrocarbons(1) mg kg-1 0.25-0.75 0.1-0.6 ds Di(2-ethylhexyl)phthalate mg kg-1 ds 50-130 20-60 Nonylphenol 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)

Compound Country WWTS Reference

Influent Effluent (µg.l-1) (µg.l-1) PAHs Austria: Total PAHs - EPA15 147-625 20-70 Gans et al.,1999 Germany: Hagenmaier et al, Total PAHs 0.79 1986 Benzo(a)pyrene 0.08 Benzo(k)fluoranthene 0.05 Greece: Manoli et al, 1999 Benzo(a)pyrene 0.022 0.005 Fluoroanthene 0.24 0.029 Indeno (1,2,3-cd) pyrene 0.015 0.005 France 0.05-0.44 0.02-0.09 ADEME, 1995 Germany 33 Koch et al, 1989 & Balzer et al 1991 UK 51.8 (5.6 to 30.8 (2.4- Morris et al, 1994 349) 147) DEHP Austria 4.4 0.3 Hohenblum et al., 2000 Germany 122 (7-232) 15 (5.6-184) Faltin, 1985 Anionic Italy 290-4800 - Braguglia et al, Surfactants 2000 Detergents France 1-26 0.1-2.7 ADEME, 1995 LAS Austria 400-3500 11-55 Scharf et al., 1995 Germany 5400 67 Feijtel et al 1995 Greece 129 (35- Kilikidis et al. 1994 325) Italy 4600 43 Feijtel et al 1995 Netherlands 4000 9 Feijtel et al 1995 Spain 9600 140 Feijtel et al 1995 UK 15100 10 Feijtel et al 1995 NPE Austria: Hohenblum et al., Nonylphenol- 2,096,000 363,000 2000 monoethoxylate Nonylphenol- 13,093,000 639,000 diethoxylate Italy: Di Corcia et al. NP 4 1994 NPEO 27 NPEC 145 Sweden 0.5-6.0 Paxéus 1996a Germany 0.02 0.002 Koppe et al, 1993

88 3. Organic Pollutants

Other organic pollutants: France 0.1-0.4 <0.1-0.5 ADEME, 1995 Chlorophenols 300 Chlorinated 1.15 organics 10 Pesticides VOCs Germany Ternes et al, 2000 Iodinated X-Ray 4.3 contrast 3.3 substances: 0.18 iopamidol 0.17 diatrizoate 1.6 iothalamic acid 7.5 iomeprol iopromide Total Phenols Italy 2.5-300 Italian Regional Environmental Protection Agency Dioxins Italy: 0.024-16.9 Italian Regional Environmental Protection Agency

PAHs: wastewater from 8 different sewage treatment influents was investigated in 1996 by the UBA [Gans, et.al., 1999]. Similar PAHs content were determined, except for the influent from a chemical plant, which had an approximately 1,000 times larger concentration. PAHs especially, with a low molecular weight were found in high concentrations. Apart from the higher PAH content of the wastewater from the chemical plant, no significant differences could be detected between municipal and industrial influents. The PAHs content of the effluent 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 and Energy 1996] sets certain cut off values for the maximum concentrations of organic contaminants 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 PAH concentration of the nine selected compounds were all found to have mean concentrations above the concentrations permissible for use on agricultural land in Denmark.

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

Danish Standards 1997 - cut off values 2000 - cut off values mg.kg-1 DS mg.kg-1 DS LAS 2,600 1,300 nonylphenol (including nonylphenol 50 10 ethoxylates) PAHs* 6 3 DEHP 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 sludge concentrations. If 50% of pyrene and phenathrene is from food sources and gives sludge concentrations of > 300mg.kg-1 ds, then this emphasises how difficult these standards are to achieve.

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

89 3. Organic Pollutants

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

It can be seen that government and other institutions are trying to introduce limits for certain pollutants and that concern for wastewater pollution reduction is increasing. Nevertheless, it is 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 the expected increase in sludge production and the reinforcing of the legislation in relation to the concentration limits for potentially toxic elements and organic pollutants, it seems necessary throughout Europe to harmonise analysis techniques and the pollutants targeted in the control of wastewater and sludge quality. Discharge standards to UWW collecting systems for industries and possibly reformulation of certain domestic products should be determined in order to reduce pollution entry into the systems.

Table 3.14a) Survey of organic pollutants in sewage sludge: mg kg-1 DS (a)PAHs PAHs Country Mean Median Min. Max. Year/s of Survey B[A]P Austria 0.30 0.22 0.09 0.67 1994/95 I[1,2,3-cd]p 0.27 0.21 0.07 0.58 (24) S PAHs* Germany 6.4 2.6 15.3 1996 (10) B[a]p (municipal) 0.35 0.1 1.1 1996 (10) Fl.thene 1.2 0.6 2.7 1996 (10) I[1,2,3-cd]p 0.3 0.1 0.8 (10) B[a]p 0.5 0.4 0.1 3.4 1995 (14) B[a]p Denmark 0.15 0.07 <0.01 1.4 (28) Fl.thene 0.3 0.1 <0.01 3.3 I[1,2,3-cd]p 0.67 0.23 <0.01 0.63 Fl.thene Spain 3.4 1.1 6.0 (19) B[a]p France 0.04 11 1994 (6) Fl.thene 0.15 31 1994 (6) B[a]p Greece 0.24 0.24 0.1 0.36 (15) Fl.thene 1.1 1.3 0.38 1.4 I[1,2,3-cd]p 0.11 0.12 0.05 0.15 S PAHs Sweden 1.2-2.2 0.7-1.4 1995/98 Fl.thene 0.01 0.7 (26) (29) S PAHs** UK 27.8 6.0 83.8 1994 (27) B[a]p Fl.thene 2.6 0.1 7.5 1989 (33) 2.3 2.5 1.1 4 1991 (29) S PAHs*** Switzerland 0.50 0.35 0.04 1.83 (29) (municipal) B[a]p Italy <0.05 2000 (34) I[1,2,3-cd]p <0.05 Fl.thene <0.05 S PAHs Poland 72.3 74.4 32.7 114.3 1999 (2) EU B[a]p USA 13.8 4.7 154 1988 Fl.thene 9.95 154 (30/31)

Limits Agricultural Soils Sewage Sludge S 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

90 3. Organic Pollutants

**** 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 of Survey PCB (28,52,101, Austria 0.07 0.05 0.02 0.27 1994/95 138, 153,180) (24) Germany 0.01-0.04 1991-96 0.5 0.05 15 (13) 1985-87 (7) Denmark 0.05 0.03 <0.03 0.2 (28) Spain 0.05 0.93 (20) PCB France 0.03 0.4 1994 (6) (101,118,138) Sweden 0.1 0.1 1995/98 (26) UK 0.34 0.01 21.5 (18) EU USA 1.46 1.48 14.8 1988 (30/31)

Limits Agricultural Soils Sewage Sludge EU 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)DEHP DEHP Country Mean Median Min. Max. Year/s of Survey Austria 23.4 34.4 (11) Bis-(2ethylhexyl)- Germany 20-60 1991-96 phthalate <2.4 320 (13) (7) Denmark 38 25 3.9 170 (28) Bis-(2ethylhexyl)- Sweden 6.7 28 (29) phthalate EU Bis-(2ethylhexyl)- Canada 68.0 11 959 (3) phthalate USA 110 17 891 1988 (30/31) Limits Agricultural Soils Sewage Sludge EU 100 (proposed) (9)

91 3. Organic Pollutants

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

LAS Country Mean Median Min. Max. Year/s of Survey Austria 8107 7579 2199 17955 1994/95(24) Germany 5000 50 16000 1985-87(7) Denmark 2700 530 11 16100 (28) Aerobic Spain 100 500 (1) Anaerobic 12100 17800 (21) Finland 9700 (17) Italy 11500 14000 (4) UK 8700 10400 60 18800 (12) EU Aerobic Anaerobic USA 152 (16) 4680 1680 7000 (22) Limits Agricultural Sewage Sludge Soils 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 of Survey Austria 24 12 69 1994/95(24) Germany 60-120 1988/89(13) 51 3.8 96.3 1996 (10) NP1EO 20 5 80 (7) NP2EO 10 <3 80 (7) Denmark 15 8 0.3 67 (28) Sweden 400 26 1100 1990 (32) 13-27 10-26 1995/98(26) UK 326-638 256 824 (27) EU USA

Limits Agricultural Soils Sewage Sludge EU 50 (proposed) (9)

Table 3.14g) Survey of organic pollutants in sewage sludge: mg kg-1 DS (g)PCDD/F DIOXINS & Country Mean Median Min. Max. Year/s of FURANS (NG Survey TEQ/KG DS) Germany 15-45 1991-96 (13) Spain 55 42 7 160 1994-98 (8) 620 29 8300 1979-87 (8) Sweden 24 23 25 (23) UK 40.2 7.6 192 (29) EU USA 82.7 37.4 0.49 2321 1988 Dioxins 90.4 1820 (30/31)

Limits Agricultural Soils Sewage Sludge

EU 100 (proposed) (9)

92 3. Organic Pollutants

References 1. Berna JL et al, 1989 11. Hohenblum, P, et al 22. Rapaport RA, et al 2. Bodzek, B. et al, 2000. 1990. 1999. 12. Holt MS et al 1992. 23. Rappe et al 1989 3. Bridle, T.R. et al 1983 13. Leschber, R. 1997 24. Scharf, S, et al. 1997 4. Cavelli L, et al 1993 14. Litz. N, et al, 1998. 25. Smith, S.R. 2000 5. Chang, A.G. et al 15. Manoli, E. et al 1999. 26. Statistika 1995. 16. McAvoy DC, et al meddelanden 1998 6. Conseil supérieur 1994. 27. Sweetman 1994 d'hygiène publique de 17. McEvoy & Giger 1986 28. Tørsløv J, et al 1997. France, 1998 18. McIntyre. A,E, et al 29. UKWIR 1995 7. Drescher-Kaden et al 1984 30. USEPA 1992 1992, 19. Moreda, JM, et al 31. USEPA 1999 8. Eljarrat. E, et al 1999. (1998a) 32. Wahlberg,.C, et al 9. European Union, 2000 20. Moreda, JM, et al 1990 10. Hessische (1998b) 33. Wild. S,R, et al 1989. Landesanstalt fur 21. Prats D, et al 1993. 34. Braguglia et al 2000 Umwelt (1991-96).

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 Occurrence in 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

93 4 Health and Environmental Effects of Pollutants in UWW and SS

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 toxic elements, and above certain concentrations, may interfere with or inhibit the actions of cellular . Effects are summarised in Table 4.1, together with the Maximum Concentration Limits (MCL) in drinking water (Source: EPA and WHO).

Potentially toxic elements at high concentrations, are acutely toxic to humans. High concentrations 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 low concentrations over longer time periods. This is chronic exposure and may have more subtle effects.

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

94 4 Health and Environmental Effects of Pollutants in UWW and SS

Health effects of potentially toxic elements

Table 4.1: Summary of the acute and chronic effects of the pollutants PTE Drinking Water Acute Health Effects Chronic Health Effects Carcinogenicity Notes Standards EU WHO US EPA Cu 0.1-3 2 1 Irritation of mouth and throat, Liver and kidney damage, "pink disease", No evidence mg.l-1 mg.l-1 mg.l-1 headaches, dizziness, nausea, cirrhosis diarrhoea, gastric ulcers, jaundice, renal damage, death Zn 0.1-5 5 Stomach and digestion Immune system damage, and interferes No evidence Taken up by plants, mg.l-1 mg.l-1 problems, dehydration, impaired with the body's ability to take in and use phytotoxic muscular coordination other essential elements such as copper and iron Cd 0.005 0.003 0.005 Digestive tract irritation, colitis, Half-life = 10-40 years. Lung, kidney, and Strong evidence in Taken up by plants, mg.l-1 mg.l-1 mg.l-1 , diarrhoea, death. hematopoietic system damage due to build animals, weak phytotoxic up, fragile bones, anaemia, nerve or brain evidence in damage in animals humans Cr 0.05 0.05 0.1 Allergic responses in skin, Damage nose and lungs, increases risks of Evidence in High potential for VI mg.l-1 mg.l-1 mg.l-1 chromium VI irritates nose, non-cancer lung diseases, ulcers, kidney, humans and in lungs, stomach, and intestines, and liver damage. Birth defects and animals aquatic organisms convulsion, death reproductive problems in mice Hg 0.001 0.001 0.002 Nausea, vomiting, diarrhoea, Brain, lung, kidney, and damage to Evidence in mice Binds to dissolved mg.l-1 mg.l-1 mg.l-1 increase in pressure, skin developing foetus, neurological disorders, matter not rashes, eye irritation, renal depression, vertigo, and tremors phytotoxic, High failure potential for bioaccumulation in aquatic organisms Ni 0.05 0.02 0.04 Allergic reactions, lung damage Chronic bronchitis and reduced lung Evidence of lung Phytotoxic, very mg.l-1 mg.l-1 mg.l-1 function, lung disease. Affects blood, liver, and nasal sinus mobile in water kidney, immune system, reproduction and cancers in development in mice and rats humans

95 4 Health and Environmental Effects of Pollutants in UWW and SS

Pb 0.05 0.01 0.015 Anaemia, constipation, colic, Non-specific: damage to nervous system, Evidence in Not usually mg.l-1 mg.l-1 mg.l-1 wrist and foot drop, renal kidneys, and immune system. In children, animals phytotoxic, damage. In children, symptoms can cause decrease mental ability and bioconcentration in are irritability, loss of appetite, reduced growth. In adults, can cause a shellfish vomiting, and constipation. decrease in reaction time and affect memory, miscarriage, premature births, and damage to male reproductive system As 0.05 0.01 0.05 Nausea, vomiting, diarrhoea, Skin keratosis, decrease in the production Evidence in Taken up by plants, mg.l-1 mg.l-1 mg.l-1 damage to tissues including of blood cells, bone marrow suppression, humans, of which are sensitive nerves, stomach, intestines and abnormal heart function, liver/kidney increased risk of to lower skin, damage, impaired nerve function, damage liver, bladder, concentrations than to foetus in animals kidney, and lung animals, cancer, may inhibit bioaccumulation in some DNA repair fish and shellfish, mechanisms persistent in the environment Ag 0.01 - 0.05 Breathing problems, lung and Argyria and may affect brain and kidneys No evidence mg.l-1 mg.l-1 throat irritations, and stomach pains, allergic reactions, necrosis, haemorrhage, and pulmonary oedema Se 0.01 0.01 - Dizziness, fatigue, irritation, Brittle , deformed nails, and loss of Suspected human mg.l-1 mg.l-1 collection of fluid in lungs, feeling and control in arms and legs, : liver bronchitis, rashes, swelling, reproductive effects in rats and monkeys, and lung tumours and birth defects in birds observed in rats and 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 of chemical substances in the environment. Washington DC (1984).

96 4 Health and Environmental Effects of Pollutants in UWW and SS

4.2 Organic pollutants

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

Examinations of the surfactant toxicity revealed that anionic and non-ionic surfactants are non-toxic to man upon oral intake, but can be very toxic for marine life. Table 4.2 shows the toxicity 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 Toxicity LD50 fish (mg/l) LD50 Daphnia (mg/l) LAS 3 - 10 8.9 - 14 NPE 1.5 - 11 4 - 50

Effect of Sludge Treatment Processes on Organic Contaminants

Sewage sludge is treated to reduce its fermentability, nuisance (particularly odour) and vector attraction potential as well as to reduce bulk and improve its physical characteristics to aid handling, dewatering and acceptability for use on agricultural land. These techniques include physical, chemical and microbiological manipulation of the sludge and may involve mechanical dewatering, heating, chemical reactions and microbiological transformation processes that may influence the loss, or potential formation, of organic contaminants. Loss mechanisms 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 others and certain treatment processes can increase the concentration of conservative compounds that are retained as volatile solids are removed during stabilisation action are due to the formation 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 of sludge is typically in the range 60-80%, depending on the extent of stabilisation treatment and volatile solids destruction. The vast majority of the organic content is environmentally benign material that confers part of the agronomic benefit associated with sewage sludge use on farmland by improving the physical and structural characteristics of treated soil. However, certain anthropogenic compounds are potentially toxic and would be detrimental to human health if they were to transfer from sludge-treated soil to the foodchain in toxicologically significant amounts. The range of organic compounds known to exist in sludge is extensive and diverse. For example, Drescher-Kaden et al. (1992) reported that 332 organic substances, with the potential to exert a health or , had been 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 by uptake into the edible parts of crop plants. However, despite the increasing intensity and extent of scientific investigation into the potential environmental consequences of organic

97 4 Health and Environmental Effects of Pollutants in UWW and SS contaminants applied to farmland in sewage sludge, there is no evidence for soil-crop transfer. This is because those compounds exhibiting some solubility and potential for plant uptake are also susceptible to rapid degradation processes in soil or are lost through volatilisation, whereas other more persistent compounds usually have very low solubilities and are strongly adsorbed by the soil matrix in non-bioavailable forms. The principal concern and theoretical mechanism of entry to the human foodchain of organic contaminants in sewage sludge applied to agricultural land is from the surface application of liquid sludge and the intake of organic contaminants by livestock grazing treated pasture and the accumulation of lipophilic compounds in meat fat and milk.

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

New and recently identified organic compounds Uptake of organic contaminants by crop plants from sludge-treated soil has been demonstrated experimentally to be negligible and the human dietary intake of organics from plant 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 of surface applied sludge adhering to grazed pasture and the potential for transfer to milk and meat fat consumed as food. However, it is emphasised that surface application of conventionally treated sludges to grazing land has been phased out in the UK, it is not practised in a number of European countries and may not be permitted in the revised the Sludge Directive. The 3rd Draft of the Working Document on Sludge (CEC 2000) states that deep injection followed by a no grazing interval of six weeks after application is acceptable for sludge treated by conventional processes (eg mesophilic anaerobic digestion). Deep injection burial of sludge below the soil surface improves pasture hygiene and avoids the potential for sludge ingestion by grazing livestock. Advanced treated sludges may be spread on grazing land, but current trends suggest thermal drying of sludge may expand and pasture contamination and animal ingestion is significantly reduced by the application of dried, 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 in sewage sludge applied to farmland. A comprehensive list of compounds that could be present in sludge and maintain their integrity during treatment and following application to agricultural land has been developed by Alcock et al. (1999). The compounds selected included:

Chlorinated paraffins – There are >200 commercial chlorinated paraffin formulations in use as plasticisers in PVC and other plastics, extreme pressure additives, flame retardants, sealants and paints. Total production of short chain paraffins in the EU is approximately 15000 t y-1. Illicit disposal of oils containing paraffins into the wastewater system may be a significant route of entry into the environment. Approximately 50 % of the lubricating oils used by industry in Sweden may be released into the air or WWTS.

Brominated diphenyl ethers (PBDEs) - This group of compounds is used, in descending order of importance, for flame retardation in high impact polystyrene, and in flexible polyurethane 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 and the increased use of plastic material and synthetic fibres. Total annual global consumption is currently 40,000 t. A survey in Sweden (Sellstrom, 1993) indicated relatively high

98 4 Health and Environmental Effects of Pollutants in UWW and SS concentrations of the TeBDE and PeBDE congeners were present in sludge (15 and 19 mg kg-1, respectively).

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

Quintozene (pentachloronitrobenzene) – A small amount of this compound is produced in the EU (21.5 t y-1) and it is registered for use in the UK, Spain, Greece and Cyprus. It has low 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 industrial and consumer products including lubricants, electrical insulators and antifoams. They are hydrophobic and partition onto the sludge solids and concentrations in sludge were reported in 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 they are relatively persistent in soil and can take months to years to fully degrade.

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

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

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

99 4 Health and Environmental Effects of Pollutants in UWW and SS

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

Polyelectrolytes Polyelectrolytes based on polyacrylamide and cationic copolymers are used extensively in sludge treatment to aid dewatering. The polyelectrolyte concentration in mechanically dewatered 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 a rate of 10 % per year (Azzam et al., 1983). Acrylamide is a common monomer associated with 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 in polyelectrolytes used in drinking water treatment has resulted in them being withdrawn from this application in Japan and Sweden, and stringent controls on their use are in place in Germany and France (Letterman and Pero, 1990). These factors give rise to potential environmental concerns about the long-term accumulation of polyelectrolytes in sewage sludge and sludge-treated soil.

The limited available evidence suggests that no environmental problems have been identified from the application of polyelectrolytes to soil. Acrylamide monomer residuals are rapidly biodegraded in the environment and are unlikely to represent a risk to human health in sludge (Gustavsen, 1998). Furthermore, polyelectrolytes are frequently applied directly to soil as soil conditioning agents to maintain soil structural conditions and as a carrier gel for fluid drilling pre-germinated seeds in agricultural and horticultural practice (Fordham and Biggs, 1985). In the fluid drilling technique seedlings of sensitive pregerminated crops emerge directly from the injected polyacrylamide gel. This practical example illustrates that polyelectrolytes do not have a phytotoxic action in soil. An investigation of the potential impacts of polyelectrolytes on nitrification processes in soil (Cartmell et al., 1998) showed no effect on nitrate accumulation at ten times the normal rate of addition of polyelectrolyte to soil in sludge, but nitrification was partially inhibited when the dose was increased to 100x the 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 is likely to significantly reduce their availability and potential toxicity to soil .

Compared to metals, organic pollutants have only recently been identified as having potential adverse human health effects. Most organic pollutants are present in the environment at very low concnetrations. Exposure to these compounds, for example through drinking water, is very low. However, as some of these compounds may bioaccumulate or have effects at low concentrations chronic health effects are starting to be investigated for some of these compounds. Table 4.3 presents acute and chronic health effects of the major organic pollutants covered in this report, and the limits set for these compounds in drinking water.

100 4 Health and Environmental Effects of Pollutants in UWW and SS Table 4.3 Summary of health effects for organic pollutants Compound MCL in Drinking water Acute Health effects Chronic Health effects Carcinogenicity Notes

EU WHO USEPA

DEHP 10 mg.l-1 8 mg.l-1 6 mg.l-1 Usually low acute Damage to liver and Some evidence in Synthetic substance, Emulsified toxicity: mild testes, reproductive rats and mice, adsorbs to soil and Hydro- gastrointestinal effects, birth defects mutagenic at high sediments, carbons disturbances, doses bioconcentrates due to nausea, vertigo lipophilicity PCBs 1.0 mg.l-1 0.5 mg.l-1 Acne-like eruptions Similar to acute Some evidence in Adsorbs to soil and Organo- and pigmentation of poisoning: irritation of rats sediments, persistent, chlorine the skin, hearing and nose, throat, and GI bioconcentrate, and vision problems, tract, changes in liver accumulate in fat spasms function and reproductive systems PAHs 0.2 mg.l-1 0.7mg.l-1 0.2 mg.l-1 Red blood cell Developmental and Some evidence from Adsorbs to soil and Benzo(a Benzo(a) damage, anaemia, reproductive effects, humans and animals sediments, weak )- - suppressed immune birth defects degradation, lipophilic, pyrene Pyrene system bioaccumulate LAS 200 mg.l-1 Gastrointestinal Kidney and liver toxicant Synthetic compound, toxic Surfactants problems to marine life: LC50= 3-10 mg/l, biodegrades easily in aerobic conditions, no accumulation in sediments and biological tissues NPE 200 mg.l-1 1-5 mg.l-1 Allergies, endocrine Has an oestrogenic Lipophilic, mainly affects Surfactants disrupting chemical effect on fish aquatic organisms, partially biodegraded in the environment metabolites are very toxic and persistent PCDD/PCDF 0.00003 Chloracne, rashes, Thymic atrophy, damage Lymphoma, Highly volatile, binds to mg.l-1 discoloration, to the hormonal and melanoma in soil and particles, USEPA excessive body hair, immune systems, humans, can accumulates in fat, website large species spontaneous abortions, increase risk of persistent, birth defects in differences in toxicity, alteration in glucose several types of animals very specific metabolism cancer in man. EU: Water treatment directive WHO: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. Washington DC (1984). USEPA website: http://www.epa.gov/safewater/mcl.html 101 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge

5. REVIEW OF EU AND NATIONAL MEASURES TO REDUCE THE POTENTIALLY TOXIC ELEMENT AND ORGANIC POLLUTANT CONTENT OF WASTEWATER AND SEWAGE SLUDGE.

5.1 Approaches to reducing pollutant content of wastewater and sludge As has been shown so far in this report UWW treatment is not sufficient to deal with all pollution. Most of the pollutants are concentrated in sewage sludge although different contaminants end up to varying degrees in effluent from WWTPs. This effluent ends up in aquatic environments. Concentrating pollutants in sewage sludge, while it may protect receiving 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 varying success, on different levels to reduce UWW pollution closer to at source. These instruments are consistent with an overall strategy of , polluter pays, and reduction at source and include individual regulatory, economic and voluntary and educational instruments. Table 5.1 Summary of Available Instruments Instruments Regulation Economic Voluntary Educational

Direct Licensing of “Downstream” Agreements and Research into Measures disposal to Charges and Management alternative (Applying to UWW taxes on Systems products and direct inputs Licensing of disposal to directed at processes, to UWW) product use UWW Integration of Subsidies for minimization or and Waste Water alternative alternative Minimisation Management technologies, collection Training and UWW into substances or Land Use and collection Spatial Planning services Trading of pollution reduction credits Indirect Product and “Upstream” Agreements, Research into Measures Substance Bans Taxes and Management alternative (applying to Product and Charges on Systems and products and indirect or Substance Use Particular Inputs Codes of processes, diffuse Licensing Subsidies for Practice aimed Management sources) Integration of alternative at alternative and Waste Water management management Minimisation Management and products strategies Training and UWW into Trading of Labelling and Land Use and pollution Specific and Spatial Planning reduction credits General Publicity Measures

The following selected examples demonstrate strategies for the minimisation of waste at source 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

102 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge

· Subsidies to 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 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 to product use and substitution, to specific processes and industries and households, or to specific localities and catchment areas.

5.2 Regulatory Instruments

The use of traditional command and control instruments involves licensing of the use and discharge of material to UWW. This approach is notoriously difficult to monitor and control effectively. In the case of particularly difficult pollutants, targeted product bans can be a more cost effective solution where clear alternatives are available and the production of the material 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 in the electro-technical industry to produce batteries or fluorescent tubes or in dental practices for fillings. Other industries require mercury for the production of instruments such as manometers, barometers and thermometers. It is estimated that the total production of mercury 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 mercury directly when the product breaks, but they also may be exposed after it is disposed and accumulates in UWW and sludge. It can also leach from landfills through the water waste pathway or be directly released in the wastewater system after disposal in sinks and – raising concern for the risk of human exposure and environmental contamination. It is also estimated that before the introduction of the ban on the marketing of clinical mercury thermometers, an average of 15 – 20 million mercury-thermometers were in circulation in France. These are equivalent to 12 tonnes of mercury – about 20% of the yearly amount produced. It is also estimated that an average 1.5 – 5 million of these thermometers were broken every year, releasing between 3 and 10 tonnes of mercury in the environment yearly and 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 nationally and at EU level, as from March 1, 1999. However, the enactment does not prohibit export to non-EU nations. The main objectives of the ban being that of minimising mercury waste and prevent its release into the environment, hence reducing risk to humans.

Costs: Although difficult to evaluate precisely, the implementation of the regulation resulted in two types of costs: 1) Reprocessing Costs associated to the reprocessing mercury thermometers 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

103 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge per unit, whilst a professional device costs around 305 EURO. The cost of an eardrum thermometer (an infra-red device) ranges around 69 EURO. The professional timpanic device costs between 305 and 610 EURO. Assuming that at least 15% of mercury thermometers are collected in the first year, the total cost of the scheme amounts to about 3.81 million EURO. Effectiveness: Although still at its early stage, the ban has been very successful and other nations such as Germany and Sweden are already following similar policy measures. It is estimated 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 taken for other industrial sectors using mercury - and French policy is seemingly moving towards this direction – so as to further minimise the quantity of mercury present in the environment and 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 a material and even product substitution. Nevertheless; evidence suggests that taxes and charges need to be carefully designed and targeted so as to promote continuing waste reduction. The social and economic implications of taxes and charges may make their introduction controversial, and the costs of their administration can make them as complex to administer as command and control regulation.

Wastewater Tax in Germany

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

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

Effectiveness: In terms of offering an overall incentive to reduce discharges, the impact of the measure was found to be limited. A study carried out by Jass (1990) conclude that while the charge had lead to considerable improvements of the purification of sewage and the environmental 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-, 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 environmental charges 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 creates an ongoing incentive to reduce the concentrations (Swedish EPA, 1991) in fertilizers. Since November 1994, the charge rate is SEK 30 (EURO 3.3) per gramme of cadmium if the cadmium content exceeds 5 mg per kg phosphorus (about 2.2 mg Cd per kg P2O5) (Swedish EPA, 1997).

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

104 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge charge) to about 23 mg in 1994/95 and 16 mg in 1995/96. The Board conclude that the Cd charge in combination with the demand (for a low content of Cd in fertiliser) by the agricultural 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 a standard) and voluntary efforts has been successful in reducing the content of cadmium in phosphorus fertilisers. However, it was not possible to estimate the relative importance of the different measures.

Costs: State tax gross revenues in 1996 were around 10 million SEK (1 million EURO), and the administrative costs are estimated to be around 1% of the gross revenues (Drake and Hellstrand, 1998). The tax is administrated by the National Tax Board, together with the tax on nitrogen. Importers and producers report quantities and contents every month. The only control made seems to be ‘tax audits’ (concerning the accounting of the firms involved). In 1999, 25 such audits were performed (corresponding to about one ‘man-year’). The authorities do not measure the actual cadmium content of fertilisers. Some problems with illegal imports by small firms of fertilisers from Poland and the Baltic states (probably of Russian 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 it cheaper or more effective to subsidise alternative production processesses or waste collection strategies than to treat a particular pollutant.

For the businesses using chlorinated solvents, currently only two main options are available to deal with the waste: regeneration of the solvent (10%) and incineration (90%). The problem with regeneration is that the process may affect the added stabilisers, which are required for the stability of the solvent at high temperature and humidity. Hence, it is necessary to treat the regenerated solvent with new stabilisers. Nevertheless, it is usually cheaper 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 organise widespread regeneration of solvents, and these are starting to cover other solvent using sectors. These organisations co-ordinate the collection and treatment of wastes. The Agences de l'Eau subsidise dry cleaners who adhere to these associations in order to help finance the cost of the collection of their wastes.

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

5.4 Voluntary Instruments and Government Industry Cooperation:

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

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

105 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge collection of problematic wastes, voluntary and regulatory or economically inspired waste minimisation 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 and harmful waste from homes. While effective in its own terms it is not a long-term solution to the problem of discharges to UWW. It may be effective to deal with continuing risks of contamination from smaller and diffuse sources, and be used in connection with the adoption of 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 in 1989, 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 total Waste 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 of Boisset-Gaujac (Gard) where 114 kg of chemical products, 169 kg of paint, 12 kg of aerosols, 27 kg of solvents, and 29 kg of products were collected and dispatched separately 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 waste reception centres for handling special wastes rather than organising random collection days. In contrast to targeted waste collection, the adoption of a comprehensive waste minimisation strategy 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 a particular sector in the long-term management of industry specific problems.

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

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

106 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge being chromium III, sulphur, chlorides, solvents and organic wastes. (Vicenza Dept. of Environment, 1997). The tanning process comprises two phases - covering and fixing, where the pelts are sprayed with paints and chemicals for treatment. It is estimated that around 80% of the solvents used become and 1997data from the Dept. of Environment of Vicenza indicate that the solvents used for the finishing process amount to 20.000 tonnes per year, 85% for the covering phase, 15% for the finishing phase. Metoxy propanol is the solvent most used in tanneries and is classified as Class III hazard substance under Italian law, which has 5 classes of hazardous substances for which maximum allowed emissions are set.

Measure: In 1997, the Dept. of Environment of Vicenza introduced new measures to tackle the environmental problems of the spray finishing process. The intervention focused on two main aspects: 1) the introduction of innovative machinery, such as the more efficient low pressure airbrushes or high pressure airless brushes instead of high pressure devices; plus 2) the use of chemicals dissolved in solution. ‘Gruppo Conciario Veneto’ a tannery association comprising 4 companies (La Veneta, Conciaria Adriatica, Sacpa and Veneta Conciaria Valle Agno), with a total of 340 employees, 26.500 m2 of indoor plant and a 1997 turnover of 103 million. EURO, adopted a scheme reducing use of, and ultimately replacing, the solvent metoxy propanol with alternative substances. Other substitutes are currently being tested.

Effectiveness: As a result of the waste minimisation scheme, the use of Metoxy was significantly 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 (Class IV), which has in turn increased in consumption. However, the total quantity of solvent used decreased from 10,650 kg to 5,007 kg per month (- 53%) and therefore less has been generated and released by the tannery (no quantitative data available).

Costs: The reduction in utilised solvents also brought significant economic benefits, with a reduction in monthly consumption equivalent to 64%, from 10,590 EURO in August 1998 to 3,794 EURO in December 1998. These economic savings appear to be the main incentive and persuasive tool to company innovation.

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

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

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

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

107 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge

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

Average Average French Standards concentration concentration (Tours) (France) References Limits Cd 5.3 5.3 20 40 Cr 96 80 1000 2000 Cu 805 334 1000 2000 Hg 7.3 2.7 10 20 Ni 106 39 200 400 Pb 153 133 800 1600 Se 3.5 7.4 100 200 Zn 977 921 3000 6000

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

Effectiveness: As the WWTP receives less contaminated wastewater, the manufacturer pays less 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 in Denmark.

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 of Ringkjøbing. This type of business consumes large amounts of water, energy and chemicals.

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 per annum and the residue consists of excipients such as detergents, phthalates etc. The wastewater from wet treatment is typically heavily dyed and has large contents of salt, detergents, post-treatment agents and other chemicals. The total amount of wastewater within this business makes up 6.6 billion cubic metres per annum. Most of the dye works discharge the wastewater at the municipal WWTPs but there are 4 dye works which have their own water treatment plants with a subsequent discharge to recipient waters. (Genbrug af procesvand fra reaktivfarvning af bomuld, Miljøprojekt nr. 374 1998, p. 15. Miljø- og Energiministeriet)

Measure: The County of Ringkjøbing, in close collaboration local authorities and the Federation 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 party established a score system to govern chemicals. The score system is based on four

108 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge parameters. Parameter A represents the estimated amount of the chemicals discharged into the environment as wastewater. Parameter B is a score on biodegradability, C is a score on bioaccumulation, and D is a score on toxicity. Importantly, a lack of information about a certain product will automatically result in a high score. The score system was implemented in 1992-1993 and the dye works in the County of Ringkjøbing undertook to inform the supervisory authorities about their consumption of chemicals according to the score system.

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

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

109 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge

5.5 Education and Information

Education and Information Campaigns can be general or specific, comprising public education campaigns and specific labelling initiatives. The first example is a general information campaign addressing discharges to water the impact of which is difficult to assess. The second is a consumer information campaign that seeks to promote product substitution. The third is both an incentive and an information measure in the form of a prize for effective environmental performance.

Provision of Consumer Information in France

Measure: Centre d’Information sur l’Eau based in Paris launched a campaign to alert people to the problems caused by disposing of compounds down the sink or the toilet. In order to do so, they distribute a leaflet where they note that the following substances should not be disposed 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 that information about these is available, and that using a waste disposal unit when connected to the UWW collecting system is prohibited, as is discharging rainwater from gutters into the UWW collecting systems when a separate system is in place in the area. The centre had over 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 Danish Society for the Conservation of Nature (DSCN) respectively. These were aimed at phasing out undesirable chemicals in washing powders and have significantly reduced the use of p LAS (Linear Alkyl Benzene Sulphonate), and Nonylphenol. Eco-labelling campaigns were launched by both the SSNC and DSCN, against washing powders containing LAS, whilst at the same time promoting the use of ‘Swan’ or ‘Good Environmental Choice’ labelled washing powders. In the past decade, the eco-label ‘Good Environmental Choice’ has become one of SSNC most important tools for dealing with environmental problems. The initial campaign was launched in 1988 with the publication of one of many green consumer guides, which included information about household chemicals - some of which were found to be replaceable with more suitable alternatives.

Use: LAS is the most important surfactant used in Europe and is found in considerable quantities in washing powders (up to 50%). Some 300000 tonnes of LAS were sold in Europe in 1999, but concerns for its environmental threat have risen since LAS is toxic to aquatic animals and plants. LAS is only degradable in the presence of oxygen, which is not always 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 further degradation (S. Hagenfors, 2000).

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

110 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge

LAS from 6,300 tonnes per year to 260 tonnes per year” (Danish Environmental Agency 2000). The published guide was a success and prompted consumers to demand for more detailed information about domestic products and their potential environmental impact. Coincidentally one the main environmental issues in the region at the time was that of widespread of Scandinavian water bodies which subsequently induced a demand by consumers for environmentally friendly products. In brief, public awareness made the market for eco-labelled products grow in Sweden, so that after the first manufacturer (Unilever) launched its eco-labelled detergent, other multi-nationals had to follow.

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 exception of 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 in Norway and Denmark; the difference in share thought to be due to less public awareness campaigning. Yet, in the last ten years, the Swedish market for eco-labelled products has undergone important modifications. It is estimated that presently 60% of the chemical ingredients in soap, shampoo, detergents have been substituted or removed, the remaining 40% of chemicals falling within the list of substances approved by the ‘Swan’ and ‘Good Environmental Choice’ guidelines. The 38000 tons of chemicals subject to substitution have been replaced by 29000 tons of approved substances, the difference (9000 tons) having been removed. The substitute chemicals are all approved by eco-labels and are more biodegradable in sewage treatment plants and sludge, meaning less environmental risk and less treatment costs. As a result, it is estimated that the use of LAS in Sweden has been reduced by 95% (S. Hagenfors, 2000).

In addition, the development of the eco-label ‘Swan’ has prompted Stockholm Water Co. to initiate a plan for the identification of measures to reduce hazardous household wastes being flushed down the drain and into WWTP. This lead to the establishment of ‘environmental stations’ or collection points and to the launch of a thorough public awareness and information 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 by concentrating the cadmium-containing effluents through evaporation and recovered the metal 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 .

Without upstream control, UWW treatment cannot treat all wastewater pollutants, resulting in effluent and sludge quality problems. A range of measures are likely to be most effective in mitigating 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 local discretion.

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 of waste. Where the costs of treatment are transparent, it is more likely that least cost

111 5. Review of EU and national measures to reduce the potentially toxic elements and organic pollutants content in wastewater and sewage sludge 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 of specific 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 be recognized as the legally appropriate implementation measures in both EU and National Legislation, and the competent authorities should be legally empowered to establish and participate in alternative measures, including substitution and recovery schemes.

112 Section 6. Case Studies

6 CASE STUDIES

(a) Platinum Group Metals in Urban Environment

(b) Sustainable Urban Drainage

(c) Artisanal activities in Vicenza, Northern Italy

(d) Pharmaceuticals in the Urban Environment

(e) Personal Care Products, Fragrances in Urban Waste Water and Sewage

Sludge

(f) Surfactants in Urban Wastewaters and Sewage Sludge

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

(h) Landfill leachate

(i) Potentially Toxic Elements (PTE) transfers to Sewage Sludge

(j) Effect of Chemical Phosphate Removal on PTE Content in Sludge

113 Section 6. Case Studies

(a) Platinum Group Metals in the Urban Environment

Introduction

The platinum group of metals (PGMs), sometimes referred to as the platinum group elements (PGEs), comprise the rare metals platinum (Pt), palladium (Pd), rhodium (Rh), ruthenium (Ru), iridium (Ir) and osmium (Os) and are naturally present in a few parts per billion (mg/kg) in the earth’s crust. The elements are noble chemically unreactive metals, and are found in nature as native alloys, consisting mainly of platinum.

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

Increasing understanding of the environmental damage of vehicle emissions has led to the introduction of stringent emission control standards throughout the western world. Since 1974 all new cars imported or produced in the United States have had catalytic convertors fitted, down hydrocarbon and carbon monoxide emissions. In 1977 they were fitted to a substantial proportion of all cars sold in America, where at the time, this application accounted for 32% of the total Pt usage (Herbert, et al., 1980).

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

Sources

The PGMs are found in nickel, copper and iron sulphide seams (Bradford, 1988). They are currently mined in South Africa, Siberia and Sudbury, Ontario. World mine production of the PGMs, of which 40-50% is platinum, has steadily increased since 1970. This reflects the increasing world-wide use of PGM vehicle catalysts (IPCS, 1991). From 1988-1992 world mine production was essentially constant at around 255 tonnes per year (WMS, 1994). The amount of PGMs present in the earth’s crust down to a depth of 5km, and hence technologically attainable, are still enormous when compared with present requirements, but only a fraction of the pertinent ores is sufficiently rich for commercial exploitation. Of the total of 3x1011 tonnes of PGMs in the earth’s crust, 3x103 tonnes have been mined, and 7x1010 tonnes are minable (Renner and Schmuckler, 1991).

The total worldwide supply of Pt for 1999 and 2000 was 138 tonnes and 153 tonnes respectively for Pd 230 tonnes and 224 tonnes respectively, and for Rh 14.2 tonnes and 20.9 tonnes respectively (Johnson Matthey, 2000)

Uses of platinum group metals.

By far the greatest use of PGMs both in Europe and worldwide is in vehicle catalysts, with additional major uses in the chemical industry, electrical and electronics industries, petroleum industry, the manufacture of jewellery, as a cancer treating in medicine, as alloys in dentistry and in the glass industry.

Demands by application for 1999 and 2000 for PGMs are shown in Table a.1. (Johnson Matthey, 2000)

114 Section 6. Case Studies

TABLE a.1 Platinum Group Metals Demand by Application (Worldwide)

Application kg 1999 (kg) 2000 (kg) PLATINUM Autocatalysts: gross 45600 51000 Autocatalysts: recovery -12000 -13000 Jewellery 79400 83300 Industrial 38400 41400 Investment 5100 -1420 Total Demand (Pt) 159000 146000 PALLADIUM Autocatalysts: gross 166700 146000 Autocatalysts: recovery -5530 -6520 Dental 31500 24700 Electronics 56100 58600 Other 16600 15000 Total Demand (Pd) 265000 238000 RHODIUM Autocatalysts: gross 14400 16000 Autocatalysts: recovery -1870 -2240 Chemical 964 992 Electronics 170 170 Glass 851 1050 Other 312 312 Total Demand (Rh) 14900 16200 RUTHENIUM Chemical 2440 1930 Electrochemical 2040 2270 Electronics 5560 6580 Other 1160 1360 Total Demand (Ru) 11200 12100 IRIDIUM Automotive 964 397 Chemical 198 170 Electrochemical 794 680 Other 936 1450 Total Demand (Ir) 2890 2690

115 Section 6. Case Studies

Trends over time in platinum and palladium uses by application for Europe are shown in Table a.2 and a.3.

TABLE a.2 Platinum demand by application in Europe (kg) Platinum demand 1992 (kg) 1994 (kg) 1996 (kg) 1998 (kg) 2000 (kg) Autocatalyst: gross 16300 17200 14600 15500 17900 Autocatalyst:recovery -142 -284 -567 -851 -1130 Chemical 1420 1420 1700 1700 2410 Electrical 851 709 709 1280 2270 Glass 425 851 1130 709 709 Investment:small 992 1276 142 142 0 Jewellery 2410 2840 3540 4540 5670 Petroleum 567 709 425 425 284 Other 1560 1840 2130 2410 2840 Totals (Pt) 24400 26500 23800 25800 30900

TABLE a.3 Palladium demand by application in Europe (kg) Palladium demand 1992 (kg) 1994 (kg) 1996 (kg) 1998 (kg) 2000 (kg) Autocatalyst: gross 1130 7370 24400 38800 51600 Autocatalyst:recovery 0 0 142 -142 -425 Chemical 2100 1700 1840 1840 2690 Dental 8500 7230 7230 5950 3120 Electronics 5950 7230 8500 7660 7370 Jewellery 992 851 851 1420 1280 Other 425 709 567 709 567 Totals (Pd) 19100 25100 43200 56300 66200

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

Catalytic convertors

A catalytic converter is a unit about the size of a small silencer that fits into the exhaust system of a car. The metal catalyst is supported on a ceramic honeycomb monolith and housed in a stainless steel box similar in shape to that of a conventional silencer. About 1-3g of PGM is contained in some vehicle exhaust catalysts, approximately 50g of PGM per cubic foot of catalyst (Steger, 1994). Due to the commercial sensitivity of these products it is difficult to obtain data on the exact amounts in each of the many different formulations of catalyst. The honeycomb made of cordierite contains 300 to 400 square channels per square inch (6.45cm2), and is coated with an activated high surface area alumina layer called the washcoat (Farruato, 1992) containing small amounts of the precious metals, platinum, palladium and rhodium in varying proportions. The conventional three-way catalysts typically contain 0.08% platinum, 0.04% palladium and 0.005-007% rhodium (Hoffman, 1989).

These metals convert over 90 percent of carbon monoxide (CO), hydrocarbons (HC) and nitrous oxides (NOx) into carbon dioxide (CO2), water (H 2 O) and nitrogen (N2). Platinum is an effective oxidation catalyst for carbon monoxide and hydrocarbons, but it is more sensitive to poisoning than palladium and so can only be used in cars which use unleaded petrol. Palladium is becoming increasingly used instead of platinum due to the higher costs of the latter. The rhodium oxidises the hydrocarbons and reduces the NOx emissions. Base

116 Section 6. Case Studies metals are also incorporated, cerium being the most frequently used; others include calcium, strontium, and iron.

Chemical fingerprinting of ground autocatalyst materials has been undertaken by laser ablation and analysis by ICP-MS for 31 elements (Rauch et al, 2000). Variations in composition were found to occur in agreement with the known fact that variations occur from one manufacturer to another and from one year to another. An association between PGMs and Ce in road sediments was ascribed to the emission of PGMs as abraded washcoat particles onto which PGMs are bound and of which Ce is a major component.

Recycling

Of the total platinum consumption in the United States, approximately thirty per cent is accounted for by vehicle catalysts (IPCS, 1991). The recovery of spent autocatalysts from vehicles at the end of their lives is regarded as important and substantial secondary sources of platinum as well as palladium and rhodium (Torma and Gundiler, 1989). The quantity of spent autocatalysts greatly increased in the United States from 1984 to 1988. These scrapped autocatalysts present an important secondary source of the platinum group metals.

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

Platinum group metals in the environment.

The average concentration of platinum group metals in the lithosphere is estimated to be in the region of 0.001-0.005 mg.kg-1 for Pt, 0.015 mg.kg-1 for Pd, 0.0001 mg.kg-1 for Rh, 0.0001 mg.kg-1 for Ru, 0.005 mg.kg-1 for Os and 0.001 mg.kg-1 for Ir (Greenwood and Earnshaw, 1984).

Although a rapid increase in Pd in sediments from the Palace Moat, Tokyo, Japan was reported by Lee (1983) between 1948 and 1973, it seems unlikely that this was connected with car catalytic convertors since there were few in use in Tokyo by 1973.

Concentrations of Pt and Pd in Boston Harbour have been investigated to evaluate Pt and Pd accumulation and behaviour in urban coastal sediments (Tuit et al, 2000). Increased levels of both metal of approximately 5 times above background concentrations were ascribed to anthropogenic activity with catalytic convertors a major source. It was concluded that anthropogenic enrichments can significantly influence coastal marine inventories of PGMs. The study also indicated that Pt associated with catalytic convertors is much more soluble than expected or alternatively that there is an additional source of dissolved Pt to the harbour. Further study of the biogeochemical behaviour of Pt and Pd was recommended.

Urban pollution with PGMs from catalytic convertors

Emissions of PGMs arise as a result of deterioration of the catalytic convertors, mainly due to thermal or mechanical strain and acid fume components, and are intensified by unfavourable operational conditions (misfiring, excessive heating) which may even destroy the converter (Schäfer and Puchelt, 1998).

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

117 Section 6. Case Studies particle sizes in the nm range, bound to small articles of washcoat material (Schlögl et al., 1987).

Several workers have reported accumulation of Pt, Rh and Pd in road dusts and soils (Zereini et al, 1993; Schäfer et al 1995; Farago et al 1996; Heinrich et al, 1996; Schäfer et al, 1996).

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

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

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

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

Schäfer et al (1999) measured time-dependent changing PGM depositions and contents of dusts and soils at a typical urban location at Karlsruhe in Germany. Daily deposition rates at 2 m distance from the traffic lane were within the range 6-27 ng m2 Pt, 0.8-4 ng m2 Pd. Concentrations of PGMs in the dusts sampled over an 8 monthly period illustrated the steady inputs. Using data for Pt concentration in soil at a site near Pfarzheim and a daily passage of around 15,000 Pt emitting cars per day, the authors calculated for a total number of 11 million converter-equipped vehicles over 2 years, a total emission of at least 3,000 ng of Pt per km along the traffic lane, giving a mean emission rate of 270 ng/km per vehicle. This value significantly exceeds the Pt emission rates of 2-86 ng.kg-1 measured in stationary motor vehicle experiments (König et al, 1992).

A recent estimate of total Pt emission in the vicinity of roads in Germany over the period 1985-2018 was 2,100 kg (using emission factors of 0.65 mg.kg-1 for highways, 0.18 mg.kg-1 for federal and national streets and 0.065 mg.kg-1 for district and city streets) (Helmers and Kummerer, 1999). These different emission rates reflect the increase in Pt load of exhausts with increasing speed of the car.

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

118 Section 6. Case Studies

Richmond, Pt concentrations ranged up to 33 ng.g-1 in road dusts and 8 ng.g-1 in soils. Pt in road dusts was highest at major road intersections (mean 21 mg.g-1) compared with along major roads (13 ng.g-1) and intermediate and minor roads (2ng.g-1). The local background concentration for soils was 1 ng.g-1, similar to that obtained in rural Scotland.

More recently a study in the city of Nottingham, UK, compared Pt and Pd concentrations in garden soils and road dusts taken in 1996 and 1998 and archived samples taken in 1982 (which represented levels before the introduction of catalytic convertors) (Hutchinson, 2001). Significant increases for both Pt and Pd were found in road dusts (see Tables a.4 and a.5 and Figure a.1) with values ranging up to 298 ng.g-1 and 556ng.g-1 respectively for Pt and Pd in 1998 .

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

Year Sample N Range Mean Geomean Median Nottingham 1982 Soil 42 0.27-1.37 0.61 - 0.59 1996 Soil 42 0.19-1.33 0.80 0.75 0.73

1982 Road dust 10 0.46-1.58 0.90 0.80 0.75 1996 Road dusta 8 0.82-6.59 2.78 2.29 2.06 1998 Road dustb 20 7.3-297.8 96.78 69.55 76.72

Table a.5 Summary data for Pd in garden soils (0-5 cm) and road dusts in Nottingham (ng.g-1) a Residential streets with low traffic densities; b Includes major roads with high traffic densities (from Hutchinson, 2001)

Year Sample N Range Mean Geomean Median Nottingham 1982 Soil 42 0.64-0.99 0.05 - 0.04 1996 Soil 42 0.21-1.11 0.18 - 0.10

1982 Road dust 10 0.69-4.92 1.24 - 0.22 1996 Road dusta 8 0.19-1.43 0.75 0.64 0.60 1998 Road dustb 20 5.6-556.3 92.9 40.95 35.84

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

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

119 Section 6. Case Studies

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

Soluble forms of PGMs emitted (in dilute HNO3) were significant for the fresh catalyst but less than 5% of the total amount. A previous study had reported 10% of the total Pt emission to be water soluble for fresh petrol catalysts (König et al, 1992).

At 30,000 km the amount of soluble PGMs released was similar or slightly higher than at 0 km. One possible explanation suggested for the relatively high amount of soluble PGMs related to the relatively high chloride concentration in fresh washcoat (i.e. one example quoted of 3.4 wt %). The authors suggested that the formation of soluble PtCl6, PdCl2 , PdCl4 or RhCl3 could be favoured at the high temperature and humidity that can be reached in the catalyst. The chloride concentrations in aged catalysts are normally very much lower. However, further laboratory tests using spiked solutions showed the instability of these chloro-complexes in the final exhaust fumes solution. It was thus thought possibly that the soluble or labile PGM fraction of the exhaust could be higher than those measured (Palacious et al, 2000).

An important conclusion from this study was that “no clear relation could be observed between the labelled amount of PGMs in the different catalysts studied and the measured amount released through car exhaust fumes. Different catalytic converter manufacturers, different car engines, even if running under the same conditions during the sampling period, and the well-characterized non-uniform behaviour of the catalyst could account for the lack of an observed correlation”.

Emissions of Platinum in effluents from hospitals

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

The authors monitored effluent samples from the University Hospital of Freiberg and two communal hospitals and found total inputs of platinum of around 330g/year from the University Hospital and 12 g/year from the Community Hospital. These equated to a consumption per bed per day of 600 mg Pt for the University Hospital and 85 mg Pt the community hospital. Extrapolation on a national basis, this amounted to an upper limit for the input in Germany of 141 kg Pt per year (c. 645,000 hospital beds (German Statistical Federal Agency, 1994) and a lower limit of 20 kg/year, with an average calculated value of 28.6 kg/year. Comparisons with other sources are shown in Table a.6.

Table a.6 Sources and sinks of platinum in Germany (from Kümmerer and Helmers, 1997) Source Amount Pt (kg/year) Reference Catalytic converters 15 König et al., 1992 emissions Zereini, F., personal communication 1996 Hospital effluents 28.6 Kümmerer and Helmers, 1997 Sewage sludge 100.4 - 400.8 Laschka and Nachtwey German Statistical Federal Agency, 1995

120 Section 6. Case Studies

A broader study based on hospitals in Belgium, Italy, Austria, the Netherlands and Germany aimed to provide reliable data with which to quantify sources of platinum in the environment from hospitals with other sources (Kümmerer et al, 1999). This study was supported by the LIFE95/D/A41/EU/24 Project of the European Community.

It was shown that 70% of the Pt administered in carboplatin and cisplatin is excreted and will therefore end up in hospital effluents. Pt concentrations measured in the total effluent of the different hospitals ranged widely from less than 10 ng.l-1 (the detection limit) in the Belgian and Italian hospitals to cca. 3,500 ng.l-1 for the Austrian and German hospitals. In all cases the influent of the WWTS was below 10 ng.l-1 as a result of dilution within the waste water system.

Annual emissions by hospitals and cars in Germany, Austria and the Netherlands are listed in Tables a.7 and a.8 and compared in Table a.9.

Table a.7 Emission of platinum by hospitals (D=Germany, A=Austria, NL=The Netherlands)

D 1994 D 1996 A 1996 NL 1996 Total hospital beds (approx) 645000 645000 77500 60000 Maximum Medical Performance 45000 45000 6500 N/A Pt per bed and year (mg) - maximum medical service spectrum 154.0 130.4 58.7 22.3 - medium medical service spectrum 14.0 14.0 N/A N/A Pt emissions by hospitals - maximum medical service spectrum 6.9 5.8 0.38 1.3 - medium medical service spectrum 8.4 8.4 N/A N/A Total emissions by hospitals (kg) 15.3 14.2 N/A N/A All hospitals as maximum medical service 99.3 84.1 4.6 1.3 spectrum

121 Section 6. Case Studies

Table a.8 Emissions of platinum by cars D 1994 D 1996 A 1996 NL 1996 Number of cars 32 000 000 32 000 000 3 593 588 5 740 489 With catalytic converter 12 800 000 19 200 000 1 607 699 3 307 300

% with catalytic converter 40.0 60.0 44.7 57.6 Kilometres/car 15 000 15 000 14 374 13 538 Total Kilometers (cat only) 1.92 x 1011 2.88 x 1011 2.311 x 1010 4.491 x 1010 Emission ( g km-1) 0.65 0.65 0.5 0.5 Total emission by cars (kg) 124.80 187.20 11.55 22.46

Table a.9 Platinum emissions comparison: hospitals vs. cars D 1994 D 1996 A 1996 NL 1996 Maximum medical service spectrum 5.6 3.1 3.3 6.0 Medium medical service spectrum 6.7 4.5 N/A N/A

Total 12.3 7.6 3.3 6.0 All hospitals calculated as maximum 79.6 44.9 39.4 6.0 medical service spectrum

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

Emissions from other Sources Kümmerer et al (1999) further concluded that “emissions by traffic and hospitals cannot explain the whole amount found in sewage and other sources emitting platinum directly into sewage have to be considered like glass and electronics industries or jewellery manufacturing. For the catalytic ammonia oxidation 92 kg platinum are reported to be lost from the catalyst” (Beck et al., 1995). If all of this is emitted into the atmosphere and washed off from roads and other paved areas in urban areas, which make up 11% of the total area of Germany (Losch, 1997), 10 kg from this source would be the input into sewage. If there are local industries like jewellery and electronic industries (Lottermoser, 1994) which use platinum to a certain extent they might be the most important local contributor to the platinum content of a certain municipal sewage and sewage sludge. Thus, unspecified input directly from industrial processes into sewage must be taken into account. These possible sources include jewellery manufacture, dental laboratories, electronic industries, glass manufacturing, production of platinum-containing drugs and industrial catalysts.

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

A study of PGMs in sewage sludge incineration ashes from the municipal WWTS at Karlsruhe, Germany, showed Pd concentrations to have increased from 64 to 138 mg.kg-1 from 1993 to 1997, with Rh increasing from 4.8 to 6.3 mg.kg-1; Pd varied from 300 to 450 mg.kg-1 with no significant trend although these concentrations were 10-fold higher than in 1972 (Schäfer et al, 1999). The authors drew attention to the Pt/Rh ratio in the sludge of c. 20:1 which differs greatly from that of 6:1 typically found in environmental samples influenced by traffic emissions. They then estimated that, as more than 90% Rh is used for the production of catalytic convertors, and as Pt and Rh are emitted in a ratio of 6:1, that the contribution of traffic to the Pt concentration in sludge is only c. 30%, a result similar to that found in Munich by Laschka et al (1996). They thus suggested that the greater part of Pt in sewage sludge must come from sources other than catalytic convertors, such as hospital

122 Section 6. Case Studies and medical effluents or industrial emissions. They drew attention to the fact that “in with a jewellery industry, noble metal concentrations in sewage sludge far exceeded normal values even before the introduction of catalytic convertors” (Lottermoser, 1994).

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

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

The study of Laschka and Nachtwey (1997) concluded that “in a large industrial centre such as Munich, automobile traffic is not the dominant source of Pt in municipal sewage”.

A recent review article by Helmers and Kümmerer (1999) has attempted to quantify the sources, pathways and sinks of Pt in the environment. The authors noted that there was as yet not enough data to reliably investigate Pd and Rh fluxes, noting the lack of good quality assurance for Pd analysis and the paucicity of environmental data on Rh. An examination of archived sewage sludge ash from Stuttgart, Germany showed a continuous increase in Pt concentrations since 1984. With an estimated 5 x 1010 kg of sewage sludges for Germany in the early 1990’s and c. 250 mg.kg-1 Pt in sewage sludge, this amounts to some 12,500 kg of Pt, some 2 orders of magnitude higher than the Pt flux emitted by traffic. Much smaller Pt concentrations (mean 35 mg kg-1) have been found in smaller German purification plants.

Assuming that 50% Pt emitted by cars is received by sewage systems and taking into account the amounts of Pt in effluents from hospitals being completely received by sewage systems, the authors calculate that for Germany Pt received in the influents of WWTS from both these major sources amounted to 42.9 kg in 1994 and 56.4 kg in 1996. They thus considered an additional input of around 10 kg Pt per year from industrial sources.

If around 70% of the Pt influent is removed within the WWTS into sludge, the remaining 30% is emitted into freshwater (see Table 10). In Germany 30% of the sewage sludge is used on agricultural land and 70% disposed of as sludge or incinerated sludge ash. Losses to the atmosphere from incineration are not yet known.

Table 10. Partition of anthropogenic Pt fluxes (in kg) within German WWTS Year Received by Remaining in Disposal with Deposited Released into the WWTS the sewage sludges or agriculturally freshwater sludges ashes with sludges 1994 42.9 30.9 21.6 9.3 12 1996 56.4 40.6 28.4 12.2 15.8

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Helmers and Kümmerer (1999) consider the possibility of extrapolating these results to other European countries, taking into account traffic densities, catalytic convertor policies etc., with the qualification that “since there is no highway speed limit in Germany, highway Pt emissions of other countries may be halved in comparison with the German situation” (Helmers, 1997).

Solubility and bioavailability of PGMs in the environment

Current scientific opinion would seem to agree that PGMs emitted as autocatalyst particles remain bound to these and have limited mobility in the road and soil environment (Rauch et al, 2000). Experimental studies under laboratory conditions, in which ground catalysts have been added to soils under varying conditions of pH, chloride and sulphur concentrations have indicated that post-deposition processes in soils and waters are of minor importance and that “the risk of a health endangering contamination of the environment, and especially groundwater, at present seems negligible, as the PGM species behave relatively inertly” (Zereini et al, 1997). However transformation of PGMs into more mobile forms has not been ruled out and indeed Rauch et al (2000) suggest that this may occur “in the roadside environment, during transport through the stormwater system or in the urban river”.

In the absence of detailed study, it would seem impossible at this stage to apportion soluble PGM species in the influents and effluents of WWTS’s to specific PGM sources from traffic, hospital or industry, or to transformation/mobilisation of Pt and other PGMs in the environment and/or waste water system, or indeed in the processing plant.

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Conclusions

Platinum group metals are present in the influents of WWTS as a result of

· exhaust emissions from motor vehicles using catalytic convertors (both petrol and diesel) and subsequent runoff from road surfaces and roadside soils; · emissions in effluents from hospitals using the anti-cancer drugs cirplatin and carboplatin, · industrial uses including jewellery manufacture, electronics and glass manufacture.

Several studies over the past decade have shown a steady increase in the use of PGMs. Reliable quantitative information has shown that in general by far the greatest input of Pt and Pd into the environment and into WWTS is from vehicle exhaust catalysts, with hospital effluents accounting for some 6 to 12 per cent Pt. In large industrial centres, such as Munich, inputs from other sources (presumed industrial) may exceed those from catalytic convertors and hospitals. Quantitative knowledge of these sources is not currently available. The solubility of PGMs entering the environment and the influents of WWTS is thought to be low, though reactions within the soil/dust and wastewater environments need further study. Interactions with chloride ions and humic substances may well increase solubility and thus bioavailability.

Around 70 per cent of Pt in the influents to WWTS is removed with treatment into sludge, which may then be applied to agricultural land or incinerated. Where land application is practical, studies into uptake into pasture and foodcrops are recommended. The 30 per cent of Pt emitted into freshwater systems will potentially increase Pt levels in drinking water supplies.

At present there is no evidence of health risks arising from increasing levels of PGMs in the roadside environment, in sewage sludge or in drinking water. However, as levels of use continue to rise, it would seem prudent to focus research into factors influencing their solubility and bioavailability, their uptake and input into food crops and drinking water and into multiple exposure routes into the population.

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(b) Case Study- Sustainable Urban Drainage

Summary

Urban runoff source control practices have been the centre of an ongoing discussion involving maintenance and quality issues. This review will provide a brief overview of available techniques and structures and summarise their design characteristics. A discussion on performance will focus on water quality but comments on maintenance will also be included to allow the reader to form an overall opinion. A number of source control application case studies in Europe will be discussed from the point of view of performance.

Introduction

Urban Runoff has traditionally been treated as a water quantity problem and the usual approach to solving it has been a system of buried pipes designed to convey water downstream as soon as possible (CIRIA, 1999). Several problems in this traditional approach have been identified including possible flooding in downstream areas due to alteration of natural flow patterns, water quality issues that are not dealt with within the pipe system and largely ignored amenity aspects (such as water , landscaping potential and provision of varied wild habitat). These considerations have led to an effort of rethinking surface water drainage methods within the following framework:

· Deal with runoff as close to the source as possible · Manage potential pollution at source · Protect from pollution · Increase amenity value

This framework and the practices and drainage systems that were developed from it, are collectively referred to in the UK as “sustainable urban drainage systems” (SUDS) (CIRIA, 1999) or more generally “source control” (Urbonas and Stahre, 1993) or “best management practices” (BMPs)1 (Jefferies et al., 1999). They essentially confirm with the emergence of Agenda 21 as a local action-planning basis for strategic and integrated approaches “to halt and reverse the effects of environmental degradation and to promote sound environmental development” (United Nations, 1992). Source Control includes structures such as:

· Dry Detention Basins · Infiltration Devices · Oil and Devices · Sand Filters · Vegetative Practices · Filter Strips · Grassed Swales · Wetlands, Constructed · Wetlands, Natural and Restored · Wet Retention Ponds

1 The fact that this Report adopts the widely used terms SUDS and BMPs to refer to source control and distributed storage practices does not imply that it necessarily considers them either “sustainable” or “best”. The positive and negative aspects of these practices will be discussed in the following paragraphs.

126 Section 6. Case Studies

Basic design characteristics and principles of use of the most widely used of these systems will be presented in the following paragraph and are summarised in Table 1.

Source Control Systems

The techniques presented will be grouped in four categories according to the CIRIA recommendations: (a) filters and swales, (b) permeable surfaces (c) infiltration devices and (d) ponds (CIRIA, 1996; CIRIA, 1999). The overall structure of an urban catchment with source control can be seen in the schematic in Figure b.1.

Figure b.1. Urban Runoff and Catchment (after CIRIA, 1996)

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Filters and Swales: These are vegetated landscape features with smooth surfaces and downhill gradient. Swales are long shallow channels while filters are gently sloping areas of ground. They mimic natural drainage patterns slowing and filtering the flow and are used for the drainage of small residential areas and roads. The flow depth should be smaller than the height of the grass to ensure filtration. Operational practices include regular mowing and clearing litter. Special care should be taken not to allow the swale to erode after heavy storms. Grass swales have been used extensively in North America, but have only recently appeared in Europe. Information on treatment performance comes mainly from the US (as summarised in Ellis, 1991) and indicates removal potential for solids, potentially toxic elements and hydrocarbons). In the UK however the quality improvement potential of Figure b.2. Grass Swale (after CIRIA 1996) the swales is ignored. A typical swale structure can be seen in Figure b.2.

Permeable surfaces: These include porous pavements, gravelled areas, grass areas and other types of continuous surfaces with an inherent system of voids. The water passes through the surface to the permeable fill, allowing for storage, transport and infiltration of water. The actual amount of water stored is dependent on the voids ratio, the plan area and the structure’s depth. It acts as a trap for sediment thus removing a large number of pollutants from the runoff, but keeps them within the particular site. The principal mechanism for pollutant retention is thought to be adsorption onto materials within the pavement construction (Pratt, 1989). Maintenance should ensure that the voids are not filled by sand and silt and such an operation may prove costly, as the surface structure can deteriorate under external pressure. The US, France, Holland, Austria and Sweden have used porous surfaces for Figure b.3. Porous Pavement (after CIRIA both traffic and pavement areas (Diniz, 1976; 1994) Hogland, 1990). A typical structure can be seen in Figure b.3.

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Infiltration devices: Soakways and infiltration trenches are below ground and are filled with a coarse material. They drain water coming in the infiltration device from a pipe or a swale directly to the surrounding soil. Their operation is based on increasing the natural capacity of the soil for infiltration but effectiveness is ultimately limited by soil permeability. The volume of storage therefore is dependent on soil infiltration potential. Physical filtration can remove solids, while biochemical reactions caused by microorganisms growing on the fill or the soil can degrade hydrocarbons. The level of treatment depends on the size of the media and the length of the flow path (CIRIA, 1999). Extensive use of soakways in Sweden and the US as well as in the UK generally provide positive feedback on maintenance and operation (Pratt, 1989; CIRIA, 1996). Areas that are drained through infiltration structures of Figure b.4. Soakway (after CIRIA, 1996) different types include car parks, roads, roofs pavements and pedestrian . Pollution levels in these types of urban runoff can be however significant and there is therefore serious risk of introducing the pollutants to the groundwater. Additionally the introduction of water to the soil may cause geotechnical problems. Figure b.4 describes a typical soakway.

(d) Basins and Ponds: These are areas of storage of surface runoff that are free from water under dry weather conditions. Structures can be mixed with a permanently wet area for or treatment of runoff and an area that is usually dry to allow for attenuation. The ponds are normally situated near the end of the system due to detention and land price constraints (Makropoulos et al., 1999). Flow detention would lead to settlement of the particles and associated pollution loads. Additionally some bacterial die-off and soluble particle removal could be expected (CIRIA, 1994). Annual clearance of the aquatic vegetation and silt-removal every five to ten years should be thought of as an average operational practice. Figures b.5 and b.6 give an idea of on and off stream detention and the pond-wetland principle respectively. Figure b.5. Typical on and off stream storage ponds (after CIRIA, 1994)

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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. Research has been focusing on mathematical modelling of the system’s quality performance and actual data is generally not available in Europe. Scotland is a notable exception. BMPs have been promoted for the past five years in response to the need to combat pollution from diffuse sources in urban areas. To meet this need, a programme of investigations is being undertaken into the performance of BMPs, which have been built in Scotland. An initial awareness survey by Abertay University and SEPA indicated high levels of apparent knowledge of BMPs, but subsequent investigations showed that in many instances knowledge was very superficial and often inadequate. Jefferies et al. (1999) discuss experimental findings and theoretical considerations of that investigation and show that, in most systems, pollutants will form sludge. This in turn must be disposed of, and indeed good household practices may be the only truly sustainable drainage practice. Pollutants removed from runoff in a system such as a pond may accumulate in sediments and biota. Potentially toxic elements and trace organics in rainfall runoff are to some extent associated with soil particulates, as discussed earlier in this report, and will thus tend to be removed by sedimentation. The soluble fraction of pollutants will also to some extent precipitate following changes in pH, oxidation-reduction potential or temperature (Kiely, 1997). The activity of the pollutants however is not ended with their concentration in the sediments. Polluted sediment may be resuspended or pollutants may be released during high stream flows (Pitt, 1995). The quality of groundwater may also be affected by exfiltration of contaminants from BMP systems. Studies in the US have shown that, when disposed in soakways, organophosphates have appeared in watercourses 400 metres away only two hours after disposal (ENDS, 1993). The entire range of toxic pollutants identified as possible input to urban rainfall runoff may leak this way to the groundwater. A new problem has appeared in the form of methyl butyl tertiary ether (additive to unleaded petrol), which is ten times more soluble in water than other constituents in petrol and thus would tend to spread readily in groundwater (Kiely, 1997). When soil is used as a filtration medium in source control systems (as in infiltration trenches and even grass swales), it must be regularly checked as the adsorbed pollutants may be remobilised under various conditions. Furthermore, possible degradation of pollutants inside the systems may give rise to hazardous by-products which may be more soluble or toxic than the original forms (Hallberg, 1989). Biotransformation of TCE for example results in hazardous products such as vinyl chloride, which is a confirmed human carcinogen (Burmaster, 1982). Table b.1 summarises the main functions and water quality attributes of source control.

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Table b.1. Functions and water quality attributes of different source control structures (after CIRIA, 1994)

METHOD PRIMARY SECONDARY WATER QUALITY ATTRIBUTES FUNCTION Infiltration Collection and Sediment and Can remove pollutants associated pavements disposal of pollutant removal with sediments and dissolved surface water pollutants but may be lead to increase in nutrient levels Swales Conveyance of Storage sediment and Can remove suspended and possibly surface waters pollutant removal; dissolved pollutants but may be a risk disposal to groundwater quality if not sealed Infiltration Disposal of Storage; sediment and Can remove suspended and possibly basins surface water pollutant removal dissolved pollutants but may be a risk to groundwater quality Storage Storage of Sediment and Can remove pollutants associated ponds surface water pollutant removal with sediments and provide some biological treatment Wetlands Pollutant removal Storage Can remove and treat various pollutants

France has also had experience in particular aspects of SUDS (for detention basins and ponds). Nascimento et al. (1999) provide an overview of the French experience in detention basins use and performance. In France, use dates back to the 1960s, together with the construction of the “Villes Nouvelles”. Their use was limited, but these facilities are increasingly popular as indicated in Deutsch et al. (1990) (Reported in Nascimento et al., 1999). Recent research has focused in the quality side of performance of the ponds with the QASTOR database created by CEREVE as a centre point (Saget et al., 1998). The database aims to collect and analyse all French national data related to urban wet weather discharge that have been collected in 19 catchments since 1970. The efficiency of detention basins in reducing pollutants is the result of a large number of variables including physical, chemical and biological characteristics of pollutants, precipitation regime, detention time and quality of maintenance services (Nascimento et al., 1999). Table b.2 identifies the potential annual and short-term efficiency of detention basis recorded by Adler (1993) as reported in Nascimento et al, (1999). The Table draws from studies conducted by the French institution CEMAG-REF. The presented data were for basins installed in separate drainage systems and the results indicate that such storage facilities can have a reasonable performance even over quite short time scales.

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

Yearly Inflow Yearly Outflow Reduction Reduction after (kg/ha imp) (kg/ha imp) (%) 2h (%) Pb 0.893 0.054 94.0 65 Zn 5.12 0.66 87.1 77 Cd 0.0310 0.0051 83.7 - Cu - - - 69 Hydrocarbons 65 4 94.2 -

However, despite such evidence, when the issue of integrating detention basins into the urban context is concerned, the outcome may be very different according to the specific case. For the UK the high pollutant loading to urban detention basins, which has been reported, has led to concerns about long-term siltation (and loss of effective storage volume) and water quality (especially in terms of health risks). Sansalone (1999) describes the

131 Section 6. Case Studies results of measurements taken in a field scale infiltration trench. Figure b.7, summarises some of the findings, indicating significant potentially toxic element removal efficiency exceeding 80% after 1 year of runoff loadings.

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

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

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

In particular, a multi-year research project sponsored by the US EPA addresses the potential problem of groundwater contamination due to stormwater infiltration (Pitt, Clark and Palmer, 1994; Pitt et al., 1997; Pitt, Clark and Field, 1999). In the case of pesticides the research found that heavy repetitive use of mobile pesticides, such as EDB, on site with infiltration devices likely contaminates groundwater. Fungicides and nematocides must be mobile in order to reach the target pest and hence, they generally have the highest contamination potential. Pesticide leaching depends on patterns of use, soil texture, total organic carbon

132 Section 6. Case Studies content of the soil, pesticide persistence, and depth to the water table (Shirmohammadi and Knisel 1989). A pesticide leaches to groundwater when its residence time in the soil is less than the time required to remove it, or transform it to an innocuous form by chemical or biological processes. The residence time is controlled by two factors: water applied and chemical adsorption to stationary solid surfaces. Volatilization losses of soil-applied pesticides can be a significant removal mechanism for compounds having large Henry’s constants (Kh), such as DBCP or EPTC (Jury, et al. 1983). However, for mobile compounds having low Kh values, such as atrazine, metolachlor, or alachlor, it is a negligible loss pathway compared to the leaching mechanism (Alhajjar, et al. 1990).

Restricted pesticide usage in areas with high infiltration potential has been recommended by some U.S. regulatory agencies. The slower moving pesticides were recommended provided they were used in accordance with the approved manufacture’s label instructions. These included the fungicides Iprodione and Triadimefon, the insecticides Isofenphos and Chlorpyrifos and the herbicide Glyphosate. Others were recommended against, even when used in accordance with the label’s instructions. These included the fungicides Anilazine, Benomyl, Chlorothalonil and Maneb and the herbicides Dicamba and Dacthal. No insecticides were on the “banned list” (Horsley et al, 1990).

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

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

Compounds Mobility Abundance Fraction Contamination Contamination Contamination (worst case: in storm- filterable potential for potential for potential for sub-surface sandyl-1ow water surface infilt. and surface infilt. with injection organic soils) no pre-treatment sedimentation with minimal pre-treatment Nutrients nitrates mobile low/moderate high low/moderate low/moderate low/moderate Pesticides 2,4-D mobile low likely low low low low g-BHC (lindane) intermediate moderate likely low moderate low moderate malathion mobile low likely low low low low atrazine mobile low likely low low low low chlordane intermediate moderate very low moderate low moderate diazinon mobile low likely low low low low Other VOCs mobile low very high low low low organics 1,3-dichlorobenzene low high high low low high anthracene intermediate low moderate low low low benzo(a) anthracene intermediate moderate very low moderate low moderate bis (2-ethylhexyl) intermediate moderate likely low moderate low? moderate phthalate butyl benzyl phthalate low low/moderate moderate low low low/moderate fluoranthene intermediate high high moderate moderate high fluorene intermediate low likely low low low low naphthalene low/inter. low moderate low low low penta- chlorophenol intermediate moderate likely low moderate low? moderate phenanthrene intermediate moderate very low moderate low moderate pyrene intermediate high high moderate moderate high Potentially nickel low high low low low high toxic cadmium low low moderate low low low elements chromium inter./very low moderate very low low/moderate low moderate lead very low moderate very low low low moderate zinc low/very low high high low low high Salts chloride mobile Seasonally high high high high high

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Conclusions

The control of diverse pollutants requires a varied approach, including source area controls, end-of-pipe controls, and . All dry-weather flows should be diverted from infiltration devices because of their potentially high concentrations of soluble potentially toxic elements, pesticides, and pathogens (Pitt et al., 1999) Similarly, all runoff from manufacturing industrial areas should also be diverted from infiltration devices because of their relatively high concentrations of soluble pollutants. In areas of extensive snow and , winter snowmelt and early spring runoff should also be diverted from infiltration devices.

All other runoff should include pre-treatment using sedimentation processes before infiltration, to both minimize groundwater contamination and to prolong the life of the infiltration device (if needed). This pre-treatment can take the form of grass filters, sediment sumps, wet detention ponds, etc., depending on the runoff volume to be treated and other site-specific factors. Pollution prevention can also play an important role in minimizing groundwater contamination problems, including reducing the use of galvanized metals, pesticides, and fertilizers in critical areas. The use of specialized treatment devices can also play an important role in treating runoff from critical source areas before these more contaminated flows commingle with cleaner runoff from other areas (Pitt et al., 1999). Sophisticated treatment schemes, especially the use of chemical processes or disinfection, may not be utilised, provided there is no danger of forming harmful treatment by-products (such as THMs and soluble aluminium).

The use of grass swales and percolation ponds that have a substantial depth of underlying soils above the groundwater is preferable to using dry , trenches and especially injection wells, unless the runoff water is known to be relatively free of pollutants. Surface devices are able to take greater advantage of natural soil pollutant removal processes. However, unless all percolation devices are carefully designed and maintained, they may not function properly and may lead to premature hydraulic failure or contamination of the groundwater (Pitt et al., 1999).

It should be clear that although SUDS have great potential in both quantity and quality control in urban runoff, each case should be assessed individually, and an incremental approach containing both high tech and low-tech solutions is the most likely development scenario (Butler and Parkinson, 1997). Direct application of such methods across different regions and countries is not always appropriate and must also include consideration of the local socio-economic and administrative circumstances associated with the operational design, which can be primary inhibitors to the implementation of innovative technology.

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(c) Artisanal Activities: Pollutant Sources and load in Urban Wastewater in Vicenza, Northern Italy; Gold Jewellery – Best Environmental Practice

Pollutant sources and load in urban wastewater in Vicenza, northern Italy

Introduction

In Northern and Central Italy there is a high density of small-scale, artisanal enterprises and activities. For example, in the region of Veneto (North-Eastern Italy), artisanal activities account for 20% of the regions exports. The area of Vicenza, a provincial town in the Veneto region, represents one of the largest agglomerations of artisanal activities in Italy. With a population of 109,000 inhabitants, the municipality of Vicenza has a total of about 1600 small to medium-scale enterprises (SMEs).

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

Due to the presence of specific pollutants, wastewater discharge is regulated in the same way as industrial wastewater (i.e. in terms of pollutant concentration limits), even though artisanal wastewater flow rates may be orders of magnitude lower than industrial ones [Italian law by decree n. 152, 1999]

EBAV (Ente Bilaterale Artigianato Veneto, a non-profit bilateral organisation representing the interests of both artisanal workers and enterprises) sponsored a study on the origin and contribution of pollutants to urban wastewater. The aim of the study was to assess the load of pollutants from different artisanal activities, in comparison to the total load originating from the urban wastewater system. In addition, the pollutant load from artisanal activities was subdivided into load from discharged wastewater and load from concentrated liquid wastes (which are separated and collected by external firms), to assess if wastewater segregation could significantly affect pollutant load from artisanal activities.

Description of the study The EBAV study was carried on in the period 1994-1995 in the municipality of Vicenza. According to local authorities, the only notable change (in terms of residential population and type of SMEs) since the time of the study, has been the rapid increase in the number of “service” enterprises (for example software companies), which do not contribute specific wastewater. Therefore, this growing number of “service” companies does not affect the conclusions of the study, which still may be considered valid today.

Wastewater for the whole Vicenza area is treated by four WWTS (Table c.1). The total capacity is about 137,900 p.e. (population equivalent) with a total flow rate of about 23.2 million m3 year-1.

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

WWTS Capacity (p.e.) Treated flow rate m3 year-1 CASALE 71,900 10,202,000 LAGHETTO 3,500 394,000 LONGARA 3,500 826,000 S. AGOSTINO 59,000 11,800,000 Total 137,900 23,222,000 In this area there are about 1,579 artisanal enterprises discharging their wastewater into the UWW collecting system. Table c.2 shows the most common artisanal activities in the area of Vicenza and the number of enterprises involved in each activity. For each activity a representative number of enterprises was selected for further investigation. Typically one wastewater sample was drawn from each enterprise. In some cases an additional sample of concentrated, segregated wastewater was also drawn.

Table c.2 Main artisanal activities in the area of Vicenza Type of activity No. of samples No. of enterprises in the municipality of Vicenza Food workshops 10 53 Car-repairers 20 (14+6*) 175 Ceramic and photoceramic 7(6+1*) 23 Artisanal galvanic shops 8(5+3*) 18 Printing shops 21(14+7*) 140 Wood manufacturing 18(3+15*) 92 Marble manufacturing 5 140 Metallurgists and mechanics 15(8+7*) 155 Dental practices 21 88 Gold manufacturing shops 34 258 Hairdressers 19 310 Laundrettes and dry-cleaners 25 88 Textile shops 2 16 Artisanal glass manufacturing 3 23 TOTAL 208 1579 *concentrated wastewater, segregated and committed to external firms

Total load Along with artisanal wastewater samples, influent and effluent samples from the four municipal WWTS were analysed for a large number of pollutants including: B, Cd, Cr(III), Cr(VI), Mn, Ni, Pb, Cu, Zn, and anionic surfactants. From each of the four WWTS a large number of influent and effluent samples were taken and analysed. Using the specific wastewater flow rate and the influent concentration, the pollutant load was calculated for each plant. Table c.3 reports the total load as sum of the pollutant loads of the WWTS, assuming that urban wastewater is treated by only one hypothetical centralized plant.

Total pollutant load in Table c.3 is reported as average value, based on the average concentration from 20-50 samples. In addition, the maximum load and the upper 95% confidence interval are given. The last column reports the average removal efficiency based on the comparison of influent and effluent pollutant concentrations.

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Table c.3 Pollutant load to the hypothetical centralized WWTS of Vicenza municipality

POLLUTANT TOTAL LOAD REMOVAL (g/day) EFFICIENCY(%) AVERAGE MAX(95%) MAX Cd 64 75 960 40 Cr(III) 636 759 13244 75 Cr(VI)

A typical example of the analytical work performed for each category of artisanal activity is reported in Tables c.4 and c.5 for car repair shops. A rough statistical analysis of the results has been performed to obtain the average concentration and the upper limit of the 95% confidence interval (Max 95%). In addition the maximum value (Max) is also reported.

Table c.4 Pollutant concentrations (mg l-1) in discharged wastewater of 14 car-repair shops Pollutant Average Max 95% Max CI COD 329 547 1800 Cd 0.01 0.03 0.14 Cr(III) 0.1 0.2 0.8 Cr(VI) 0 0 0 Mn 0 0 0 Ni 0 0 0 Pb 0 0.1 0.4 Cu 0.1 0.1 0.5 Zn 4.5 11.1 55

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

Pollutant Average Max Max 95%CI COD 10293 15346 19120 Cd 0.3 0.7 1.3 Cr(III) 0.2 0.3 0.5 Cr(VI) 0 0 0 Ni 2.5 6.1 12 Pb 22.1 52 102 Cu 33 74.6 144 Zn 31.9 73.8 145

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

138 Section 6. Case Studies the wastewater is discharged directly into the UWW collecting system. Similar calculations were performed using concentrations and volumes of segregated wastewater (spent baths). So, for the enterprises that separate wastewater for treatment by external firms (for recovery and detoxification), two potential pollutant loads were given with reference to the two hypothesized scenarios:

· all the wastes (concentrated and diluted wastewater) are discharged into the urban wastewater system without pre-treatment (Table c.6 D and S); · the spent baths are treated externally, whereas wastewater is directly discharged into the urban wastewater system (Table c.6 D only).

Car-repairers (175 shops)

The most significant pollutants are: suspended material, COD, oils, surfactants, organic solvents, copper, and zinc.

Considering both discharged and segregated wastewater the total pollutant load for Zn, Cd, Cu and, above all, Pb is very high. However upon careful segregation of concentrated wastewater (D only), the pollutant load is significantly reduced. The careful segregation of spent baths induces a decrease of about one order of magnitude in the percentage of lead originating from car repair shops.

Ceramics and photoceramics (23 shops)

The principal pollutants from these artisanal activities are: suspended solids, lead, ammonia, nitric nitrogen, and surfactants. As can be seen in Table c.6 the pollutant load from the ceramic and photoceramic shops is minimal due to the low wastewater flow rate. Only lead seems to represent a significant load. In the case of ceramic shops, segregation of concentrated wastewater is not very significant in terms of reducing pollutant load.

Galvanic (18 shops)

The principal pollutants from these enterprises are: suspended solids, chromium (VI), nickel, lead, copper, and cyanide. Segregation and external treatment of concentrated wastes does not appear to significantly reduce the lead load because this load originates mainly from the discharged diluted wastewater.

Printing shops (140 shops)

Printing activities generate several pollutants: suspended material, COD, cadmium, chromium, lead, copper, zinc, sulphites, sulphates, chlorides, ammonia, total phenols, aldehydes, aromatic organic solvents, and surfactants. Due to the low contribution to the overall UWW flow rate (0.15%), only Cd and Cu loads from the printing activities are significant with respect to the total load. Segregation of concentrated wastewater reduces the load of metals to negligible values.

Wood processing and furniture making shops (92 shops)

The most significant pollutants originating from this activity are: suspended solids, COD, lead, copper, zinc, total phenols, organic solvents, and surfactants. Table c.6 shows that by simply segregating concentrated wastes the pollutant load to the WWTS from wood processing and furniture manufacturing shops is dramatically reduced. Other specific wood processing pollutants such as arsenic were not analysed.

139 Section 6. Case Studies

Metallurgists and mechanics (155 shops)

Several pollutants originate from metallurgists and mechanic shops: suspended solids, COD, cadmium, chromium, nickel, lead, zinc, copper, sulphates, chlorides, phosphorus, oils, solvents, and surfactants. Pollutant concentrations in wastewater are typically much lower than in segregated wastewaters (see Table c.7) which have high average concentrations of pollutants such as; nickel, lead, copper and zinc. Separation of concentrated wastes reduces the load to the WWTS significantly.

Goldsmiths (258 shops)

The area of Vicenza represents one of the most important districts for gold manufacturing in Italy, with up to 258 artisanal goldsmith shops. The main pollutants originating from gold manufacture are: COD, boron, cadmium, copper, zinc, and surfactants. However goldsmiths shops are characterized by very low wastewater flow rates, with an average of about 1 cubic meter per day per unit. In terms of contribution to the pollutant load, the 258 goldsmiths shops represent a high contribution of Cd, Cu, and Zn. This is mainly due to the fact that goldsmith shops used to add spent concentrated baths to the diluted wastewater in order to recover precious metals during the wastewater treatment before discharging it into the sewer.

Food workshops (confectioners, ice-cream parlours, bakeries) (53 shops)

The most significant pollutants originating from this activity are: COD, fat, oils, and surfactants. Cu and Zn are the only metals to be above the limits of detection in wastewater.

Dental technicians (88 shops)

Principal pollutants originating from this activity are: suspended solids, COD, and surfactants. The contribution of the 88 shops to the total metals pollutant load is generally low, due to the low contribution in terms of flow rate, that is an average of about 0.07% of the total UWW flow rate. Mercury is an important pollutant in wastewater linked to dental practices, which is not considered in this particular study but is referred to in section 2.1.2 of this report.

Hairdressers (310 shops)

This is the category with the largest number of shops in the Vicenza’s municipality. The main pollutants originating from hairdressers are: suspended solids, COD, and surfactants.

Laundrettes and dry-cleaners (88 shops)

The main pollutants originating from laundrettes and dry-cleaners are: suspended solids, COD, surfactants, chlorides, and solvents.

The most significant load of pollutant from the 88 laundrettes and dry-cleaners is for Cd although overall the levels for this pollutant were very low .

140 Section 6. Case Studies

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

Flow Average Pollutant Load (g/day) rate m3 per day Cd Cr III Mn Ni Pb Cu Zn Car repair shops 153 4 6.3

Impact of all artisanal activities on urban wastewater treatment plants

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

141 Section 6. Case Studies

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

POLLUTANT TOTAL POLLUTANT LOAD FROM PERCENTAGE ARTISANAL ACTIVITIES (g/day) OF TOTAL LOAD TO WWTS average max(95%) max B 2909 4774 38832 Cd 30 66 529 46.59 Cr(III) 54 106 243 8.55 Cr(VI) 14 49 43 Mn 14 19 38 0.28 Ni 1180 3153 8114 16.63 Pb 452 1064 2676 71.12 Cu 903 1872 6387 25.48 Zn 2231 3998 10589 19.88 MBAS 53404 93158 353795 25.45 flow rate m3/day 1633 2526 5559 2.57 p.e. 7697 13295 46221 5.58

Table c.8 Total pollutant loads from artisanal activities in the Vicenza municipality, considering discharged wastewater only POLLUTANT TOTAL POLLUTANT LOAD FROM PERCENTAGE ARTISANAL ACTIVITIES (g/day) OF TOTAL LOAD TO WWTS average max(95%) max B 2909 4774 38832 Cd 24 51 498 36.9 Cr(III) 11 27 100 1.8 Cr(VI)

In terms of flow rate the contribution of all artisanal activities is typically low (2.6%). It remains relatively low (4%) in the worst scenario case, where the artisanal activities discharge at the upper limit of the confidence interval of their cumulative wastewater flow rate, while the total UWW flow rate remains at the average value. In contrast, the percentage contribution of artisanal activities to the total load is very high for pollutants such as: Pb, Cd, Cu, Zn, Cr III, and surfactants.

Figures c.1 and c.2 compare the load of each specific activity to the total load of a hypothetical centralized WWTS, for potentially toxic elements and surfactants, respectively. Figure c.1 clearly shows that only car-repairers, goldsmiths shops, metallurgists and mechanics contribute significant amounts of potentially toxic elements to UWW, with respect

142 Section 6. Case Studies to the total metal load of the all activities. The main activities responsible for the surfactant load in UWW are; hairdressers, goldsmiths, and food workshops (Figure c.2).

Figure c.1 Heavy metal load from artisanal activities

9000

8000 7000

6000

5000 4000

3000

average load (g/day) 2000 1000

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

AL: Food workshops AU: Car repairers CE: Ceramics GA: galvanic GR: Printing shops LG: Wood manufacturing MA: Marble manufacturing ME: Metallurg.and mechanics OD: Dental practices OR: Goldsmiths PR: Hairdressers

Figure c.2 Surfactant load from artisanal activities

60000

50000

40000

30000

20000 average load (g/day) 10000

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

143 Section 6. Case Studies

With the exception of surfactants and cadmium, the high pollutant loads from artisanal activities are notably reduced when the concentrated wastes are not considered (Table c.23). The poor effect of waste segregation on surfactant load is explained by the fact that the major surfactant contribution derive from activities that do not practice waste segregation: hairdressers, goldsmiths shops, and food workshops (Figure c.2).

Impact of artisanal activities on sewage sludge characteristics

Assuming that the fate of sewage sludge produced from the hypothetical centralized WWTS is used in agriculture, the maximum admissible pollutant concentration in sludge must be considered. From this value a maximum admissible pollutant concentration in the UWW may be calculated according to the following equation:

Ci Qf 100 %H 2O C R% Q where C is the maximum influent concentration permitted for sludge disposal Ci is the maximum pollutant concentration allowed in sludge to land R% is the removal efficiency (%) in the treatment plant 3 -1 Qf is sludge flow rate in m day calculated on the basis of the production of 1.87 l/inhabitant with 95.5% of humidity Q is the influent wastewater flow rate % H2O is the water content of the sludge (95.5%)

Table c.9 Pollutant contribution of artisanal activities to the admissible load for sludge disposal in agriculture, considering both discharged and segregated wastewater

Pollutant Regulatory limits for Admissible(*) Admissible(*) percentage of total agricultural use concentration load admissible load (mg/kg dry sludge) (%) D.Lgs 99/92 Veneto (mg/l) (g/day) average Cd 20 10 0.01 580 5.11 Cr(III) 500 0.15 9670 0.56 Ni 300 200 0.17 10549 11.18 Pb 750 500 0.21 13599 3.33 Cu 1000 600 0.28 17582 5.14 Zn 2500 2500 0.66 42045 5.34 (*)on the basis of equation (1)

144 Section 6. Case Studies

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

Pollutant Regulatory limits for Admissible(*) admissible(*) percentage of total agricultural use conc load admissible load (mg/kg dry sludge) (%)

D.Lgs 99/92 Veneto (mg.l-1) (g/day) average Cd 20 10 0.01 580 4.04 Cr(III) 500 0.15 9670 0.12 Ni 300 200 0.17 10549 0.31 Pb 750 500 0.21 13599 0.37 Cu 1000 600 0.28 17582 1.92 Zn 2500 2500 0.66 42045 2.35 (*)on the basis of equation (1)

In Tables c.9 and c.10 the concentration limits imposed by law for sludge used in agriculture are reported. From these limits the concentration limit in the influent and the consequent admissible load were calculated. In the last column the impact of artisanal activities on the sludge characteristics are reported in terms of percentage load from artisanal activities with respect to the total admissible load for sludge disposal (calculated using average values).

Table c.10 considers the highest possible pollutant loads in the hypothesis that waste segregation does not take place. Even in this pessimistic hypothesis, the impact of artisanal activities is typically low. The highest metal contribution is for Ni, representing 11% of the admissible load. In the worst case scenario, considering all the maximum loads, artisanal activities by themselves almost reach the admissible Cd and Ni loads.

The average impact on sewage sludges, without considering the concentrated spent baths (Table c.9) is less than 5% for all the metals.

Validation of the Case Study with results from a recent study on hairdressers’ shops in a different area of Veneto region

To validate the results obtained in the study described above, EBAV undertook an additional study specifically addressed to activities of hairdressers and beauticians. This study, performed during 1999-2000, considered a group of shops representing all the hairdressers and beauticians located in the district of Valdagno (Vicenza), which discharge their wastewater into the UWW collecting system of the municipality of Trissino (Vicenza). In this validation study the number of shops examined was about 22% of those present in the area, compared to only 6% of hairdressers in the Vicenza Case Study. Typically one wastewater sample was drawn from each shop. A rough statistical analysis of the results has been performed to obtain the average concentration; then the average values have been increased by 10% to take into account the effects of highly polluted wastewater (due to typical products such as shampoos or dyes).

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

145 Section 6. Case Studies

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

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

Table c.27 Pollutant concentrations and pollutant loads from hairdressers and beauticians in the district of Valdagno (Vicenza) POLLUTANT AVERAGE POLLUTANT LOAD* PERCENTAGE OF LOAD CONCENTRATION (g/day) TO WWTS (mg/l) (%) Cr(III) <0.1 <5 <0.1 Anionic surfactants 44 3915 8.2 Non ionic surfactants 51 4590 6.2 Total surfactants 99 8775 7.2 Flow rate 0.6 (m3/day x shop) 81 (m3/day) 0.33 *based on average + 10%

Conclusions

Cases like the Vicenza district, with artisanal activities deeply rooted in residential areas are common in Italy and in other EU regions. As shown in the validation case of Valdagno, in other districts the contribution of artisanal activities to the pollutant load of the UWW system may be lower.

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

The main conclusion of this study was that, by segregating concentrated liquid wastes, the contribution of artisanal activities to the pollutant load was dramatically reduced, at least for potentially toxic elements. However, only some of the artisanal activities in this Case Study practised wastewater segregation. One issue raised by artisanal representatives was the economic cost of segregation of wastewater. It is felt that the stringent environmental requirements concerning wastewater from Italian artisanal shops, considered industrial wastewater, do not compensate efforts for waste segregation.

Artisanal activities that did not practice segregation of concentrated liquid wastes include goldsmiths, hairdressers and food manufacturing shops, which are also the main enterprises responsible for the surfactant load in wastewater. As a consequence, neglecting the contribution of segregated liquid wastes did not significantly affect the total load of surfactants from artisanal shops. This load typically represents one fourth of the total surfactant load to the WWTS. It may be anticipated that, if careful segregation of the concentrated liquid wastes were extended to all the artisanal activities, a dramatic decrease of potentially toxic element and surfactant load from the artisanal activities would be obtained. Even though the wastewater flow rate from artisanal shops would not decrease significantly, the pollutant load could be reduced to negligible values with respect to the total pollutant load.

146 Section 6. Case Studies

GOLD JEWELLERY PRODUCTION IN ITALY- BEST ENVIRONMENTAL PRACTICE

Introduction

In Italy there are about 6,000 small to medium gold and jewellery manufacturing shops, most of which are concentrated in three main production districts: Arezzo (Tuscany), Vicenza (Veneto) and Valenza Po (Piemonte). In the period 1995-1999 the Italian Government and the Association of Artisanal Activities sponsored a large research programme, carried out by the National Research Council, in support of Craft Goldsmiths Production and Trade. The programme tackled problems relating to:

· innovation in production cycles, · fast analytical tools for the assay of precious metals and their alloys, · safety and health of artisanal workers, · environmental impact of gold manufacturing shops.

The Italian Water Research Institute (IRSA) carried out a survey on management practices for wastewater, produced in small to medium gold manufacturing shops in Arezzo (Marani, 1997). According to the Association of Goldsmiths, the results obtained in the survey of Arezzo district may confidently be extended to draw conclusions about national practices in goldsmiths’ shops. There is no information on the losses of Au, Ag, PGMs from the gold and jewellery shops to the wastewaters in this research programme. More data on PGM is presented in Case Study (a).

The gold manufacturing district of Arezzo

Gold manufacturing processes

The most prevalent processes are those starting from wire or plate to produce rings, chains and medals. Hollow bars may be used to produce lighter objects. Gold or silver goods may also be produced using micro-casting or electrolytic processes (electro-forming). All production cycles have common final steps: object assembly, polishing, finishing and cleaning.

Wastewater origin and characteristics

Different production cycles and processes generate wastewater in the gold manufacturing shops. Casting operations to prepare wire or plate do not typically require aqueous solutions, with exception of small volumes for washing crucibles. In contrast, the preparation of hollow bars requires nitric acid, hydrochloric acid, caustic soda, and ammonia solutions.

The wastewater resulting from micro-casting comes from water used to break the gypsum mould and rinse waters. The “gypsum” waters are segregated from the other wastewater produced in the workshops and recycled after a settling step.

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

147 Section 6. Case Studies re-circulated after elution through an column (anionite). The concentrated cyanide solutions produced by column regeneration are sent to external firms.

In the final steps of assembly and finishing, the operations producing liquid wastewater are:

· acid pickling, · galvanic treatments, · surface shining, · washing steps.

Usually spent pickling baths are treated by external firms. The waters derived from surface shining generally contain metal powder. In addition, the final step generates large amounts of wastewater as well as surfactants present in the spent baths.

Finally hand and floor washing waters, do not typically require pre-treatment and could be sent to the sewer. However as they may contain gold and silver powder they are not directly discharged into the sewer. Instead they are sent to the internal wastewater treatment plant. Here the insoluble precious metals are concentrated in the sludge which, is dried and sent to specialised firms for recovery of precious metals.

Wastewater may be divided into four classes:

Soapy waters contain high concentrations of detergents, along with fatty substances and metal powder. Other pollutants such as phosphates and ammonia typically originate from the detergents used in these workshops.

The acids contained in the acid wastewater are: sulphuric acid, nitric acid, hydrochloric acid, hydrofluoric and fluoboric acid. This wastewater may also contain high concentrations of potentially toxic elements such as copper, zinc, iron and nickel.

“Gypsum” waters containing suspended gypsum particles, are recycled several times after sedimentation of the suspended material. Then they are committed, together with the sedimented gypsum, to external firms for final disposal.

The cyanide waters contain free cyanide and soluble cyano-complexes of gold and silver. For safety reasons these waters are treated separately to oxidize the cyanide before sending them to the wastewater treatment plant.

148 Section 6. Case Studies

Wastewater management and fate About 40% of the goldsmiths’ workshops of Arezzo province do not declare any wastewater production. These workshops: · have a particular production step that does not produce wastewater; · accumulate little quantities of wastewater to be treated by external firms specialized in wastewater treatments; · treat their own wastewater with systems that permit the complete recycle of the treated water in the productive cycle; · evaporate the wastewater, obtaining a concentrate to dispose with other solid wastes.

Regarding the remaining 60% of workshops with authorisation to discharge their wastewater, Table c.28 reports the number (and relative percentage) of workshops that discharge their wastewater into the UWW collecting system or into surface waters. In terms of flow rate, the 694 workshops discharge about 121600 m3.year-1 of wastewater. Assuming an average number of 6.3 workers per unit, the average specific wastewater flow rate is about 0.5 m3.week-1 per worker.

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

Destination No. of workshops % of workshops UWW collecting 567 81.7 system Surface waters 29 4.2 Unknown 98 14.1 TOTAL 694 100

Survey on a representative group of goldsmiths’ shops The workshops were selected with the aim of choosing representative establishments in terms of number of workers and in terms of type of manufacturing processes. Twelve small to medium workshops were selected for the study. The survey included both interviews on production processes, wastewater flow rate, wastewater treatment, direct sampling and analysis of treated and untreated wastewater. Typically, several wastewater samples were collected from each workshop, in March, May and October of 1996. The chemical characterisation of samples was performed analysing a large number of parameters, including: boron, potentially toxic elements like Cd, Cr, Cu, Ni, Pb, and Zn, and surfactants.

Conclusions In examining the pollutant concentrations in these wastewaters (considering the results obtained for the 12 shops sampled), the pollutants most often detected with concentrations higher than admissible limits for discharge are: surfactants, copper, zinc, cadmium and boron. Surfactants, copper and zinc are detected in all samples whereas boron and cadmium are present in 75 % and 60% of samples respectively. Surfactants, derived generally from washing processes, are present in wastewater with an average concentration of 34 mg.l-1 MBAS and a maximum value of 118.5 mg.l-1 MBAS. The average boron concentration found in the wastewater of small jewellery shops examined is 13.5 mg.l-1, with a maximum value of 100 mg.l-1. The presence of boron in these wastewaters may be due to processes such as soldering (where boron is used as borace), voiding and washing, or through the use of other materials e.g. hydrofluoboric acid, detergents.

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

149 Section 6. Case Studies

(D) Pharmaceuticals in the Urban Environment

Introduction

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

Current research demonstrates that drugs and their metabolites entering water supplies and the food chain may pose a real threat, both to the ecosystem and to human health, and risk assessments are slowly being carried out. However, many problems must be overcome, such as the fact that these compounds are very changeable and are usually present in mixtures and at low concentrations. Furthermore, pharmaceuticals have a wide variety of structures and activities and that they may act synergistically (Alcock, 1999). There are many different pharmaceuticals substances and approximately 3000 pharmaceutical compounds are discharged into UWW collecting systems (ENDS, 2000). Sewage sludge is predominantly disposed of on agricultural land, as is manure from farms and both of these products will contain large amounts of pharmaceutical substances. Unfortunately, very little is known about the fate of these compounds in the environment and the potential long-term impacts. Many pharmaceutical substances though, have the same characteristics as organic compounds; i.e. they are lipophilic, which tends to be a requirement to be able to pass membranes, and some are designed to be persistent so that they are not inactivated before achieving their healing effect (Halling-Sørensen, 1998).

Sources and fate in the environment There are two major groups of pharmaceuticals; human and veterinary drugs, and they will enter the environment through different pathways (Figure d.1).

A large amount of pharmaceutical products from both categories are prescribed each year. For example in Germany, 100 tonnes of human drugs were prescribed in 1995 (Ternes, 1998c). This probably reflects the amounts prescribed in other countries of Western Europe, relative to population size. Over the counter pharmaceutical sales will also increase this figure. It is likely that a high concentration of drugs may find their way into wastewater, making wastewater and sewage sludge major vectors for the entry of these compounds into the environment. However, this will depend on the chemico-physical behaviour of the pharmaceuticals in question.

150 Section 6. Case Studies

Figure d.1: Scheme for the main fates of drugs in the environment after application [after Ternes, 1998c.]

The main entry routes of pharmaceutical substances into the environment are through; disposal of wastewater treatment end products: effluent and sewage sludge; and manure spreading onto agricultural land or even from the excreta of grazing animals (Figure d.2). Fish farms also use medical substances as feed additives, but most of the food is not eaten and is deposited straight onto the sea-bed (Halling-Sørensen, 1998).

151 Section 6. Case Studies

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

The compounds For the purpose of risk assessments, pharmaceutical compounds have been divided into four activity groups: antibiotics, antineoplastic drugs, antiparacetic drugs, and hormone disrupters. However, there are many other groups of pharmaceutical compounds, such as lipid regulators, which have been found ubiquitously in the environment and are highly persistent compounds (Daughton, 1999).

152 Section 6. Case Studies

Antibiotics: Antibiotics are widely used as medicines for human and animals treatment, also used widely as growth promoters in veterinary use. According to the Swiss environmental research institute, in the EC 54000 tonnes of antibiotics were used in human medicine in 1997. Veterinary use amounted to 3500 tonnes of medicines and 1600 tonnes of growth promoters (ENDS, 2000). Due to the effect of bans the use of growth promoters is expected to decline. According to a study carried out by Halling-Sorensen (1998), most antibiotics are not very persistent in the environment, particularly in soils, and the most widely used growth promoters have been shown to have no effect on invertebrates, even at relatively high concentrations. However soil bacteria may be more sensitive (ENDS, 2000).

Veterinary drugs tend to end up in manure and so have the potential to contaminate soils where manure or slurry is spread. Levels of antibiotics in soil around a pig farm studied reached up to 1400mg.kg-1, due to presence of antibiotics in the animals’ feed (ENDS, 2000). The increased use of antibiotics has led to an increase in drug resistant micro flora. This resistance is actually favoured by low concentrations of antibiotics (Jorgensen, 2000), thus, the presence of antibiotics in the environment may be an important problem.

Anti-neoplastic drugs These anti-cancer drugs are mainly used in hospitals rather than in households. They are primarily used for chemotherapy and are found sporadically, in a range of concentrations in the environment (Daughton, 1999). Anti-cancer drugs act as non-specific alkylating agents, which means that no receptors are required, hence, they have the potential to act as mutagens, , teratogens, and embryotoxins (Daughton, 1999). The most widely used substance is cyclophocamide (CP). In Denmark, approximately 13 to 14 kg of CP is used in hospitals each year, and approximately 6 kg are prescribed by pharmacies (Christensen, 1998). Thus, it is assumed that a total of 20 kg are used per year (Christensen, 1998). Anti-cancer drugs are also referred to in the Case Study on Platinum Group Metals.

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

Endocrine-disrupting substances There is increasing concern about compounds that interfere with the hormonal system. Endocrine disrupting substances block or trigger oestrogenic effects by binding to receptors. -specific responses are particularly problematic as they can affect people for which they are not intended (Christensen, 1998). An endocrine-disruptor may have an ‘agonistic effect’ where it binds to the receptor instead of the natural hormone and causes a response, or it may have an ‘antagonistic effect’ where the binding of the compound prevents the natural one from binding and producing the required response (Environment Agency of England and Wales, 1998). Other effects may also occur, showing that the process is very complex and affects many systems in the body.

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

153 Section 6. Case Studies

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

Category Examples Uses Modes of action Natural phytoestrogens Isoflavones, Present in plants Oestrogenic and anti- oestrogenic Female sex 17b-oestradiol, Produced in animals Oestrogenic hormones oestrone Man-made Polychlorinated PCBs, dioxins By-products from Anti-oestrogenic organic incineration and chemical compounds processes Organochlorine DDT, dieldrin, lindane Insecticides Oestrogenic and anti- pesticides oestrogenic Alkylphenols Nonylphenol Production of NPE and Oestrogenic polymers Alklphenol NonylPhenol Surfactant Oestrogenic ethoxylates Ethoxylate (NPE) Phthalates Dibutyl phthalate Plasticiser Oestrogenic Bi-phenolic Bisphenol A In polycarbonate plastics Oestrogenic compounds and epoxy resins Synthetic steroids Ethinyl oestradiol Contraceptives Oestogenic

Of these, oestrogens are of a major concern, as they are excreted in an inactive form but are found to be reactivated in sewage effluent. Oestrogens are organic molecules derived from , which can bind to receptors and cause a physiological response (Montagnani et al., 1996). The purpose of the endocrine system is to regulate metabolic activity, which requires a degree of interaction with the nervous system (Montagnani et al., 1996). Because oestrogen receptors are located in the cell nucleus, oestrogen-like molecules can thus enter the cell and could potentially interact with DNA, causing damage which may lead to tumour formation (Montagnani et al., 1996). Prolonged exposure to these compounds may induce female characteristics in males. There is increasing speculation that these compounds may be linked to reduction in male fertility and reproductive complications (Montagnani et al., 1996).

In the UK, research has shown that male fish exposed to the natural hormone: 17b- oestradiol, oestrone, and the synthetic hormone: ethinyloestradiol, from domestic sewage effluent, developed hermaphrodite characteristics (Alcock et al., 1999). It was also found that these hormones were present in the biologically active form, having been transformed and reactivated after excretion and not degraded during wastewater treatment (Alcock et al., 1999). However, research carried out by the Ministry of Agriculture, Food and Fisheries (MAFF) along the river Lea, UK determined that although estrogenic substances are likely to be present in wastewater effluents, the development of female characteristics in male fish present in WWTS lagoons is unlikely to be due to these substances, as the transformation is only possible at a very early stage in their development (Montagnani et al., 1996). Due to the widespread use of estrogenic substances and their entry into the environment via sludge and effluent from WWTS, aquatic environments may be acting as a sink for these substances. Natural and synthetic estrogens are extremely widely present/used. Although the contribution these compounds make to oestrogenic effects is thought to be small, they are are of concern because of their highly persistent and potent nature (ENDS, 2000).

154 Section 6. Case Studies

Consumption of pharmaceutical compounds

As already stated, it is difficult to obtain information on the quantities of pharmaceuticals used for most countries but in Denmark consumption of the most commonly used drugs is available (Table d.2).

Table d.2: Major drugs and drug groups and their consumption in Denmark, in 1997 [ENDS, 2000, Christensen, 1998, and Halling-Sørensen, 1998].

Active ingredient Major use Amount used (kg) Aspirin Analgesic 305250 Paracetamol Analgesic 248250 Ibuprofen Anti-rheumatic 33792 Pencicillin V Antibiotic 19000 Diuretic 3744 Terbutaline Anti-asthmatic 475 Enalapril Anti-hypertensive 416 Anti- 368 Anti-depressant 207 Salbutanol Anti-asthmatic 170 Bendroflumethiazide Diuretic 167 Anti-hypertensive 144 Amlodipidine Anti-hypertensive 132 Oestradiol Hormone replacement 119 Anti-hypertensive 116 17b-estradiol Oral contraception 45 Budesonide Anti-asthmatic 39 Gestodene Birth control pill 37 Cyclophosphamide Anti-cancer 20 Xylometazoline Nasal decongestant 13 Digoxin Heart drug 4 Desogestrel Birth control pill 3 Medicine groups Antibiotics 37700 28300 Hypotensiva 410 Diuretica 3800 Anti-asthmatics 1700 Psychleptics 7400

It was also found that a total of 110 tonnes of antibiotics were used as growth promoters, feed additives or as medicines, on livestock and fish farms (Halling-Sørensen, 1998). In 1994, the overall production of antibiotics in Germany was 1831 tonnes, of which contributed 624 tonnes (Hirsch, 1999). It should be noted that the amount of antibiotics used for human and veterinary purposes, 37.7 tonnes and 110 tonnes respectively, is in the same range as the amounts of certain pesticides used (Hirsch, 1999). A paper published by Goll van (1993) estimates that if the total amount of growth promoters used in the Netherlands was spread on their 2 million hectares of agricultural land, this would give a yearly average of 130mg of antibiotics and metabolites/m2 (Halling-Sørensen, 1998). The pharmaceutical substances predominantly used in hospitals must also take into account compounds such as X-ray contrast media. Iodinated X-ray contrast media are very stable biochemically, so they tend to be excreted unmetabolised. In Germany, 500 tonnes per year of X-ray contrast media are used and iopromide (CAS 73334-07-3) alone accounts for 130 tonnes per year (Ternes, 2000). Other X-ray contrast substances used in the EU are:

155 Section 6. Case Studies diatrizoate (CAS 131-49-7) an ionic X-ray diagnostic drug, iopamidol (CAS 60166-93-0) and iopromide (CAS 73334-07-3) both non-ionic X-ray diagnostic substances, iothalamic acid (CAS 2276-90-6) and ioxithalamic acid (CAS 28179-44-4) both ionic X-ray diagnostic substances.

Detection of pharmaceutical compounds

Other biologically active compounds are common in wastewater, and there is increasing evidence that such substances are widely present in the environment. Danish research has found that up to 68 different drug residues can be detected in the environment. Compounds such as caffeine, , aspirin, and paracetamol are all frequently detected (ENDS, 2000). More studies are now looking at the prevalence of these drugs and their metabolites in wastewater and sewage sludge. For example, clofibric acid (2-(4-chlorophenoxy)-2- methylpropionic acid), which is a breakdown product of lipid-regulating drugs, is highly persistent and resistant to wastewater treatment, as usually only 15-51% is removed (ENDS, 2000). Lipid regulating drugs are commonly prescribed and it is thought that the daily load of clofibric acid to UWW collecting systems in Denmark, is around a few kilograms (ENDS, 2000). Concentrations of clofibric acid have also been detected in sewage effluent in micrograms per litres, and in nanograms per litres in water bodies such as rivers and (ENDS, 2000). In the UK, clofibric acid has been detected in the 1 mg.l-1 range in the aquatic environment, and in Germany, it has been detected at concentrations up to 165 ng.l-1 (Ternes, 2000).

Antibiotics have been also widely detected. In Germany, concentrations up to 5 mg.l-1 were found in WWTS effluents, which is comparable to the data collected by Richardson and Bowron in 1985 (Hirsch, 1999). Five of the 18 compounds investigated were frequently detected in German WWTS effluent and rivers: , , , sulfamethoxazole, and trimethoprim. The highest concentration was detected for erythromycin degradation products, at a median value of 2.5 mg.l-1 in WWTS effluent and a maximum value of 6 mg.l-1 (Hirsch, 1999). The other four antibiotics were only detected at levels below 1 mg.l-1 (Hirsch, 1999). Median values for the concentrations detected in surface waters are one order of magnitude lower than those detected in WWTS effluent (Hirsch, 1999), (see Figure d.3).

156 Section 6. Case Studies

Figure d.3: Presence of antibiotics from investigated surface waters in Germany [Hirsch, 1999].

Analyses for tetracycline and found no detectable amounts in five WWTS effluents or surface waters (Hirsch, 1999). Tetracyclines tend to form stable complexes with calcium and other ions, thus contaminating the sediment rather than the water (Hirsch, 1999). Penicillins tend to be easily eliminated, as they are very susceptible to hydrolysis of the b- lactam ring (Hirsch, 1999). Ground water samples were also slightly contaminated by sulphamethoxazole and sulphamethazine (not used in human medicine), due to infiltration from application of contaminated sewage sludge or manure to agricultural land. The samples collected contained concentrations of sulphonamide residues up to 0.48 mg.l-1 (Hirsch, 1999).

X-ray contrast media are also widespread in German wastewater influent and effluent. Loads of the most frequently used compounds, such as iopromide were found to exceed 1mg.l-1 during the working week but decreased at weekends as X-rays did not tend to be performed (Ternes, 2000). Maximum concentrations detected were greater than 3 mg.l-1 (up to 15 mg.l-1 for iopamidol). Median values were around 0.25-0.75 mg.l-1, which indicates their ubiquity in German wastewater treatment effluents (Ternes, 2000). The compounds detected depended on the region and on the practices of particular hospitals in that area.

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

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

157 Section 6. Case Studies

Oestrogens. There is increasing concern about the widespread presence of oestrogenic substances in wastewater and other water bodies. The daily production rate of natural oestrogens by humans is in the microgram range, up to 400 mg of 17b-estradiol for women (Ternes, 199b). The maximum daily excretion rate is 64 mg for oestriol (Ternes, 1999b). Oestrogens are mainly excreted as inactive polar conjugates (Ternes, 1999b). Vitellogenin (precursor for production of yolk in all oviparous vertebrates) induction in male or juvenile fish has become a “biomarker” for the presence of estrogenic substances in the aquatic environment (Larsson, 1999). In the UK, caged fish downstream of an WWTS were found to produce vitellogenin. Two possibilities were investigated: the presence of ethinylestradiol and NPE. The effluent from a Swedish WWTS showed high levels of oestrogenic compounds (Larsson, 1999). It was found that exposure to large amounts of ethinyloestradiol caused accumulation in fish, as fish concentrations were found to be 104- 106 times higher than those detected in the water (Larsson, 1999). The estimated use of ethinyloestradiol is 3.5mg/day, which is close to the concentration found in the WWTS effluent (2.9mg/day), showing very low degradation of this compound during treatment (Larsson, 1999).

In the UK, analysis of sampled effluents found that natural hormones (17b-oestradiol and oestrone) are present in the range 1.4 to 76 ng.l-1, whereas the synthetic hormone (ethinyloestradiol) was only found in 3 out of the 7 effluents analysed and at comparatively low levels: 0.2 to 7 ng.l-1 (Alcock et al., 1999). The source of these is thought to be mainly from human excretion products. It was also found that the hormones were present in the biologically active form, suggesting that they had been transformed and reactivated after excretion (Alcock et al., 1999).

In Germany, raw sewage was found to contain 0.015 mg.l-1 of 17b-estradiol and 0.027 mg.l-1 of oestrone and it was also found that oestrone and 17a -ethinyloestradiol were not efficiently removed during wastewater treatment (Ternes, 1999a). In contrast, 17b-oestradiol and 16a - hydroxy-oestrone were eliminated with a higher efficiency: around 64-68% (Ternes, 1999a) (See Figure d.4).

Figure d.4: Elimination percentages and loads of estrogens during passage through a municipal sewage treatment plant located near Frankfurt/Maine over 6 days [Ternes, 1999a].

In discharges from the treatment plants, all compounds could be detected in the ngl-1 range (Ternes, 1999a), however oestrone was predominant with concentrations up to 0.07 mgl-1

158 Section 6. Case Studies

and a median value of 0.009 mgl-1. The compounds 17a -ethinyloestradiol and 16a -hydroxy- estrone were found at the detection limit of 0.001 mg l-1 (Ternes, 1999a). Oestrone was the only compound detected in 3 of the 15 rivers sampled at concentrations between 0.7 and 1.6 ng l-1 (Ternes, 1999a). Therefore, it seems that these compounds, particularly natural oestrogens, are not degraded in the treatment system and tend to accumulate in sludge and effluent. However, the loads entering receiving waters are quite low. DEHP is also an endocrine-disrupting compound. In Sweden, it has been found in all sewage sludge samples analysed, with concentrations between 25-660 mg kg-1 dry weight (Alcock et al., 1999).

In Italy, drinking water, rivers, and sediments have been analysed to determine the extent of environmental contamination by pharmaceuticals (see Table d.4), (Zuccato, 2000).

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

-1 -1 -1 DRUG DRINKING WATER (ngl ) RIVER WATER (ngl ) RIVER SEDIMENTS (ng kg ) Po (Piacenze Po Lambro Adda Lambro Adda Milan Lodi* Varese and, (Piacenze, (Milan)* (Sondrio) (Milan) (Sondrio) Cremona)* Cremona)* Atenolol

It can be seen that most drugs were measurable in these media, showing widespread contamination. The concentrations measured could potentially give rise to human exposure in the ng day-1 range, which is 3-4 orders of magnitude lower than the concentrations capable of producing pharmacological effects (Zuccato, 2000). Hence, acute exposure is assumed to be unlikely, but long-term effects must still be studied.

In Germany, occurrence of drugs in WWTS and rivers has been studied (Ternes, 1998c). Results showed maximum values in average loads of up to 3kg.day-1 for salicylic acid in the influent and up to 114g.day-1 for in the effluent (Ternes, 1998c).

159 Section 6. Case Studies

Figure d.5: Elimination of different drugs during passage through a municipal sewage treatment plant near Frankfurt/Maine over 6 days [Ternes, 1998c].

From Figure d.5, it appears that more than 60% of the compounds in the influent were usually removed during treatment of wastewater: ranging between 7-99% removal (Ternes, 1998c). Only carbamazepine, clofibric acid, and phenanzone were less efficiently eliminated (Figure d.6). However, complete elimination was not usually achieved; thus receiving waters may potentially be contaminated. Subsequently, a screening programme of 49 different German WWTS effluents was carried out, in addition to river sampling. Lipid regulating agents were found in the majority of the WWTS effluents, and in many river samples but in a much lower concentration (Ternes, 1998c). Polar metabolites of the compounds were usually detected. For example, clofibric acid was detected at levels up to 1.6 mg.l-1 in WWTS effluent and in the ng.l-1 range in rivers, which illustrates the importance of metabolites (Ternes, 1998c). Anti-inflammatories , such as, ibuprofen and naproxen, were also detected. Diclofenac was present in the highest concentration at median levels of 0.81 mg.l-1 in treated effluent effluent and 0.15 mg.l-1 in rivers (Ternes, 1998c). In the case of betablockers, the highest median concentration was found for metoprolol at 0.73 mg.l-1 in WWTS effluent and 0.45 mg.l-1 in rivers (Ternes, 1998c). b2-sympathomimetics were also present but in very low concentrations. Anti-cancer agents such as cyclophosphamide and ifosamide were detected at levels of 0.02 mg.l-1 and 0.08 mg.l-1 respectively in WWTS effluent; however, they are associated with presence of hospital effluents, and are not widespread (Ternes, 1998c). Carbamazepine, an anti-epileptic drug was widespread in the aquatic environment, with a high median value of 2.1 mg.l-1 in effluent and 0.25 mg.l-1 in rivers (Ternes, 1998c). Annual prescriptions of carbamazepine amount to approximately 80 tonnes per year in Germany, it then becomes metabolised and glucuronides are excreted. However treatment of wastewater cleaves these metabolites back to the parent compound, increasing the environmental concentrations (Ternes, 1998c).

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

160 Section 6. Case Studies

Ground water pollution has been detected in some instances, mainly due to leaching from landfill sites containing pharmaceutical wastes (Halling-Sørensen, 1998). In Berlin, clofibric acid has been detected in drinking water at concentrations between 10 ng.l-1 and 165 ng.l-1 and in all surface water samples around Berlin, suggesting extensive contamination (Halling- Sørensen, 1998).

River water is often polluted with pharmaceutical compounds. Most groups of compounds, i.e. antibiotics, antineoplastic agents, and ethinyloestradiol, have been detected between 5- 10 ng.l-1. A study conducted by Richardson and Bowron (1985), investigated the exposure of human pharmaceuticals in the river Lea in England and found that over 170 substances are used in excess of 1 tonne per year in the river’s catchment. This allowed them to predict a concentration of at least 0.1mg.l-1 in the river water (Halling-Sørensen, 1998).

161 Section 6. Case Studies

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

Compound Use/Origin Environmental occurrence Acetaminophen Analgesic/anti- Removed efficiently by WWTS, max. effluent 6mgl- inflammatory 1, not detected in surface waters Acetylsalicylic acid Analgesic/anti- Ubiquitous, removal efficiency 81%, max. effluent inflammatory 1.5mgl-1, in surface water 0.34mgl-1. Betaxolol betablocker Max. effluent 0.19mgl-1, in surface water 0.028mgl- 1. Benzafibrate Lipid regulator Removal efficiency 83%, max. effluent 4.6mgl-1, in surface water 3.1mgl-1. Biphenylol Antiseptic, fungicide Extensive removal in WWTS. Bisoprolol betablocker Max. effluent 0.37mgl-1, in surface water 2.9mgl-1. Carazolol Betablocker Max. effluent 0.12mgl-1, in surface water 0.11mgl- 1. Carbamazepine Analgesic, anti- Removal efficiency 7%, max. effluent 6.3mgl-1, in epileptic surface waters 1.1mgl-1. Chloroxylenol Antiseptic In influents and effluents <0.1mgl-1. Chlorophene Antiseptic Influent 0.71mgl-1, removal not very efficient

Clenbuterol b2-sympathomometic Max. effluent 0.08mgl-1, in surface waters 0.05mgl- 1. Clofibrate Lipid regulator River water 40ngl-1, not detected in effluent or surface waters. Clofibric acid Metabolite of Removal efficiency 51%, max. effluent 1.6mgl-1, clofibrate surface waters 0.55mgl-1, up to 270ngl-1 in German tap waters Cyclophosphamide antineoplastic Max. effluent 0.02mgl-1, not detected in surface waters, high in hospital sewage: up to 146ngl-1 Diatrizoate X-ray contrast media Resistant to biodegradation, median in German surface waters 0.23mgl-1, locally very high concentrations can occur. Diazepam Psychiatric drug Max. effluent 0.04mgl-1, not detected in surface waters. Diclofenac-Na Analgesic/anti- Removal efficiency 69%, max. effluent 2.1mgl-1, in inflammatory surface waters 1.2mgl-1. Dimethylaminophen Analgesic/anti- Removal efficiency 38%, max. effluent 1mgl-1, in azone inflammatory surface waters 0.34mgl-1. 17a -ethinylestradiol Oral contraceptive Up to 7ng.l in WWTS effluent, not detected in German surface waters above 0.5ngl-1. Etofibrate Lipid regulator Not detected in WWTS effluent and surface waters Sympathomimetic No studies but is known to be an endocrine- amine disrupting substance Fenofibrate Lipid regulator Efficiently removed, max. effluent 0.03mgl-1, not detected in surface waters. Fenofibric acid Metabolite of Removal efficiency 64%, max. effluent 1.2mgl-1, in fenofibrate surface waters 0.28mgl-1. Fenoprofen Analgesic/anti- Not detected in WWTS effluent or surface waters inflammatory Fenoterol b2-sympathomometic Max. effluent 0.06mgl-1, in surface waters 0.061mgl-1. Flurorquinolone Antibiotics Ubiquitous, led to resistance in pathogenic carboxylic acids bacteria, strongly sorbs to soil. No studies Antidepressant No studies

162 Section 6. Case Studies

Gemfibrozil Lipid regulator Removal efficiency 69%, max. effluent 1.5mgl-1, in surface waters 0.51mgl-1. Gentisic acid Metabolite of Efficiently removed by WWTS, max. effluent acetylsalicylic acid 0.59mgl-1, in surface waters 1.2mgl-1. o-hydroxyhippuric Metabolite of Efficiently removed by WWTS, not detected in acid acetylsalicylic acid effluent or surface waters. Ibuprofen Analgesic/anti- Removal efficiency 90%, max. effluent 3.4mgl-1, in inflammatory surface waters 0.53mgl-1. Ifosamide Antineoplastic Max. effluent 2.9mgl-1, not detected in surface waters, hospital sewage 24ngl-1, totally refractory to removal by WWTS. Indomethacine Analgesic/anti- Removal efficiency 75%, max. effluent 0.60mgl-1, inflammatory in surface waters 0.2mgl-1. Iohexol X-ray contrast media Very low aquatic toxicity. Iopamidol X-ray contrast media Max. effluent 15mgl-1, median 0.49mgl-1. Iopromide X-ray contrast media Resistant to biodegradation, yields refractory, unidentified metabolites, max. effluent 11mgl-1. Iotrolan X-ray contrast media Very low aquatic toxicity. Ketoprofen Analgesic/anti- Max. effluent 0.38mgl-1, in surface waters 0.12mgl- inflammatory 1. Analgesic/anti- Not detected in WWTS effluent or surface waters. inflammatory Metoprolol betablocker Removal efficiency 83%, max. effluent 2.2mgl-1, in surface waters 2.2mgl-1. Nadolol Betablocker Max. effluent 0.06mgl-1, not detected in surface waters. Naproxen Analgesic/anti- Removal efficiency 66%, max. effluent 0.52mgl-1, inflammatory in surface waters 0.39mgl-1. antidepressant No studies Phenazone Analgesic Removal efficiency 33%, max. effluent 0.41mgl-1, in surface waters 0.95mgl-1. Betablocker Removal efficiency 96%, max. effluent 0.29mgl-1, in surface waters 0.59mgl-1. Propyphenazone Analgesic/anti- Prevalent in Berlin waters. inflammatory Salbutamol b2-sympathomometic Max. WWTS influent 0.17mgl-1, in surface waters albuterol 0.035mgl-1. Salicylic acid Metabolite of Up to 54mgl-1 in WWTS effluent but efficiently acetylsalicylic acid removed in effluent, average in effluent 0.5mgl-1, in surface waters 4.1mgl-1. Sulfonamides Antibiotics Present inn landfill Terbutaline b2-sympathomometic Max. effluent 0.12mgl-1, not detected in surface waters. 3,4,5,6-tetrabromo- Antiseptic, fungicide Found in influents and effluents in Germany o-cresol <0.1mgl-1. Timolol Betablocker Max. effluent 0.07mgl-1, in surface waters 0.01mgl- 1. Analgesic/anti- Not detected in WWTS effluent or surface waters. inflammatory Antiseptic 0.05-0.15mgl-1 in water, very widely used. Cardiac drug No occurrence data

163 Section 6. Case Studies

Table d.6 summarises the occurrence of certain pharmaceuticals in the environment. It can be seen that some pharmaceuticals, such as lipid regulators, X-ray contrast media, antibiotics etc., are ubiquitous and extremely persistent in the environment, some are even present in drinking water: clofibric acid for example has been found at concentrations up to 0.27 mg.l-1 in some German waters. This would breach EC regulations if the compounds were classed as pesticides (ENDS, 2000). However only a fraction of the drugs on the market, have been investigated regarding their occurrence in the environment.

The fate of pharmaceutical compounds

Pharmaceutical compounds enter the body and are then often metabolised in the liver through oxidation, reduction, or hydrolysis to “phase I” metabolites, which tend to be more toxic than the parent compound. Other reactions, such as conjugation, metabolise the compounds into “phase II” metabolites, which tend to be inactive and more polar and water- soluble. It has also been observed that phase II metabolites are often reactivated into the parent compounds, either during treatment of wastewater and sewage sludge or in the environment. For example, chloramphenicol glucoronide and N-4-acetylated sulphadimidine (phase II metabolites of the antibiotics chloramphenicol and sulphadimidine, respectively), are reactivated in liquid manure (Halling-Sorensen, 1998). This shows the importance of the investigation of metabolites as well as parent compounds.

For example, 17b-oestradiol, is administered orally and mainly undergoes first-pass hepatic metabolism, being transformed to oestrone and oestriol, which are less potent (Christensen, 1998). Other metabolites are also formed but to a lesser extent. Experiments using the diluted slurry of activated sludge from a WWTS, were undertaken to investigate the persistence of natural oestrogens and contraceptives under aerobic conditions (Ternes, 1999b). The natural oestrogen 17b-oestradiol, was oxidised to oestrone, which is then linearly removed with time. Rapid elimination also occurred for16a -hydroxy-oestrone. However, the contraceptive 17a -ethinyloestradiol was persistent and highly stable under environmental conditions (Ternes, 1999b). Two glucuronides of 17b-oestradiol were cleaved to their parent compounds and 17b-oestradiol was re-released in an activated form (Ternes, 1999b). This indicates that the microorganisms present have the ability to deconjugate oestrogen glucuronides. It is interesting to note that glucuronide conjugates are the main oestrogen metabolites excreted by humans, so during wastewater treatment, the concentration of free oestrogen increases due to the of the glucuronide moieties from the compounds. As a result, the predominant presence of oestrone in WWTS effluents and rivers is due to; its high stability during treatment; the cleavage of glucuronide conjugates from oestrone and 17b-oestradiol; and the oxidation of the latter to oestrone (Ternes, 1999b).

Penicillin antibiotics are eliminated rapidly and have short half-lives in the body, usually 30- 60 minutes, and very high concentrations are excreted in urine: it has been determined that up to 40% of penicillin V is excreted unchanged (Christensen, 1998).

Cyclophosphamide, an anti-cancer drug is administered intravenously or orally. It is not active in itself but undergoes activation in the body when transformed to phosphoramide mustard and acrolein. The parent compound is genotoxic. Some of it is excreted unchanged: 5-20% (Christensen, 1998).

Antibiotics are generally believed to leave humans unchanged by the body metabolism (see Table d.7) (Hirsch, 1999) and it has been determined that up to 90% of the parent compounds are excreted unchanged (ENDS, 2000). These active products can be excreted either as unchanged compounds or as conjugates; 30-90% of administered antibiotics are excreted via urine as active substances (Alcock et al., 1999). This introduces the problem at

164 Section 6. Case Studies the WWTS of disruption of biological treatment processes, as pharmaceutical compounds, particularly antibiotics, can potentially affect bacteria.

Table d.7: Human prescription amounts and excretion rates of antibiotics [Hirsch, 1999].

Antibiotic Amount prescribed Excretion (%) (t/a) Unchanged Other Metabolites Amoxicillin 25.5-127.5 80-90 10-20 Ampicillin 1.8-3.6 30-60 20-30 Penicillin V 40 40 60 Penicillin G 1.8-3.6 50-70 30-50 Sulphamethoxasole 16.6-76 15 Trimethoprim 3.3-15 60 Erythromycin 3.9-19.8 >60 Roxithromycin 3.1-6.2 >60 Clarithromycin 1.3-2.6 >60 Minocycline 0.8-1.6 60 40 Doxycycline 8-16 >70

A study looking at the amounts of antibiotics in human faeces found trimethoprim and doxycycline at concentrations between 3-40 mg.kg-1, and erythromycin at concentrations around 200-300 mg.kg-1 (Hirsch, 1999). Elimination at treatment plants is usually incomplete, ranging between 60-90% (Ternes, 1998b). Polar antibiotics are probably not removed efficiently because elimination is mainly due to adsorption on activated sludge, which is mediated through hydrophobic interactions (Hirsch, 1999). As a consequence, receiving waters and other environmental media may become contaminated. Furthermore, erythromycin and other drugs such as naproxen and sulphasalazine, have survived in the environment for over a year (Zuccato, 2000). Clofibric acid was also found to survive for 21 years and although its use has been stopped, it is still detected in rivers and lakes in Italy (Zuccato, 2000).

Many pharmaceutical compounds have the same physico-chemical characteristics as organic compounds, such as persistence and lipophilicity; much less is known though about their entry into the environment and their subsequent fate (Alcock et al., 1999). Over 30% of all drugs produced between 1992 and 1995 were lipophilic, i.e. solubility less than 100 mg.l-1 (Halling-Sørensen, 1998). The fate of pharmaceutical substances may be divided into three groups:

· Mineralisation to CO2 and water, for example aspirin. · Retained in sludge, if the compound is lipophilic and not readily biodegradable. · Emitted to receiving water due to transformation into a more hydrophilic form but still persistent, for example clofibrate.

Richardson and Bowron (1985) investigated degradation of pharmaceuticals during wastewater treatment and found that many common compounds are biodegradable, although cortisteroid compounds and ethinyloestradiol, among others, were non- biodegradable (Alcock et al., 1999).

In work by Kummerer (2000), two clinically important groups of antibiotics have been studied with regards to their biodegradability. Chinolones and nitromidazoles possess different chemical structures, actvity spectra and modes of action. The study found low rates of biodegradation for , , and metronidazole; it also found that the genotoxicity 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 treatment and the results are varied. Compounds such as the analgesics, ibuprofen and naxoproxen, have been found to have removal efficiencies between 22-90% and 15-78% respectively (ENDS, 2000). It has also been determined that 70-80% of the drugs administered in fish farms, are transferred into the environment (Halling-Sorensen, 1998).

Table d.6 gives an overview of the present knowledge on the environmental fate of specific pharmaceuticals. It can be seen that most hormones, such as oestrogen, are persistent in all areas and that most of the antibiotics used for human treatment are not biodegradable. The majority of other compounds used for human treatment are also non-biodegradable, with the exception of the following: paracetamol, , ibuprofen, caffeine, and aspirin. The compounds used in veterinary treatment tend to be more biodegradable than human pharmaceuticals, although the speed of degradation will depend on environmental conditions, such as pH, and temperature.

It has also been determined that iodinated X-ray contrast media are not degraded during wastewater treatment, due to their high polarity (log Kow of iopromide = -2.33) (Ternes, 2000). These compounds are designed to be highly stable to give optimum results during X-ray, so are not readily biodegradable. Ninety percent of X-ray contrast media are excreted unmetabolised (Ternes, 2000); hence, receiving waters will also tend to be contaminated. Concentrations up to 0.49 mg.l-1 for iopamidol were detected in receiving rivers (Ternes, 2000). It appears that groundwater may also become contaminated, as concentrations of up to 2.4 mg.l-1 were identified for iopamidol as a result of infiltration by polluted surface water (Ternes, 2000). The concentration of X-ray contrast media in receiving water bodies is lower than that detected in wastewater effluent; however, due to the high persistence of such media in the environment, this reduction seems to be less important than for other pharmaceutical compounds (Ternes, 2000). This shows that pharmaceutical compounds have the ability to infiltrate and survive for many years. , clofibric acid, benzafibrate, diclofenac, and carbamezepine, have all been found in aquatic environments, persisting there for up to 20 years (Ternes, 2000).

With the exception of Denmark, there is very little data available on the use and quantities of pharmaceuticals, which renders the task of studying the fate of these compounds in wastewater very difficult. The list of pharmaceutical substances could be exhaustive and prioritisation is necessary.

Certain physical processes occur that may be used to degrade these contaminants: sorption to solids, volatilisation, chemical degradation, and biodegradation. The effectiveness of sorption and volatilisation can be determined using the octanol water partition coefficient (Kow) and Henry's law constant (Hc) [Rogers, 1996]:

· if log Kow is less than 2.5, the compound has a low sorption potential (i.e. it will not adsorb onto soil particles and will not be very lipophilic), · if log Kow is between 2.5 and 4, the compound has a medium sorption potential, · if log Kow is greater than 4, the compound has a high sorption potential and is very lipophilic. -4 -9 · if Hc is greater than 1x10 and Hc/Kow is greater than 1x10 , the compound is thought to have a high volatilisation potential, -4 -9 · if Hc is less than1x10 and Hc/Kow is less than 1x10 the compound is thought to have a low volatilisation potential.

166 Section 6. Case Studies

Several models, using Kow, Koc, and Hc, have been developed in order to take all these characteristics into account in order to prioritise pollutants. This has enabled an assessment to be made of the exposure risk to such pollutants, through consuming food derived from sludge-amended soil. Two parameters are important when trying to determine movement of contaminants through the food chain: persistence, and non-polarity. Easily metabolised or polar compounds do not move through the food chain. As bioconcentration increases with lipophilicity, compounds with high log Kow values will tend to accumulate in the food chain (Duarte-Davidson, 1996).

Unfortunately, sewage sludge may contain a wide variety of pharmaceutical compounds, and for many, information on their characteristics is not readily available. Therefore, it becomes difficult to eliminate or prioritise pharmaceuticals using this screening process. In addition, without proper information on the physico-chemical properties of these compounds, it is not possible to predict their fate in the environment, or their concentrations in sewage sludge (Alcock et al., 1999). In the absence of detailed knowledge, it can be presumed that most pharmaceutical substances have the same properties as pesticides and other organic pollutants (Halling-Sorensen, 1998). Also, as mentioned already, many of these compounds have lipophilic characteristics, so they are likely to accumulate. Some pharmaceuticals may even be metabolised during treatment or in the soil, to more readily available compounds, increasing their potential for plant and animal uptake (Engwall et al., 2000).

The low concentration of individual pharmaceutical compounds, coupled with their metabolic characteristics to incomplete removal in WWTS (Daughton, 1999). They tend to be non-volatile, so transport and movement through the environment will occur via aquatic media. In fact, their polarity and non-volatile characteristics will often prevent them from leaving the aquatic environment (Daughton, 1999). As it has been seen earlier, metabolites also tend to be cleaved to the parent compound during wastewater treatment and then released afterwards.

Nutraceuticals/Herbal Remedies

During the last several years, the popularity of nutritional supplements was codified by the creation of a new term for the subclass of highly bioactive food supplements called nutraceuticals (Daughton and Ternes, 1999) also referred to as nutriceuticals. Nutraceuticals are a rapidly growing commercial class of bioactive compounds, usually botanicals, intended as supplements to the diet. Nutraceuticals and many herbal remedies can have potent physiologic effects. These are a mainstay of alternative medicine and have enjoyed explosive growth in use in the United States and Europe during the last decade. Many are used as food supplements that have either proven or hypothesized biologic activity but are not classified as drugs by the FDA, primarily because a given botanical usually has not one but an array of distinct compounds whose assemblage elicits the putative effect and because these arrays cannot be easily standardized. As such, they are not regulated and are available over the counter (heavily promoted via the Internet). Even in those cases in which the natural product is identical to a prescription pharmaceutical (e.g., the Chinese red- yeast product Cholestin newly introduced to the United States contains lovastatin, an active ingredient in the approved prescription drug Mevacor used to lower cholesterol levels), a recent ruling (Borman , 1998; Zeissel, 1999). prevented the FDA from regulation.

The significance of dietary supplements in the United States led to the creation of the Office of Dietary Supplements (ODS) via the DSHEA in 1995 under the National Institutes of Health (NIH) (DSHEA, 1994). The ODS maintains a searchable database (International Bibliographic Information on Dietary Supplements [IBIDS]) of published scientific literature on dietary supplements (NIH Office of Dietary Supplements, 1999).

167 Section 6. Case Studies

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

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

Another example is , which is prepared from the of methysticum, used of its mild effect among other effects. The active ingredients in Kava are believed to be a suite of lipophilic comprising substituted -pyrones (, , , and others) (Shao, et.al.,1998). These compounds display a host of effects in humans, but nothing is known about their effects on other organisms or fate in WWTS.

There are many nutraceuticals, both new and ‘traditional’, experiencing increased consumption. These few examples illustrate the unknowns regarding whether these compounds are being excreted, surviving WWT, and then having possible effects on aquatic organisms. Nutraceuticals and herbal remedies would have the same potential fate in the environment as pharmaceuticals, with the added dimension that their usage rates could be much higher, as they are readily available and taken without the controls of prescription medication. However, because these compounds are natural products, they would be expected to biodegrade more easily .

Legislation and policy for risk assessment

At the beginning of the 1980s, environmental risk assessment was introduced for new chemicals but it took a decade later for drugs to be included in the discussion. In Europe, since the 1990's, there has been a distinction made between compounds for human use and those used in veterinary practice. For several years legislation has been implemented for veterinary medicines. The EU Directive 81/852/EEC, (Amended 1993), introduced the requirement for a tiered environmental risk assessment of new veterinary products, and attempts are being made to implement this for a review of existing substances (ENDS, 2000). Currently, environmental risk assessment consists of examining the likely environmental sectors and if levels of pharmaceutical compounds exceed the trigger values set, such as 100 mg.l-1 in manure, further data is required (ENDS, 2000). The technical directive [Directive 81/852/EEC, amended 1993] concerning veterinary medical products outlines the basic requirements for conducting an environmental risk assessment (Halling- Sorensen, 1998). The technical directive [Directive 75/318/EEC, amended 1993] concerning human medical products does not refer to any ecotoxicology or ecotoxicity tests and no guidance is given on how to carry out an environmental risk assessment for drugs used by humans. However, a draft directive for human pharmaceuticals is currently being devised, proposing that risk assessment should be part of the approved procedure of new medical substances (Halling-Sorensen, 1998). The EC is proposing a similar programme for human

168 Section 6. Case Studies medicines: if drug concentrations in surface waters are predicted to exceed 0.01 mg.l-1, toxicity testing is required to find the no effect level (NOEC) (ENDS, 2000). Although environmental assessment of the potential impacts of newly developed drugs has been expected in the EU since the 1st of January 1995 (Christensen, 1998), it is should be noted that the end point of human exposure is not usually investigated. Also, assessment of individual compounds is usually based on a limited number of tests but pharmaceuticals in the environment may affect a large number of different organisms and species, so this ability should be reflected in the tests carried out (Stuer-Laurisden, 2000). Pharmaceuticals may not affect the standard test species and give rise to false negative results (Stuer-Laurisden, 2000).

A risk assessment study was carried out in Germany looking at salicylic acid, paracetamol, clofibrinic acid, and methotrexate (Henschel, 1997). As seen previously in this Case Study, these compounds were present in the environment and had toxic effects in at least one standard ecotoxicological test. The most sensitive reaction however, was to a non-standard test incorporating relevant end points for the pharmaceuticals (Henschel, 1977), so proving the limitations of standard tests.

Risk assessment of pharmaceuticals

Risk assessment for pharmaceuticals in the environment has not usually been carried out due to the lack of data and the need for more precise and sensitive measures in the environmental sectors. Some high consumption compounds, such as antibiotics and clofibric acid, are being released into the environment and have been found to be widely present in aquatic environments, sediments and soils. Although the liver often metabolises pharmaceutical compounds to more easily hydrolysed compounds, the metabolites can be cleaved back to the more hydrophobic, active parent compounds by bacteria, which can then persist and bioaccumulate.

There is a strong opinion that there are more pressing environmental problems than pharmaceuticals and that these compounds do not pose a large risk because they are present in such low concentrations (ng.l-1), with most effects only seen in the mg.l-1 range (ENDS, 2000). As already stated though, disease resistance to pharmaceuticals is favoured by low concentration exposure and compounds such as antineoplastic agents and hormones have effects at very low levels. The effects of active compounds in the low, ng.l-1 range, cannot be excluded, as experience with pesticides shows, impacts can be significant at low levels (Stuer-Laurisden, 2000).

At the moment, most toxicity tests performed investigate acute impacts on specific species. However, as most compounds in question are persistent and are discharged continuously into the environment at low levels, it would appear to be more relevant to look at the chronic, long-term impacts of exposure to low concentrations over all trophic levels. Some of the long-term, chronic impacts that may be of concern are genotoxicity and reproduction impairment. It has been found that Daphnia are tolerant to most antibiotics within the mg.l-1 range but that exposure to these levels over several weeks causes death, probably because of toxic effects in the food organisms (ENDS, 2000). The effects of continuous exposure to even low levels of pharmaceuticals in the environment are very complex and affect many different organisms. More studies are necessary with regards to the long-term impacts and the potential synergistic effects of exposure to a mixture of drugs.

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

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A study carried out in Italy (Zuccato, 2000), found that most drugs were present in drinking water, river water and sediments, but that human exposure would only be in the ng day-1 range, which is much lower that the concentration where effects are expected to be observed. The study concluded that risk seems negligible but possible long-term exposures and impacts still need investigation. The same was found with X-ray contrast media, which were ubiquitous in the aquatic environment and highly persistent. No acute toxic effects were observed, in Daphnia magna or in bacteria, algae, fish, and crustaceans (Ternes, 2000). Long-term exposure was not investigated.

Risk assessments were carried out on three pharmaceuticals, using the computer program EUSES, which was developed as a support to the technical guidance document for risk assessment on new and existing substances (Christensen, 1998). The three compounds assessed were; the synthetic estrogen, 17a -ethinylestradiol; the antibiotic, penicillin V; and the antineoplastic agent, cyclophosphamide. This program estimated environmental fate and human exposure based on worst-case scenarios, using data on the physico-chemical properties of the compounds and amounts consumed. The results indicated that for all three there was negligible human risk. However, the author stressed the point that many uncertainties are associated with this method and that the drugs, although seemingly insignificant, still contribute to the total toxic load in the environment, and that interactions may have ecotoxicological impacts.

Research in Denmark attempted to carry out a risk assessment for the 25 most highly used drugs in the primary health sector in Denmark, including furosemide, paracetamol, ibuprofen, and estradiol (Stuer-Laurisden et al., 2000). Different parameters are used to calculate environmental exposure: biodegradation, bioaccumulation, and bioavailability are three of the most important. Nevertheless, it has been seen that biodegradation of pharmaceuticals does not happen very often. Bioaccumulation in the human body does not happen for drugs, as they are metabolised to more polar compounds that can then easily be excreted. The bioavailability of drugs is different if it is bound to solids, adsorbed, or dissolved. However, as seen above, the octanol-water partition coefficient can be used to determine bioaccumulation, and other parameters can be used to determine bioavailability. However, this information on the properties of the compounds is not readily available. They found that ecotoxicology data was available only for 6 of the 25 compounds, and biodegradation data only for 5 (Tables d.9 and d.10) (Stuer-Laurisden et al., 2000). Predicted environmental concentrations (PEC) should be determined for the system where it is anticipated that the highest values would be found, i.e. aquatic ecosystems and sewage sludge (Jorgensen, 2000). In order to do so, modelling could be a useful tool; however, not many models have yet been developed and validated due to the lack of data in this area (Jorgensen, 2000 and Halling-Sorensøn,1998). The predicted environmental concentrations for the 6 compounds were calculated and all exceeded 0.001mg.l-1, which is the cut off value in EU legislation for carrying out more investigations. The majority of the PECs are between 1-100ng.l-1, it is only for the top 5 compounds that these reach the mg.l-1 range (Stuer- Laurisden, 2000). The predicted no effect concentration (PNEC) is based on ecotoxicological data and the PEC/PNEC ratio was found to exceed one only for ibuprofen, paracetamol, and acetylsalicylic acid and below one for estrogen, diazepam, and digoxin (Stuer-Laurisden, 2000). This showed that data is only partially available, preventing complete risk assessments. Nevertheless, it was concluded, with this data, that ibuprofen, aspirin, and paracetamol may pose a risk; hence, contradicting other studied that had concluded these were efficiently removed by treatment and did nor reach hazardous levels in the environment (ENDS, 2000). The efforts are hampered by the fact that concentrations measured in sludge and effluent vary extensively, and furthermore comparisons of predicted concentrations in sludge based on Kow, sludge-water partition coefficients (Kd), or acid-base constants (pKa) also reveal large variations (Stuer-Laurisden, 2000).

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The cost/benefit stage of risk assessment is extremely important: the indirect and direct effects of the drug on the environment and the human body must be known in order to make an informed decision (Jorgensen, 2000 and Halling-Sørenson,1998). This can allow selection of a substitute drug that has the same benefits but fewer environmental impacts. Hence, lack of data for toxicity but also environmental concentrations prevent calculation of risks.

Examples of good environmental practice

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

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

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

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

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

171 Section 6. Case Studies multi-stage biofilm process removing all organic matter and the toxicity is no longer measurable (Rosen, et.al. 1998).

Conclusion and Recommendations

Pharmaceutical compounds must become priority substances in the same way as persistent organic pollutants are. Judgements on the relative priorities are based on the knowledge at the time and the priority list will obviously change over time as more studies are carried out and more data is gathered. Pharmaceuticals are widely used and mainly disposed of through the system, allowing their entry into the environment continually, as removal rates can be compensated by replacement rates (Daughton and Ternes, 1999). They are concerning because they are biologically active and are usually lipophilic and potentially bioaccumulating. Many resist biodegradation, within the WWTS and in the environment, and can end up in surface and ground waters, as well as sediment and soils and are found to be highly stable under environmental conditions. Furthermore, metabolites tend to be cleaved and transformed back to the parent compounds once in the environment, increasing the concentrations and justifying the importance of metabolites. Many can have unpredicted and unknown side effects particularly after long-term exposure to low concentrations. Aquatic ecosystems are the most vulnerable, as this is the main environmental compartment where pharmaceuticals are found ubiquitously.

Analytical methods must be improved to detect pharmaceuticals at very low levels and sampling procedures must be of very high quality so as not to cause contamination. More information on the physicochemical, ecotoxicity, and ecotoxicological characteristics, using appropriate tests that better accommodate subtle end points, of drugs and their metabolites should be obtained in order to allow environmental risk assessments to be carried out. Furthermore, this may lead to the validation of modelling techniques that could speed up the whole process. Furthermore, use patterns of drugs in all countries is still very limited and should be determined, as it is essential for the elaboration of amounts released into the environment.

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

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

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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 having intended end uses, primarily on the human body (products not intended for ingestion) or in the household. In general, these chemicals alter odour, appearance, touch, or taste without displaying significant biochemical activity (Daughton and Ternes, 1999). Most of these chemicals are used as the active ingredients or preservatives in cosmetics, toiletries and fragrances. They are not used for treatment of disease, but some may be intended to prevent diseases (e.g., sunscreen agents). In contrast to drugs, almost no attention has been given to the environmental fate or effects of personal care products, the focus has traditionally been on the effects from intended use on human health.

Personal care products differ from pharmaceuticals in that large amounts can be directly introduced to the environment and unlike medicinal compounds, there are rarely recommended doses. These products can be released directly into recreational waters or volatilised into the air. Because of this direct release they can bypass possible degradation in UWWT. Also, in contrast to pharmaceuticals, less is known about the effects of this broad and diverse class of chemicals on non-target organisms, such as aquatic organisms. Data are also limited on the potential adverse effects on humans. For example, common sunscreen ingredients, 2-phenylbenzimidazole-5-sulfonic acid and 2-phenylbenzimidazole, can cause DNA breakage when exposed to UV-B (Stevenson and Davies, 1999).

The quantities of personal care products produced commercially can be very large. For example, in Germany alone the annual output was estimated to be 559,000 tonnes for 1993 (Statistisches Bundesamt, 1993). A few examples are given below of common personal care products that are ubiquitous pollutants, which may possess varying degrees of bioactivity. Table e.1 Personal care and fragrances produced in Germany (1993)

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

· Preservatives

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

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· Disinfectants/Antiseptics

Triclosan, a chlorinated diphenyl ether: 2,4,4´-trichloro-2´-hydroxydiphenyl ether, is an antiseptic agent that has been widely used for almost 30 years in a vast array of consumer products. Its use as a preservative and disinfectant continues to grow; for example, it is incorporated at < 1% in Colgate's “Total” toothpaste, the first toothpaste approved by the FDA to fight gingivitis. While triclosan is registered with the U.S. EPA as a pesticide, it is freely available over the counter. Triclosan's use in commercial products includes footwear (in hosiery and insoles of shoes called Odour-Eaters), hospital hand-soap, acne creams (e.g., Clearasil), and rather recently as a slow-release product called Microban, which is incorporated into a wide variety of plastic products from children's toys to kitchen utensils such as cutting boards. Many of these uses can result in direct discharge of triclosan to UWW collecting systems, and as such this compound can find its way into receiving waters depending on its resistance to microbial degradation. Okumura and Nishikawa [1996] found traces of triclosan ranging from 0.05 to 0.15 µg.l-1 in water. Although triclosan has long been regarded as a biocide (a toxicant having a wide-ranging, nonspecific mechanism(s) of action - in this case gross membrane disruption) McMurry et al. [1998] report that triclosan actually acts as an antibacterial, having particular enzymatic targets (lipid synthesis). As such, bacteria could develop resistance to triclosan. As with all antibiotics in the environment, this could lead to development of resistance and change in microbial community structure (diversity).

A wide range of disinfectants are used in large amounts, not just by hospitals but also by households and livestock breeders. These compounds are often substituted phenolics as well as other substances, such as triclosan. Biphenylol, 4-chlorocresol, chlorophene, bromophene, 4-chloroxylenol, and tetrabromo-o-cresol [Ternes et al 1998] are some of the active ingredients, at percentage volumes of < 1-20%. A survey of 49 WWTPs in Germany [Ternes et al 1998] routinely found biphenylol and chlorophene in both influents, up to 2.6 µg/L for biphenylol and up to 0.71 µg.l-1 for chlorophene, and effluents. The removal of chlorophene from the effluent was less extensive than for biphenylol, with surface waters having concentrations similar to that of the effluents.

· Sunscreen Agents

The occurrence of sunscreen agents (UV filters) in the German lake Meerfelder Maar was investigated by Nagtegaal et al. [1997]. The combined concentrations of six sunscreen agents (SSAs) identified in perch (Perca fluviatilis) in the summer of 1991 were as high as 2.0 mg.kg-1 lipid and in roach (Rutilus rutilus L) in the summer of 1993, as high as 0.5 mg.kg- 1 lipid. Methylbenzylidene camphor (MBC) was detected in roach from three other German lakes. These lipophilic SSAs seem to occur widely in fish from small lakes used for recreational swimming. Both fish species had body burdens of SSA on par with PCBs and DDT. The bioaccumulation factor, calculated as quotient of the MBC concentration in the whole fish (21 µg.kg-1) versus that in the water (0.004 µg l-1), exceeded 5,200, indicating high lipophilicity. The fact that SSAs (e.g., 2-hydroxy-4-methoxybenzophenone [oxybenzone] and 2-ethylhexyl-4-methoxycinnamate) can be detected in human breast milk (16 and 417 ng.g-1 lipid, respectively) [Hany et al 1995] shows the potential for dermal absorption and bioconcentration in aquatic species. No data have been published on newer SSAs such as avobenzene (1-[4-(1,1-dimethylethyl)phenyl]-3(4-methoxyphenyl)-1,3-propanedione).

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

174 Section 6. Case Studies example in cosmetics, cleaning agents and air fresheners. Perfume ingredients may enter the urban wastewater system directly from domestic sources, such as in the washing agents or from being washed off skin in the case of perfumes and cosmetics.

The organic compounds found in perfumes that may be of environmental or health concern include

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

Details of the physical properties of these compounds are included in Appendix B. Health and environmental effects of the compounds discussed are also briefly introduced within this case study.

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

Synthetic musks comprise a series of structurally similar chemicals (which emulate the odour of the more expensive, natural product, from the Asian musk deer), used in a broad spectrum of fragranced consumer items, both as fragrance and as fixative. Included are the older, synthetic nitro musks (e.g., ambrette, musk ketone, musk xylene, and the lesser known musks moskene and tibetene) and a variety of newer, synthetic polycyclic musks that are best known by their individual trade names or acronyms.

The major musks used today are, the polycyclic musks (substituted indanes and tetralins), which account for nearly two-thirds of worldwide production and the inexpensive nitro musks (nitrated aromatics), accounting for about one-third of worldwide production. These substances are used in nearly every commercial fragrance formulation (cosmetics, detergents, toiletries) and most other personal care products with fragrance; they are also used as food additives and in cigarettes and fish baits (Gatermann, et.al. 1998)

The nitro-musks are under scrutiny in a number of countries because of their persistence and possible adverse environmental impacts and therefore are beginning to be phased out in some countries. Musk xylene has proved carcinogenic in a rodent bioassay and is significantly absorbed through human skin; from exposure to combined sources, a person could absorb 240 µg/day [Bronaugh et al 1998]. The human lipid concentration of various musks parallels that of other bioaccumulative pollutants, such as PCBs [Schmid 1996]. Worldwide production of synthetic musks in 1988 was 7000 tonnes [Gattermann et al 1998] and worldwide production for nitro musks in 1993 was 1,000 tonnes, two-thirds of which were musk xylene [Kfferlein 1998]

Synthetic musks first began to be identified in environmental samples almost 20 years ago [Yamagishi 1981 and 1983]. By 1981, Yamagishi et al. had identified musk xylene and musk ketone in gold fish (Carassius auratus langsdorfii) present in Japanese rivers and soon after [Yamagishi et al 1983] in river water, wastewater, marine mussels (Mytilus edulis), and oysters (Crassosterea gigas). This was followed by a number of studies in Europe, some of which are summarised in table e.2.

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Table e.2 Concentrations of various musk compounds in environmental samples. Location Tissue/Substance Product Concentration Reference North Germany Freshwater Fish Musk Xylene 10-350mg.kg-1 Geyer et al, 1994 (Fillet) Musk Ketone 10-380mg.kg-1 Geyer et al, 1994 Ruhr River, Bream and Perch Galaxolide, Average Eschke et al 1998 Germany (Fatty tissue) Tonalide and concentrations Celestolide between 2.5 and 4.6 mg.kg-1 (ppm) Berlin, Germany Surface waters Galaxolide, Maximum Herberer et al Tonalide and concentrations 1999 Celestolide above 10m L-1 Elbe River, Particulate matter Musk ketone 4-22 ng/g Winkler et al. 1998 Germany from river samples Galaxolide 148-736 ng/g Tonalide 194-770 ng/g Italy Freshwater Fish Galaxolide, 4 ng.g-1 – Draisci et al 1998 (Fatty tissue) Tonalide 1054 ng.g-1

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

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

Although the significance of the aquatic toxicity of the nitro and polycyclic musks is debatable (genotoxicity from the polycyclics seems not to be a concern) [Kevekordes 1998], the aminobenzene (reduced) versions of the nitro musks can be highly toxic. These reduced derivatives are undoubtedly created under the anaerobic conditions of sewage sludge digestion. Behecti et al. [1998] tested the acute toxicity of four reduced analogs of musk xylene on Daphnia magna. The p-aminodinitro compound exhibited the greatest toxicity of the four, with extremely low median effective concentration (EC50) values averaging 0.25 µg.l-1 (0.25 ppb).

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

176 Section 6. Case Studies

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

Amino nitro musk transformation products are · more water soluble than the parent musks, · still have significant octanol-water partition coefficients (high bioconcentration potential), · more toxic than the parent nitro musks, therefore more attention should be focused on these compounds.

Because synthetic musks are ubiquitous; used in large quantities; introduced into the environment almost exclusively via treated wastewater; and are persistent and bioconcentratable, they are prime candidates for monitoring in both water and biota as indicators for the presence of other personal care chemicals. Their analysis, especially in biota, has been thoroughly discussed by Gatermann et al. [1998] and by Rimkus et al. [1997].

It is thought that musk compounds can bioaccumulate in human tissue [spinnrad website 2000] and act as hormones, because they bind to the hormone receptors of the cells [Gerhard, I umweltmedizin website 2000]. However, there is insufficient data for an adequate toxicologically assessment for both the nitro- and the polycyclic musk scents [Antusch, 1999].

Emission Data

At the present time the quota of the polycyclic musk scents amounts to approximately 85 % of total musk production worldwide, and the quota of nitro-musk scents is approximately 12 % [Rebmann et al., 1998].

Musk Compounds in Wastewater: The use of musk compounds in cosmetic and detergent products, which are used primarily in domestic situations or in buildings connected to the UWW collecting system, implies that their presence in surface waters occurs via municipal wastewater treatment plants. The mean concentrations found in biologically clarified wastewater from 25 German municipal WWTP were,

o for musk-xylene: 0.12 µg.l-1 (concentration range: 0.03 – 0.31 µg.l-1) and o for musk-ketone: 0.63 µg.l-1 (concentration range: 0.22 – 1.3 µg.l-1).

The mean emission levels in Germany were quantified as 20 µg/inhabitant/day for musk- xylene and 90 µg/inhabitant/day for musk-ketone [Eschke et al., 1994].

In Vienna, Austria, extensive testing of wastewater was carried out at the pilot plant of the city’s main WWTP, during 1999. Concentrations of musk compounds in WWTP influent and effluent are shown in Table e.3.

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Table e.3: Musk compound concentrations in influent and effluent of the pilot WWTP Simmering, Austria in 1999 [Hohenblum et al., 2000].n: number of samples analysed. LOD: limit of detection Compound Type of Sample Range Mean (µg.l-1) sample number > (µg.l-1) Value LOD (µg.l-1) Musk-xylene Influent 4 0.023 - 0.037 0.031 (n=4) Effluent 0 - - Musk-ambrette Influent 0 - - (n=4) Effluent 0 - - Musk-moskene Influent 0 - - (n=4) Effluent 0 - - Musk-tibetene Influent 0 - - (n=4) Effluent 0 - - Musk-ketone Influent 4 0.049 - 0.069 0.056 (n=4) Effluent 4 0.038 -0.053 0.049

Musk Compounds in Sewage Sludge: The result of analyses into the presence of musk compounds in sewage sludge and the sediment of UWW collecting systems in German commercial and residential areas, are presented in Table e.4. In all samples noticeably high musk scent concentrations were detected. Sediment samples from the UWW collecting system for the residential area had slightly higher concentrations of the three polycyclic compounds ADBI (celestolide), HHCB (galoxolide) and AHTN (tonalide). This is probably due to the more frequent use of perfumes and detergents in domestic areas. Table e.4: Concentration of different musk scents in sewage sludge and UWW collecting system sediment in mg/kg DS, Germany [Antusch, 1999]. N = number of samples analysed. N>LOD number of samples over the limit of detection Compound N> Sediment: Sediment: Sewage sludge LOD industrial area residential area (n=17) (n=2) (n=2) mean range mean range mean range Musk-xylene 6 0.028 <0.005-0.20 0.095 0.066-0.134 <0.005 < 0.005 Musk-ketone 7 0.12 <0.01-1.78 0.25 0.15-0.36 0.03 <0.01-0.06 ADBI 12 0.051 <0.01-0.28 0.35 0.19-0.52 0.20 0.12-0.29 HHCB 17 1.43 0.08-5.2 15.5 9.1-21.8 8.87 4.3-13.4 AHTN 17 2.08 0.13 - 8.9 23.1 9.5 - 36.7 8.30 4.0 - 12.6

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Musk Compounds in Watercourses:

The pollution of the river Ruhr (Germany) with musk compounds was found to be relatively low, with maximum concentrations of 0.08 µg.l-1 and 0.03 µg.l-1 for musk-ketone and musk- xylene, respectively. Fish from the Ruhr contained residues of musk compounds in their muscle flesh, at concentrations below 10 µg.kg-1 wet weight [Eschke et al., 1994].

The polycyclic musks, HHCB (galaxolide, abbalide) and AHTN (tonalide, fixolide) were found in German receiving waters at concentrations up to the µg.l-1 level. In the Wuhle, a small stream consisting almost totally of wastewater effluent, maximum concentrations were 12.5 µg.l-1 for HHCB and 6.8 µg.l-1 for AHTN. Additionally, the polycyclic musk ADBI (celestolide, crysolide) and musk-ketone were detected at low concentrations in the majority of samples. Two other nitro-musks, moskene and xylene, were only detected in a single surface water sample [Heberer et al., 1999].

The concentration of the three compounds tonalide, celestolide and galaxolide, were measured in different watercourses in Germany. The results of this study are shown in Table e.5. The Elbe concentrations for musk-xylene were approximately 0.2 µg.l-1.

Table e.5: Concentration of nitro-musk compounds in different watercourses in the länder Sachsen and Sachsen-Anhalt, Germany [Lagois, 1996].

ADBI (celestolide) HHCB (galaxolide) AHTN (tonalide) [µg.l-1] [µg.l-1] [µg.l-1]

Sample date 22.5.95 12.6.95 22.5.95 12.6.95 22.5.95 12.6.95 Elbe at Torgau <0.08

Conclusions There is very limited information on the health and environmental effects of personal care products, such as the musk compounds found in perfumes. Many of these compounds have the potential to bioaccumulate, which is why there is concern about their presence in wastewater. Though these products may be used in large quantities there is insufficient data though to establish whether the presence of these compounds in UWW could cause any detrimental environmental or health effects.

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(f) Surfactants in Urban Wastewaters and Sewage Sludge

Introduction

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

· Anionic surfactants are anion-active, amphiphilic compounds in which the hydrophobic residues carry anionic groups with small-sized counter-ions, such as sodium, potassium or ammonium ions. These counter-ions have only a slight influence on the surface active properties. Examples include soaps, alkylbenzene sulphonates (ABS), alkylsulphates (AS), and alkylphosphates (AP). · Non-ionic surfactants (NIS) – are amphiphilic compounds that are unable to dissociate into ions in aqueous solutions, for example, alkyl- and alkylphenyl polyethylene glycol ethers, alcohol ethoxylates (AE) and alkylphenol ethoxylates (APE), fatty acid alkylolamides, sucrose fatty acid esters, alkylpolyglucosides, trialkylamine oxides. · Cationic surfactants, cation-active amphiphilic compounds in which the hydrophobic groups exist as cations with counter-ions such as chloride, sulphate or acetate. Examples include tetraalkyl ammonium chloride, N-alkylpyridinium chloride and others. · Amphoteric surfactants have zwitterionic* hydrophilic groups (*electrically neutral ions with both positive and negative charges), such as aminocarboxylic acids, betaines and sulphobetaines.

Uses and sources of surfactants in the environment

The largest proportion of surfactants is used in detergents and cleansing agents for domestic and commercial use [Falbe, 1987]. Surfactants are also used in:

· fabric softeners (cationic), · foam cleaning agents (sulphosuccinates, LAS, AE), · general cleansing agents (LAS, alkylbenzenes, fatty alcohol ether sulphates), · domestic washing up liquids (betaines, NPO, alkylpolyglucosides), · industrial cleansing agents (alkylbenzene sulphonates, alkanesulphonates, fatty alcohol ethoxylates, alkylphenol ethoxylates, fatty amine ethoxylates, ethoxylates, propylene oxide adducts, and others), · bodycare products (ethersulphates, ether carboxylates, betaines, sulphosuccinates of fatty alcohol polyglycol ethers, isethionates, amineoxides, alkyl polyglucosides).

Textile manufacturers uses surfactants extensively as washing agents, also for cleaning, lubricating, bleaching, de-sizing or shrinking (where the mutual adhesion of fibres is reduced), mercerising (cotton treatment that requires wetting agents), and finishing. Wool washing is done with NIS, while cotton is washed with anionic surfactants. The leather industry uses non-ionic and cationic surfactants as wetting and cleansing agents, and also for leather conditioning. Surfactants are also used as emulsifiers and dispersants, and also as food additives (natural substances only, such as glycerides, fatty acid salts, etc.). Pharmaceuticals manufacturing uses surfactants, and they are also used in agricultural applications such as crop-protection and pest control agents. Metal working and machining, petroleum extraction and processing, ore flotation and dressing, mineral oil industry, road construction and maintenance work, cement industry, plastics production, pulp and paper industry and printing, electroplating, manufacturing, all use surfactants for their unique properties.

180 Section 6. Case Studies

Non-ionic surfactants are detergents which possess specific physicochemical properties, including relative ionic insensitivity and sorptive behaviour [deVoogt et.al, 1997] which makes them particularly suited for use wherever interfacial effects of detergents, foaming- defoaming, de-emulsification, dispersion or solubilisation can enhance product or process performance. The major part of the non ionic surfactants group consists of alcohol ethoxylates (AE) and alkylphenol ethoxylates (APE) of which, nonylphenol ethoxylate (NPE) is the main representative. Because of the formation of persistent metabolites in the environment, OSPARCOM member states have decided to phase out the use of NPE and to replace the APEs with AEs. The production of non-ionic surfactants in the USA and EU amounts to about 750 000 t/a and includes some 300 000 t/a of APE [Holt et.al., 1992].

Ionic surfactant molecules contain both strongly hydrophobic and strongly hydrophilic groups. They thus tend to concentrate at interfaces of the aqueous system including air, oily material and particles. The hydrophobic group is generally a hydrocarbon radical (R) of 10 to 20 carbon atoms. The hydrophilic portion may ionise or it may not. Ionic surfactants may be either anionic or cationic. Ionic surfactants constitute approximately two-thirds of the surfactants used. Cationic surfactants constitute less than 10% of the ionics and are used for fabric softening, disinfection and other specialized applications. The predominant class of anionic surfactants is linear alkylbenzene sulphonates (LAS).

The concentrations of linear alkylbenzene sulphonates (LAS) in raw wastewater range from 3 mg.l -1 to 21 mg.l-1 (Brunner et al., 1988, De Henau et al., 1989, Ruiz Bevia et al., 1989). Although LAS and other common surfactants have been reported to be readily biodegradable by aerobic processes, much of the surfactant load into a treatment facility (reportedly 20-50%) is associated with suspended solids and thus escapes aerobic treatment processes, being directed via primary sedimentation into sludge management processes. Because LAS is not biodegraded by anaerobic biological processes usually employed in sludge stabilization (McEvoy and Giger, 1985; Swisher, 1987), it may be found in the gram per kilogram range in anaerobic sludges. Given these concentrations and the major effects of surfactants in particle surface modification, deflocculation, and surface tension reduction, it seems clear that the performance of certain treatment processes as thickening, conditioning, and dewatering may be strongly influenced by these materials.

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

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

181 Section 6. Case Studies

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

Country Influent mg/l Country Effluent mg/l Mean (+/- 95% CIs) Mean (+/- 95% CIs) UK 15.1 (2.3) UK 0.010 (0.002) Germany 5.4 (6.1) Germany 0.067 (0.076) Italy 4.6 (5.1) Italy 0.043 (0.065) Netherlands 4.0 (4.0) Netherlands 0.009 (0.008) Spain 9.6 (9.6) Spain 0.14 (0.14)

As it can be seen from this study, the UK plants had significantly higher influent concentrations than in Germany, Italy and the Netherlands (Spain had a very small data set and therefore there is uncertainty in these results, which is reflected in the large confidence intervals, which overlap with those of the UK). Having a high influent of LAS was unrelated to the concentration in the effluent and the UK samples had a lower concentration of LAS in its effluent than Germany, Italy or Spain. The level of biodegradation of LAS during the wastewater treatment process is high and also varied between the plants, with the UK WWTS having the highest level of biodegradation breaking down greater than 99.9 percent of the LAS.

The study above, used UK data from Holt et al [1995] in which the levels of LAS in the influent corresponded to estimates of LAS usage in homes. However a subsequent study [Holt et al 1996], based on usage data on the influent entering six WWTS in the Yorkshire region, found much lower levels of LAS than expected. In no cases did the concentrations of LAS reaching the plant approach the level predicted from consumer usage. The previous study had been carried out in March while the second was carried out in a ‘warm dry summer’. This study suggested there was significant biodegradation of LAS under certain conditions in the UWW collecting systems. Even in the relatively short residence time in the Yorkshire UWW collecting system, for a couple of hours, up to 60% of the LAS was removed prior to the wastewater treatment plant. This was then followed by removal of between 70 to 99 percent of the remaining LAS in the treatment plants.

Given the right treatment conditions, LAS are biodegradable and more research is necessary to compare risks associated with alternative chemicals used in detergents.

The impact of LAS in wastewater effluent has been studied on several UK and Dutch rivers and steams sediment. In some cases (such as the small stream into which the Owlwood WWTS discharges its effluent the LAS load in sediment upstream of the treatment plant was higher than that downstream [Waters and Feijtel 1995] This was hypothesised to be because that the LAS contribution upstream was due to unregulated discharges of untreated wastewater to the aquatic environment while the input of the WWTS effluent actually served to dilute the concentration of this pollutant downstream. A similar case was found in the Netherlands when concentrations of LAS could be higher upstream of a WWTS due to direct discharges from storm tanks. In other locations downstream sites were found to have very slight elevations of LAS in sediment of between 0.49 and 3µg/g. A more careful policy of discharges has to be followed across the EU.

A new study by NERI [NPE/DEHP in sewage sludge in Denmark, http://www.dmu.dk/beretuk98/society.htm#c4] shows that only small amounts of NPEs and DEHPs are discharged in the urban wastewater by the commercial sector and that they do not accumulate in agricultural soils treated with sewage sludge in moderate amounts. Figure d.1 shows the level of surfactants and plasticisers in soils with the depth of the soil.

182 Section 6. Case Studies

Fig. d.1 Vertical distribution of various nonylphenols and phthalates in agricultural soils fertilized with large amounts of sewage sludge (17 tonnes dry matter per hectare per year). [NERI, 2000]

Fate and effect of surfactants in the environment

Fate of surfactants during wastewater treatment: LAS and other common surfactants have been considered to be readily biodegradable by aerobic processes, based on laboratory studies (Swisher, 1970). Figure d.2 shows schematically the fate of LAS in the environment (deWolf and Feijtel, 1998). LAS evidently undergoes nearly complete biodegradation, with 97-99% removal rates found in some wastewater treatment plants (Brunner et al., 1988; Bevia et al., 1989; De Henau et al., 1989). However, the mass loadings indicated above suggest that even at these removal rates, appreciable amounts are released to receiving waters. Ventura et al. (1989) identified LAS and a variety of other anionic, cationic and nonionic surfactants in both surface and drinking water extracts.

183 Section 6. Case Studies

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

Alkylphenol ethoxylates such as NPnEO are evidently less biodegradable than LAS with laboratory results ranging from 0-20% based on oxygen uptake (e.g. Swisher, 1970; Steinle, 1964; Pitter, 1968) and a wider range of removals from 0-90% based on specific analyses such as UV and IR spectroscopy (Swisher, 1970). This suggests that only partial degradation occurs, such as conversion from polyethoxylates to nonylphenol diethoxylate (NP2EO), nonylphenol monoethoxylate (NP1EO), and nonylphenol (NP). Mass balances carried out on treatment plants in Switzerland (Brunner et al., 1988) support this.

The findings of Brunner et al. and other reserachers, also show that the nearly complete removal of surfactants from treated waters is not entirely due to biodegradation. Brunner et al. indicated that 19% of the surfactant load entering a treatment facility is associated with suspended solids, and other studies report levels up to 27% (Rapaport and Eckhoff, 1990), or even in excess of 50% (Bevia et al., 1989). The surfactant load linked to suspended solids is directed into sludge treatment processes via primary sedimentation. Surfactants such as LAS are not biodegraded by either mesophilic or thermophilic anaerobic digestion (McEvoy and Giger, 1985; Swisher, 1987) so a large proportion of these materials simply escapes treatment and becomes associated with sludge solids.

The resulting concentrations of surfactants in sewage sludges can be substantial. LAS concentrations measured in sludges often make up between 0.5% and 1.5% of the dry solid mass, particularly for anaerobically digested sludges (McEvoy and Giger 1986; De Henau et al. 1989; Holt et al. 1989; Marcomini et al. 1989). Bevia et al. (1989) reported LAS between 2% and 4% of the sludge solids weight. In a study of 29 Swiss treatment plants, LAS concentrations averaged 4.2 and 2.1 g kg-1 respectively in anaerobic and aerobic sludges. NP exceeded 1 g kg-1 dry sludge and, in some instances NP1EO and NP2EO exceeded 0.1 g kg-1 dry sludge (Brunner et al. 1988).

184 Section 6. Case Studies

Effects of surfactants on wastewater treatment

As stated previously, given these surfactant concentrations and the considerable effect that surfactants can have on the properties of suspensions such as sludges, the performance of such processes as thickening, conditioning, and dewatering may be strongly influenced by these materials. For example, Bierck and Dick (1988) have shown that surface tension of sludge solids is directly related to the capillary pressure available for solids compression during the latter stages of vacuum filtration: Ps,s = v [1/R1 + 1/R2]

Where, Ps,s = the pressure or effective stress producing solids shrinkage, v = the surface tension, R1 and R2 = principal radii of curvature of the solid surface.

Thus the effect of surfactants, in lowering the surface tension, is to decrease the compressive dewatering by allowing gas penetration of the solids cake. Campbell et al. (1984, 1986) showed that a detergent could decrease the dewaterability of a sludge even before the compressive phase, as indicated by capillary suction time (CST) measurements. Household detergent added to anaerobically digested sludge at 0.2 and 0.3% by volume caused significant increases in the CST (poorer dewaterability) which could not be compensated for even by doubling the addition of cationic polymer used as the sludge conditioner.

The implications of surfactants' influence on dewatering should not be underestimated. Costs for the sludge conditioning polymer are the greatest operating cost for dewatering at a WWTS such as Wilmington, and sludge dewatering and disposal represent up to 50% of the total cost of wastewater treatment (Evans, 1988).

The biodegradation mechanism of LAS was described by Balson and Felix(1995). The mechanism of breakdown of LAS involves the degradation of the straight alkyl chain, the sulphonate group and finally the benzene ring. The breakdown of the alkyl chain starts with the oxidation of the terminal methyl group (w-oxidation) through the alcohol, aldehyde to the carboxylic acid as follows (see Fig. d.3a). The reactions are catalysed by alkane monooxygenase and two dehydrogenases. The carboxylic acid can then undergo b- oxidation and the two carbon fragment enters the tricarboxylic acid cycle as acetylCo-A. It is at this stage that problems arise with branched alkyl chains, a side chain methyl group or a gem-dimethyl-branched chain cannot undergo b-oxidation by microorganisms and must be degraded by loss of one carbon atom at a time (a-oxidation, Figure d.3b). (Scott and Jones, 2000).

185 Section 6. Case Studies

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

The second stage in LAS breakdown is the loss of the sulphonate group. The loss of the alkyl and the sulphonate group from LAS leaves either phenylacetic or benzoic acids. Microbial oxidation of phenylacetic acid can result in fumaric and acetoacetic acids and benzene can be converted to catechol .

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

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Effects of the surfactants on the wider environment:

The presence of surfactants in sewage sludge may have undesirable environmental effects if land application is the chosen disposal method. The surfactant molecules may leach to groundwater contributing to groundwater contamination. Federle and Pastwa (1988) studied the percolation of anionic and nonionic surfactants through a soil column. Most of the surfactant was mineralised, but this process was found to be highly dependent on the number of organisms present in the soil. A number of reports (e.g. Bevia et al., 1989; Holt et al., 1989) attribute observed decreases of LAS concentrations over time in sludge-amended soils to biodegradation, without evaluating possible migration. Marcomini et al. (1989) reported a fraction of LAS in sludge-amended soil to be resistant to biodegradation over long time periods. Table d.2 contains data on fate and persistence of surfactants in sludge amended soils.

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

Application Country Surfactant/ Soil Monitoring Final Half Life Author Form derivative Concentration period Soil (days) post Conc. application -1 -1 (mg.kg ) (mg.kg )

Sludge onto SP LAS 22.4 6 months 3.1 Not Prats et al soil 12 months 0.7 reported Sludge onto CH LAS 45 12 months 5 9 Marcomini et al soil NP 4.7 0.5 Surfactant D LAS Not reported 2 months Not 5-25 Litz et al onto soil 6 months reported summer 66-117 winter Sludge onto D LAS 16 76 days 0.19 13 Figge and soil 27 106 days 0.44 26 Schoberl Surfactant USA LAS 0.05 40 days Not 1.1- Knaebel et al onto soil LAE 0.05 reported 3.6 Sludge onto SP LAS 16 90 days 0.3 26 Berna et al soil 53 170 days Not 33 reported Sludge onto UK LAS 2.6-66.4 (*) 5-6 months <1 7-22 Water et al soil Holt et al Sludge onto UK LAB 0.3-9.5 (*) 55 days 0-0.38 15 Holt and soil Bernstein Composted AUS NPE 14 000 14 weeks 1200 Not Jones and wool scour reported Westmoreland sludge (* estimated cumulative load)

Not only may surfactants migrate to groundwater, but they may also carry hydrophobic organic pollutants with them. The degree of partitioning of hydrophobic organic pollutants to particles depends on the hydrophobicity of the pollutant and the amount of organic matter contained in the particle. Dissolved organic matter tends to decrease the potential for sorption by providing an additional aqueous phase to which the pollutant can partition (Enfield et al. 1989). Partitioning of surfactant to sludge particles in the sewage treatment plant would be expected to enhance the partitioning of organic pollutants to sludge. When applied to land, desorption of surfactant could lead to pollutants also being released. Kile and Chiou (1989) studied the effect of anionic, cationic and nonionic surfactants on the water solubility of DDT and trichlorobenzene. The results were extremely surprising. As would be expected, the solubility was enhanced when the surfactant was present at

187 Section 6. Case Studies concentrations greater than the critical micelle concentration. There was also a solubility enhancement at surfactant concentrations less than the critical micelle concentration.

In addition to the effects of surfactants in sludge on pollution of groundwater, the surfactants may effect soil texture and water retention through processes similar to those discussed with respect to sludge dewatering. Holtzclaw and Sposito (1978) determined LAS content in a sludge amended soil to be high enough (1% of the fulvic acid fraction) that could be affected.

The fate of surfactants in sludges disposed of in landfills is somewhat surprising. Concentrations of LAS up to 1% by weight have been found in recently deposited material, with some amounts above 1 g.kg-1 even after 15-30 years (Marcomini et al., 1989). Given that landfills function in a similar manner to anaerobic digesters, the persistence of LAS is evidently due to its poor degradability in such environments. The role of surfactants in mobilizing less hydrophobic contaminants into landfill leachate is thus a relevant concern.

Behaviour of nonionic surfactants:

Recent studies have revealed that fish living downstream of wastewater treatment plants show oestrogenic effects [Purdom et al. 1994] as a result of alkylphenol polyethoxylates (APE) and nonylphenol (NP) present in the water. Male fish produce vitellogenin, a yolk which is formed under the influence of oestradiol and therefore is typically produced by females. Hermaphrodite fish species have been found as well. The decomposition products of APE, are considered as a potential cause, since their decomposition products formed in WWTS (Giger et al. 1984) show slightly oestrogenic effects (Soto et al. 1991, Jobling and Sumpter 1993).

Alkylphenol polyethoxylates (APE) usually enter surface waters via WWTS, where they are degraded - but not totally - by microorganisms. In a first rapid step the ethoxylate groups are split off by hydrolysis, and the metabolites nonylphenol (NP), nonylphenol ethoxylate (NP1EO) and nonylphenol diethoxylate (NP2EO) are formed. These metabolites are more toxic than the original substances. Due to the hydrophobic properties of the aromatic group the second step of biodegradation occurs much slower. The interim products can also be biodegraded to alkylphenoxy ethoxylate carboxylic acids (APEC). The second, slower, step of biodegradation, does not always occur, and the fact that the metabolites are more lipophilic than the parent compounds can cause an accumulation of interim products in sludge and sediment. Nonylphenol, for example, was determined in digested sludge in concentrations between 0,45-2,53 g.kg-1 dry weight (Giger et al. 1984). Approximately 50 % of the APE occurring in the wastewater are estimated to reach the sludge as NP (Brunner et al. 1988). Before prohibition of APE in washing agents NP, NP1EO and NP2EO concentrations between 36-202 µg/l were found in drain channels of WWTS in Switzerland. Now NP concentrations in drain channels from WWTS, are found at concentrations between 1 and 15 µg.l-1 in Switzerland and Germany; other metabolites (NP1EO, NP2EO, NP1EC) are normally determined to be between 1 and 40 µg.l-1 (Ahel et al. 1994a, Ahel et al. 1994b, Giger 1990). Concentrations of 15 µg.l-1 in drain channels of WWTS were determined in the USA. In highly polluted streams average nonylphenol concentrations are determined to be in the range 0.3 to 3 µg.l-1 (Ahel et al. 1994a), polyphenoxy carboxylic acids products are predominant, whereas in sediment NP was the dominating degradation product. Due to their high octanol/water partition coefficient (log 4.0-4.6) nonylphenol, NP1EO and NP2EO show a tendency towards bioaccumulation in organisms. This was confirmed by residue analyses (Table d.3). The bioconcentration factor in fish is approx. 300, in one case, however it amounts to 1300.

188 Section 6. Case Studies

Table d.3: Environmental concentrations of degradation products of nonionic surfactants Environmental Substance Concentration Reference compartment Sewage sludge NP 0,45 - 2,53 g kg-1* Giger et al., 1984 0,03 g kg-1* Giger and Alder, 1995 WWTS-drain NP, NP1EO, NP2EO 36 - 202 µg l-1 Stephanou and Giger, NP 1 10 µg l-1 1990 NP1EO, NP2EO 1 - 40 µg l-1 Streams NP 0,3 - 45 (2-3) µg l-1 Ahel et al., 1994b NP1EO, NP2EO < 3 - 69 µg l-1 NP1EC, NP2EC < 2 - 71 µg l-1 Stream sediment NP 0,5 - 13 mg kg-1* Ahel et al., 1994b Fish NP, NP1EO, NP2EO 0,03 - 7,0 mg kg-1* Ahel et al., 1993 Algae NP, NP1EO, NP2EO 80 mg kg-1* Ahel et al., 1993 Waterfowl (ducks) NP, NP1EO, NP2EO 0,03 - 2,1 mg kg-1* Ahel et al., 1993 * dry weight

Health effects of surfactants:

Prats.et.al, 1993 show significant differences between distribution of LAS homologs in water and solids (sludges, sediments, and soils), as compared to the original distribution in detergent formulations, yielding a lower LAS average molecular weight in water samples. The change observed in the homolog distribution of LAS implies a reduction in the toxicity to Daphnia, because a lower average molecular weight of LAS is less toxic. The risk assessment of LAS to terrestrial plants and animals reported by Mieure et al. (1990) also concludes that there are adequate margins of safety in the use of wastewater for the of plant species. Adverse effects on plant and animal species (earthworms Eisena foetids and Lumbricus terrestris) were observed at LAS concentration of 10 mg.l -1 , however LAS concentrations in wastewater effluents are in a range 0.09 mg l-1 to 0.9 mg l-1 . These figures give a safety margin in a range 10 to 100. The effect of surfactant on plant growth from the use of sewage sludge is difficult to assess because in general the sludge promotes plant growth. Adverse effects on plant growth were observed at 392 µg g-1 but long term monitoring at a range of 46 environmental sites gave LAS concentrations of less than 3µg.g-1 . These figures give a safety margin of 131. For terrestrial animals the limit of no adverse effects was 235 µg.g-1 giving a safety margin of 78. However, in looking at ecotoxicity from WWTS effluents the less toxic surfactant residues and surfactant catabolites must be considered and this requires analytical tests for these entities (Scott and Jones, 2000; Schoberl, 1997).

Amounting to 2-4 g.kg-1 the acute mammalian (mouse, rat) toxicity of APE is low. Dermal toxicity, however, is higher (500 mg.kg-1), and eye irritation is the highest with 5-100 mg.kg-1. NP can be metabolised to a glucoronide in the body and excreted via the kidney. Nonionic surfactants are more toxic for aquatic organisms than for mammals. The toxicity of APE increases with decreasing number of ethoxylate units and increasing hydrophobic chain length. Accordingly, the toxicity of the original substances is lower than the toxicity of the metabolites NP, NP1EO and NP2EO, whereas the carboxylic acids are less toxic than the ethoxylates. For instance the LC50 (48 hours) of NP16EO is 110 mg.l-1 for fish (Oryzias latipes) and decreases to 11,2 and 1,4 mg.l-1 for NP9EO and NP, respectively (Yoshima, -1 1986). The LD50 (96 hours) for algae (Skeletonema costatum) is 27 µg.l , and the value for rainbow trouts 480 µg.l-1 (Nayler 1992). The no observed effect concentration (NOEC) for reproduction for Daphnia is in the range of 24 µg.l-1. These data show that the acute toxicity of NP is considerably high.

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

189 Section 6. Case Studies based on the vitellogenin synthesis revealed that NP, NP1EO and NP1EC have the same activity (half maximum activity: around 16 µM). The oestrogenic activity, however, is 10 times lower than that of oestradiol (Pelissero et al. 1993). Other in vitro studies give hints on potential differences between fish and mammalia regarding the binding to the oestrogen receptor (Thomas and Smith 1993). However, vitellogenin synthesis in fish hepatocytes is also induced by well-known phyto-oestrogens. Studies in the UK indicate that downstream of the drain channels of WWTS vitellogenin is formed in male fish. After 1 to 3 weeks exposure of fish in 15 drain channels of WWTS, displayed a high increase of vitellogenin synthesis (Purdom et al. 1994). It is supposed that the decomposition products of APE, especially NP, are mainly responsible for this effect. The assumption is confirmed indirectly by the results of the in vitro studies with fish hepatocytes. However, it cannot be excluded that synthetic oestrogens are also responsible for this effect. On the one hand their concentrations are lower than the usual concentrations of NP, but on the other hand their activity is some orders of magnitude higher. Experimental exposure of fish to NP or metoxychlor over 7 days induced vitellogenin synthesis in male fish (Nimrod and Benson 1995). The dose required to induce the vitellogenin synthesis was 300 times (approx. 150 mg.kg-1) higher than the necessary dose of oestradiol.

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

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

Best environmental practice examples One of the main success stories regarding the use and fate of surfactants is linked with eco- labelling. Eco-labelling has been developed for the products used in dishwashing, laundry and cleaning detergents in Scandinavia, Germany, Austria and other European countries. Products with the ‘Swan’ and ‘Good Environmental Choice’ label do not contain LAS and Nonylphenol and have gained considerable market share. In Sweden, products with these labels accounted for more than 95% of sales by 1997 while in Finland they reached 15%. Norway and Denmark had lower sales (source, Danish Environment Agency 2000). More research is needed for the potential environmental effects of the alternatives used in these LAS-free and NPE-free surfactants.

This public awareness and consumer choice, lead to the use of LAS in Sweden falling from 6300 tonnes a year to 260 tonnes a year. The Danish Environment Agency launched a public campaign against LAS in September 1999. Currently about 2500 tonnes a year of LAS are used in Denmark.

During the development of the “Swan” mark, Stockholm water company identified the need to have an alternative for taking care of hazardous waste in the household rather than flushing it into the UWW collecting system. Environmental stations, or collection points were established and an extensive public information campaign was carried out about the impacts of household products on the aquatic environment [Ulmgren, 2000a, 2000b].

Detergents and cleaning agents containing alkylphenolethoxylates (APEO), such as NPE, are being gradually phased-out under various initiatives and voluntary agreements in EU. Distearyl-dimethylammonium chloride (DSDMAC), widely used in laundry softeners, was

190 Section 6. Case Studies also substituted with more degradable substances during the 1990s in Germany [Greiner, 1996] and is currently under scrutiny in the rest of the EU.

The eco-labelling combined with a public awareness campaign could therefore influence consumer choice and reduce contaminant discharges in the UWW from domestic products.

More research is necessary to experimentally determine the role of surfactants in sludge treatment processes and following sludge disposal in the environment. Specific effects to be investigated are

· impacts on sludge thickening, conditioning, and dewatering processes and · transport and mobilization of hydrophobic organic contaminants when sludges are landfilled or land-applied. Also anticipated is an improved fundamental understanding of mechanisms by which surfactants are incorporated into sludge solids.

Some important research gaps and necessary research are summarised as follows:

Effects of endocrinally active chemicals have not yet been systematically investigated in amphibian and reptiles. In this field nearly no knowledge is available. · Chemical methods for the detection of traces of synthetic oestrogens and their metabolites must be elaborated, since only very few data are available on environmental concentrations, especially regarding concentrations in drain channels of wastewater treatments plants. Furthermore, data material on NP concentrations in drinking water and organisms including humans is insufficient. · The ecotoxicological relevance of vitellogenin production in male animals has to be elucidated. Which interrelations exist between the problem of vitellogenin production and further estrogenic and ecotoxicological effects of NP and other chemicals? To answer these questions in vivo experiments using histopathological, biochemical, endocrinological and reproduction biological methods have to be conducted. Furthermore, insufficient information is available about the bioaccumulation of these chemicals. In a further step the problem should be investigated by more comprehensive field studies. · The mechanisms of chronic effects of alkylphenols (modes of action) must be studied in more detail. · Finally, in vitro assays should be elaborated to identify and estimate the oestrogenic activity of existing new chemicals in fish and other organisms.

191 Section 6. Case Studies

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

Polyacrylamide (PAM) is a widely used flocculant in water treatment applications. Some 20,000 tonnes is used in the USA for this purpose each year. Concerns with its use are that it can degrade to the acrylamide monomer which is known to affect the central and peripheral nervous systems and is also believed to be carcinogenic. Safe levels for this chemical are said to be 300 µg l-1 over a ten-day period and 2 µg l-1 over a seven year period (EPA). PAM is used in other applications such as an aid in irrigation (Trout et al. 1995) and in pulp and paper manufacture. The EPA also notes that it is used in formulating grouting for tunnels and sewers. Effluents from a sewage works which used PAM as a flocculant in the UK were reported to be 2.3 to 17.4 µg l-1. High levels of the monomers have been reported in acrylamide manufacturing plant effluents. In this case the raw effluent contained 1100 µg l- 1 and the treated effluent 280 µgl-1 (EPA). Both PAM and its monomer are very soluble in water and the presence of PAM in soil causes leaching of microorganisms by ground or irrigation water.

PAM is shown to degrade by biological action and photolytic effects (Nakamiya 1995). Experiments have shown that polyacrylamide solutions in a bottle covered with plastic film and left outside can contain significant amounts of the monomer after two weeks exposure (Smith et al. 1996). The polymer can also be degraded by turbulent shear stress in and pipes (Rho et al. 1996). Once the polymer degrades the monomer is also subjected to more rapid degradation in which it is decomposed to acrylic acid and ammonia. These are non toxic as acrylic acid degrades to CO2 and water in a day in soil (Staples et al. 2000) and is thus not an ecological problem. The degradation of acrylamide under favourable conditions by pseudomonas species immobilised in calcium aliginate took one day (Nawaz et al. 1993). The EPA say that degradation of acrylamide in river water takes 4 to 12 days and is more rapid in summer than winter. Due to its solubility adhesion of the monomer on soil is unlikely, though it is reported that it is partially removed by secondary activated sludge.

The general conclusion of the EPA paper is that the monomer is not an environmental hazard when released in small quantities to the aqueous environment. Tests with fish would indicate this. The monomer is relatively biodegradable within days compared to the time span of years for substances such as PCB. There are two areas where there could be some concern and these are control of effluents from acrylamide manufacturing and use in drinking water treatment. The fact that the monomer is detected in water treatment plants where the residence time is only a few hours suggests that the PAM flocculant could have significant residual amounts of monomer.

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

· Chlorination and the presence of potentially toxic elements can stop acrylamide degradation by passifying the bacteria present. · Degradation of the monomer is not pH dependant · 50% of PAM is removed in aerated sludge and trickle bed filters.

A significant drop in the percentage of monomer present in the PAM used for flocculation has eased the likelihood of serious contamination of water. The present level of monomer present in grades of PAM used for water treatment is 0.3% and has dropped from 0.8%. However PAM used in grouting has a much higher monomer content and the use of this particular grade has caused the monomer to leak into grouted sewers.

To conclude it seems that whenever water is sent into the environment it would be safe to use PAM as a coagulant. Problems may arise when it is used in a stream which is

192 Section 6. Case Studies subsequently sterilised by chlorine or by another disinfectant. This removes the bacteria that are needed to degrade the PAM and it would be a problem with using other organic polymers as well. It might be concluded that a polymer grade containing just 0.3% of monomer is very pure and improvements in purity might not be practically feasible.

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

193 Section 6. Case Studies

(h) Landfill Leachate

Introduction

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

With the introduction of redrafted legal conditions in Germany, strict demands have been placed on the purification performances of leachate treatment plants. Thus, the treatment of organic substances and nitrogen compounds using nitrification and , is required prior to direct discharge to a receiving watercourse.

Wastewater and leachate quality requirements in Germany

Definitions of direct and indirect discharge of wastewater are as follows:

· Direct discharge: To discharge wastewater directly must conform to standards of water quality in the receiving water. · Indirect discharge: Discharging wastewater directly into a public WWTS requires that the concentration of COD, BOD, NH4-N, AOX and potentially toxic elements must be reduced to the same levels found in domestic wastewater.

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

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

Potentially toxic Symbol Standard element Concentrations [mg/l] Lead Pb 1 Cadmium Cd 0.5 Chromium Cr 1 Chromium (VI) Cr(VI) 0.2 Copper Cu 1 Nickel Ni 1 Mercury Hg 0.1 Zinc Zn 5 Regulation AbwV: “Requirements for discharging wastewater into watercourses”

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

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

194 Section 6. Case Studies kept low by proper measures and operation at the landfill installation. The requirements listed in Table h.2 relate to the discharge site of leachate into watercourses.

Table h.2: Requirements for wastewater quality at point of discharge (qualified sample or 2 hour mixed sample) [AbwV, annex 51, 1999].

Parameter Unit Value COD* mg/l 200 BOD mg/l 20 Ntotal** mg/l 70 Ptotal mg/l 3 Hydrocarbons, total*** mg/l 10 N02-N mg/l 2 Fish toxicity GF 2 * For wastewater with a COD value (before treatment) of more than 4000 mg/l, the COD effluent value in the qualified sample or in the 2 hours mixed sample must be reduced by 95 %. ** Sum of ammonium-, nitrite- and nitrate-nitrogen (Ntotal) or total bound nitrogen (TNb). The requirement applies to a wastewater temperature of 12 °C. A higher limit concentration of 100 mg/l is permitted when the decrease of nitrogen load amounts to at least 75 %. *** The requirement relates to the qualified sample.

The requirements listed in Table h.3 below relate to leachate before mixing with other wastewaters.

Table h.3: Requirements on leachate before mixing (qualified sample or 2 hours mixed sample) [AbwV, annex 51, 1999 Parameter Unit [mg/l] AOX* 0.5 Mercury 0.05 Cadmium 0.1 Chromium 0.5 Chromium VI* 0.1 Nickel 1 Lead 0.5 Copper 0.5 Zinc 2 Arsenic 0.1 Cyanide, easily released* 0.2 Sulphide* 1 * value for the qualified sample.

195 Section 6. Case Studies

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

· fish, indicator bacteria and Daphnia toxicity of a representative sample is not exceeded (see Table h.3). It has to be stated that exceeding the GF value is not caused by ammonia (NH3). · a DOC elimination rate of 75 % is reached. · leachate shows a COD concentration lower than 400 mg.l-1 before the common biological treatment.

Table h.4: Fish, indicator bacteria and Daphnia toxicity [AbwV, annex 51, 1999]

Fish toxicity GF = 2 Daphnia toxicity GD = 4 Indicator bacteria toxicity GL = 4

Landfill Leachate

Formation: A substantial proportion of pollutant emissions from landfill sites enters percolating through the landfill. Rain water (and other sources of water) entering unsealed sections of the landfill undergo chemical and biological transformation in the body of the landfill to form leachates. Pollutants are taken up by solution processes or are carried in suspension. This loaded water, the so called leachate, is collected at the base of the landfill in a drainage pipeline.

The quantity of leachate produced depends principally on rainfall and the state of the landfill. At new, unsealed landfill sites, the total calculated rainfall collects as leachate. During the life of the landfill leachate quantity reduces to 10 – 20 % of total rainfall, with an increase in superficial sealing.

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

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

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

196 Section 6. Case Studies

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

Parameter Units Industrial and special waste Household refuse landfill landfill Range Mean value Range Mean value pH - 5.9 – 11.6 7.7 3.5 – 9 7.5 -1 COD mg O2.l 50 – 35000 5746 500 - 60000 5000 -1 BOD5 mg O2.l 41 – 15000 2754 100 - 45000 1500 Conductivity mS.cm-1 2110 – 183000 28217 - 10000 Chloride * mg.l-1 36 – 126300 13257 100 - 15000 2000 Sulphate mg.l-1 18 – 14968 2458 50 - 3000 300 Ammonium* mg.l-1 5 – 6036 921 20 - 3000 500 Nitrite* mg.l-1 0.02 - 131 7.3 - 0.5 Nitrate* mg.l-1 0.1 - 14775 606 0 - 50 3 Total-N* mg.l-1 1 - 3892 461 20 - 4000 600 Total-P* mg.l-1 0.03 - 52 7.9 0.01 - 10 1 Fluoride mg.l-1 0.1 - 50 13.3 - - Total cyanide mg.l-1 0.007 - 15 1.3 - - Easily released mg.l-1 0.008 - 1 0.2 - - cyanide Arsenic* mg.l-1 2 - 240 51 0.1 - 1000 20 Lead* mg.l-1 4.3 - 650 155 20 - 1000 50 Cadmium mg.l-1 0.2 - 2000 144 1 - 100 5 Copper* mg.l-1 1.3 - 8000 517 10 - 1000 50 Nickel* mg.l-1 14.2 - 30000 2096 20 - 2000 200 Mercury mg.l-1 0.17 - 50 5.5 - 10 Zinc mg.l-1 20 - 272442 2936 100 - 10000 1000 Chromium (total)* mg.l-1 0.009 - 300 18.1 0.02 - 15 0.2 Iron mg.l-1 0.38 - 2700 144 1 - 1000 50 Phenol index mg.l-1 0.01 - 350 26 - 0.006 Hydrocarbons mg.l-1 0.01 - 424 30 - - AOX mg.l-1 44 - 292000 32000 320 - 3350 2000 *leachate substances not influenced by the biochemical state of the landfill matter.

Treatment Practices Different methods can be used for treating leachate from landfills, consisting principally of biological, physical and chemical processes. A specific process can only treat a particular substance categories in wastewater. Because of the wide range of pollutants found, leachate treatment has to be performed using a combination of suitable processes. The choice of treatment processes depends closely on the leachate composition. A short description of processes used in Germany for treating leachate follows below.

Biological Practices: Biological process can be used to degrade leachate pollutants into mineral end products. To enable degradation specialised microorganisms must be enriched in the bioreactors by proper process conditions. Nitrogen elimination can also be obtained by nitrification and denitrification. Biological processes, especially the aerobic ones, are efficient and cost- effective in comparison with the chemical-physical processes (Rudolph et al., 1988). The activated sludge process and the biofilm process are both use to biologically treat leachate from landfills.

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

197 Section 6. Case Studies concentrations of persistent organic compounds, forming a fine, dispersed sludge, which does not settle readily. So the biomass passes through the activated sludge plant without treatment. Under these conditions biological degradation of pollutants is not possible (Albers, 1991, Wilderer et al., 1989).

Biofilms: Biofilm systems can be used to prevent the loss of biomass by washing-out, which may be experienced in the activated sludge process. Biomass growth is encouraged by attachment to support surfaces, in form of biofilm. SBBR, the so called sequencing batch biofilm reactors, are also used for cleaning leachate with high salt concentrations and a high percentage of persistent organic compounds. Advantages of the biofilm processes are the small space requirement and the high flexibility in service (Wilderer et al, 1989).

Chemical-physical methods: Flotation, precipitation and flocculation, adsorption, and thermic techniques, belong to the chemical-physical processes for treating leachate. Other chemical-physical processes are chemical oxidation and membrane filters.

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

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

Adsorption: At a leachate treatment plant, adsorption by activated carbon is always used in combination with biological pretreatment or with a chemical-physical process. Any persistent organic compounds not degraded in the pretreatment step and AOX compounds, can be separated in the back-washed carbon filters. Through adsorption processes, an agglomeration of the solute molecules takes place on the activated carbon interface. Advantages of the adsorption process are; simplicity of the technology involved; relatively low running costs; and possible thermic recycling of the exhausted carbon (Detter, 1998). Regeneration of activated carbon is problematic and expensive though.

Chemical oxidation: In the oxidation stage of a leachate treatment plant non-biodegradable and inhibitory organic substances can be oxidised or reduced. In ideal conditions, given a sufficient supply of medium for oxidation, complete mineralisation can be achieved. Substances such as potentially toxic elements and neutral salts remain in solution and are not transformed (Döller, 1998). Hydrogen peroxide/UV or ozone/UV are the mainmedium used for oxidation medium in leachate treatment. In practice, leachate is enriched with ozone (O3) or hydrogen peroxide (H2O2) and afterwards conducted to the UV radiators.

Thermal treatment: In pollutants in leachate are separated from water (stripping), concentrated (vapourising) and mineralized (combustion). Due to the different volatilities of water, organic solvent and of dissolved and suspended substances, partition by distillation can be achieved. Thus volatile hydrocarbons contained in the leachate can be separated with a stripping step. With the vapourising process, inorganic and organic residual substances are obtained separately in a chemically unchanged form. Then, the concentrated organic phases must be made inert by combustion. Proper treatment of the exhaust gases is necessary to meet air quality emission standards. During the vapourisation of critically loaded leachate, single toxic halogen organic compounds such as polychlorinated biphenyls, dibenzo-dioxins and dibenzo-furans can enter the distillate. In this case, post treatment with activated carbon is essential (Leonhard, Wilderer,. 1992).

198 Section 6. Case Studies

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

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

Conclusions

In Germany, the discharge of wastewater into public WWTS and into receiving watercourses is strictly regulated. Thus, leachate must also be treated before discharging and legal regulations set high requirements on the performance of leachate treatment plants (ATV- A115 1994 and AbwV 1999).

Prior to the discharge of non-domestic wastewater into a public WWTS, concentrations of COD, BOD, NH4-N, AOX and potentially toxic elements must be reduced to at least domestic wastewater standards. Purified leachate contributes only a relatively small proportion of the pollutant load WWTS.

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

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

199 Section 6. Case Studies

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

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

The removal of metals during secondary wastewater treatment is dependent upon the uptake of metals by the microbial biomass and the separation of the biomass during secondary sedimentation. Several mechanisms control metal removal during biological secondary treatment including:

· physical trapping of precipitated metals in the sludge floc · binding of soluble metal to bacterial extracellular polymers

In general the patterns in metal removal from settled sewage by secondary treatment are similar to those recorded for primary sedimentation. However, the general survey of removal efficiencies listed in Table i.1 suggests that secondary treatment (by the activated sludge process) is more efficient at removing Cr than the primary stage. Operational experience and metal removals measured by experimental pilot plant systems provide guidance on the overall likely removal and transfers to sludge of potentially toxic elements from raw sewage during conventional primary and secondary wastewater treatment. This shows that approximately 70 – 75 % of the Zn, Cu, Cd, Cr, Hg, Se, As and Mo in raw sewage is removed and transferred to the sludge (Blake, 1979) and concentrations of these elements in the final effluent would be expected to decrease by the same amount compared with the influent to the works. Lead may achieve a removal of 80 %, whereas the smallest overall reductions are obtained for Ni and approximately 40 % of this metal may be transferred to the sludge.

The majority of potentially toxic elements in raw sewage are partitioned during wastewater treatment into the sewage sludge or the treated effluent. However, atmospheric volatilisation of Hg as methylmercury, formed by aerobic methylation biotransformation processes, is also suggested as a possible mechanism contributing to the removal of this element during secondary wastewater treatment by the activated sludge system (Yamada et al., 1959). Whilst it some of the Hg removal observed in activated sludge may be attributed to bacterially mediated volatilisation, it is unlikely that this is a major route of Hg loss because of the significant quantities of Hg recovered in surplus activated sludge (Lester, 1981).

200 Section 6. Case Studies

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

PTE Removal (%) Primary(1) Secondary(2) Primary + Primary + secondary secondary(3) Zn 50 56 78 70 Cu 52 57 79 75 Ni 24 26 44 40 Cd 40 40 64 75 Pb 56 60 70 80 Cr 40 64 78 75 Hg 55 55 80 70 Se 70 As 70 Mo 70 (1)Mean removal (n = 5) from raw sewage and transfer to sludge during primary sedimentation (2)Mean removal (n = 9) in activated sludge from settled sewage (3)Blake (1979)

201 Section 6. Case Studies

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

The chemical treatment of wastewater to remove phosphorous is increasingly practised to control P discharges and as a measure to reduce eutrophication of sensitive water courses. This also has the advantage of increased BOD removal, reduction in polyelectrolyte coagulant consumption for sludge thickening, elimination of hydrogen sulphide in sludge digesters and reduced consumption of chemicals for scrubbing (Abendt, 1992). High rates of P removal can be achieved from wastewater using common precipitants such as aluminium sulphate (alum) and ferric chloride although this influences both the quality and quantity of sludge produced (Yeoman et al., 1988). Chemical precipitation also enhances the removal of potentially toxic elements from sewage effluent compared with conventional treatment practices, increasing the transfer of metals to sewage sludge and the content of metals in sludge. For example, Stones (1977) measured the reductions in metal concentrations in sewage effluents obtained after an 18 h settling period with aluminium sulphate (Al2(SO4)3) compared with sedimentation without Al salt. The removal of all the elements examined was increased by the addition of aluminium sulphate compared to the unamended control, except for Ni (Table j.1). The removal of Cu and Zn from the effluent was raised by approximately 50 % by chemical treatment compared with removals achieved by sedimentation without addition of Al. Lead removal increased by about 80 % and the largest overall increase relative to the control was obtained for Cr. In the case of Cr, precipitation with aluminium sulphate increased the recovery of this element in the sludge almost by a factor of three.

Table j.1 Effect of chemical precipitation on metal removals (%) from raw sewage after 18 h sedimentation (Stone, 1977)

PTE Unamended Al2((SO4))3 Removal relative control at 400 mg l-1 to control (%) Zn 50 73 48 Cu 57 90 56 Ni 19 19 0 Pb 54 96 79 Cr 22 63 193

Iron-based precipitants are marketed for use in wastewater treatment may be derived from industrial by-products of titanium oxide production. Such by-products may contain significant concentrations of potentially toxic elements (PTEs) with potentially undesirable effects on the metal content of sludge (Thiel, 1992). An example of the effects of Fe dosing with industrial by-product on the maximum potential increase in the PTE content of activated sludge is shown in Table j.2. The typical dosing -1 rates of FeSO4 are typically in the range is 15 – 30 mg Fe salt l , but may increase up to 50 mg Fe salt l-1, to comply with the discharge requirements for P in the Urban Waste Water Treatment Directive (CEC, 1991). The calculations suggest that dosing with FeSO4 may potentially increase the Cd content of activated sludge by approximately 300 % to 6 mg kg-1 ds from a typical background value of 1.5 mg Cd kg-1 ds, assuming the maximum likely dose rate of 50 mg Fe salt l-1 and that secondary sludge production is equivalent to 250 mg l-1 of total solids (UKWIR, 1997). The Ni content in activated sludge may theoretically increase by 130 % compared to sludge without Fe addition, whereas Pb and Zn concentrations may increase by about 10 % with Fe dosing. These increases in sludge content remain well within the current quality standards for agricultural use (CEC, 1986). However, the revision of the Directive on land application (CEC, 2000b) will introduce more stringent limit values for PTEs and the use of Fe salts from industrial processes could potentially penalise the acceptability of sludge for use in agriculture under the new regulatory regime. Furthermore,

202 Section 6. Case Studies the potential increase in the metal content of sewage sludge, resulting from the use of industrial-grade chemical precipitants, could also be considered as unsatisfactory because it erodes the beneficial reductions in metal inputs that have been achieved through the successful control of trade effluent discharges.

The quality and metal content of low-grade chemical precipitants for use in wastewater treatment should be examined to ensure that they do not significantly increase the metal content of sludge. In Germany, for example, composition standards are recommended for Fe and Al-based coagulants used for sewage treatment and sewage sludge conditioning (Schumann and Friedrich, 1997). The use of potable water grade Fe salts should be considered for sewage treatment (Thiel, 1992) to avoid potential problems associated with contamination with potentially toxic elements. In practice, there are few published data on the effects of chemical precipitants on sludge metal contents and Fe and Al dosing. One example from the literature (Yeoman et al., 1993) showed no consistent effects of chemical treatment with Al or Fe salts on potentially toxic elements in sewage sludge from Beckton WWTS in the UK (Table j.3). However, the significance of the direct metal inputs in chemical precipitants will increase as industrial discharges are effectively controlled and as diffuse inputs from domestic sources and run-off become the predominant sources of potentially toxic elements entering the wastewater collection system.

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

PTE FeSO4 Increase in FeCl2 salt Increase in Activated salt activated activated sludge sludge due sludge due without Fe to FeSO4 to FeCl2 Zn 348 70 26 5.2 600 Cu 5 1.0 51 10 400 Ni 160 32 120 24 25 Cd 22 4.4 3.0 0.6 1.5 Pb 64 13 22 4.4 110 Cr 32 6.4 236 47 -

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

Sludge type Salt addition Concentration (mg kg-1) ds Cd Cu Ni Pb Zn Digested None 5.4 159 24 137 231 Digested + Al Raw sludge 5.4 142 19 62 148 Digested + Al Activated sludge 4.5 254 22 168 184 Mean 4.9 198 20 115 166 Digested + Fe Raw sludge 8.5 195 40 253 300 Digested + Fe Activated sludge 4.8 95 18 155 121 Mean 6.6 145 29 204 211 (1)Sludge was collected from Beckton WWTS, sludges were digested in laboratory scale digesters (75 % raw sludge, 25 % activated sludge)

Waste products from water treatment and industrial processes, incinerator ash and have potential for re-use as P precipitants in wastewater treatment processes (Fowlie and Shannon, 1973). For example, Oostelbos et al. (1993) treated Fe-enriched sludge from water treatment with hydrochloric acid to convert ferric hydroxide to ferric chloride for use in sewage treatment for phosphate removal, and as a dewatering agent in

203 Section 6. Case Studies sludge conditioning. Verberne (1992) considered that the use of water treatment sludges as chemical precipitants for P removal was technically feasible and would depend on the agreement and acceptance of the approach by water and sewage treatment authorities. The recovery of Fe and Al from acid mine drainage is another source of chemical precipitants that can be used for P removal during sewage treatment (Bouchard et al., 1996). The re-use of secondary resources for precipitating P during wastewater treatment is intuitively attractive and also alleviates the environmental problems and impacts associated with disposal of those wastes. However, some product types derived from waste materials are potentially contaminated with potentially toxic elements that accumulate in the sludge (Fowlie and Shannon, 1973). Therefore, the metal content of waste derived products should be established, and the potential consequences for sludge quality determined, before a particular product is accepted for use as a chemical precipitant in wastewater treatment.

204 Section 7. Report Synopsis, Discussions and Conclusions

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 use through the Water Framework Directive-2000/60/EC (WFD) based on a long-term protection of water resources. A strategic aspect of this is the progressive decrease in contaminant discharges to the aquatic environment. Allied to this water policy is the amended proposal by the European Commission (COM/2001/17) listing 32 priority substances to be phased out and 11 hazardous substances that will be subject to emission 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 toxic elements and organic contaminants in sludge and metal concentrations in sludge-treated soil. The urban wastewater (UWW) collection system is one of the main facilities used for the disposal of many types of commercial and domestic wastes that contain potentially toxic elements and organic pollutants. It also receives unintentional diffuse inputs of contaminants in storm run-off from paved surfaces. However, wastewater treatment processes are effective in mitigating the discharge of many substances to surface waters with the treated effluent. This is achieved principally through aerobic biological transformations or because a significant proportion of the contaminant load is transferred to the sewage sludge. Centralised collection and treatment of urban wastewater and sewage sludge is therefore a pivotal link in defining the pathways and fate of contaminant releases to the environment.

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

The main objectives of this study were to determine the multiple sources of potentially toxic elements and organic pollutants entering UWW, to determine quantitatively the amounts 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 and organic pollutants in sewage sludge and treated effluents. This information was to be used to identify problems, to make recommendations and enable optimum reduction of pollutant inputs, and to identify areas where data are lacking and where further research is required. The significance of the inputs and types of contaminants from different sectors connected to the wastewater collection system and potential opportunities and practical measures to reduce the extent of contaminant discharges to sewer are discussed. Potentially toxic elements are routinely measured in wastewater and sludge and an extensive data-base of information has accumulated on the concentrations and fate of these elements in environmental media. Several important groups of organic contaminant have also received considerable attention (eg PAHs, PCBs, PCDD/Fs), but, in general, the amount of published information on organics is limited compared with potentially toxic elements due to the high cost of analysis, and need for specialist analytical facilities and the absence of standardised methodologies.

More information has been identified for potentially toxic elements, which have been subject to more detailed monitoring, as they are relatively easy to analyse using routine and inexpensive tests and environmental or food chain impacts are much easier to quantify where high soil loadings have occurred. Much less information is available on

Pollutants in Urban Waste Water and Sewage Sludge 205 Section 7. Report Synopsis, Discussions and Conclusions 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 lack of appropriate instrumentation, the absence of a unified methodology and expense.

The type of pollutants and the magnitude of loadings to the WWTS system are a complex 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 from anthropogenic activities is difficult to evaluate, due to the lack of information on background levels of these substances in the environment. Background concentrations relate to natural geochemical sources and biological sources, and include amounts in soils, dusts and waters derived from historical pollution.

The specific potentially toxic elements (PTEs) and organic contaminants considered in this report are listed in Table 7.1.

Table 7.1 Potentially toxic elements and organic contaminants examined in the review 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: 17b-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

Pollutants in Urban Waste Water and Sewage Sludge 206 Section 7. Report Synopsis, Discussions and Conclusions

Specific case studies on the following key issues are included in greater detail in Section 6

· 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 and sludge cannot be divorced from an understanding of their potential impacts on the environment, and particularly in relation to the use of sewage sludge in agriculture, as this provides the context for defining the significance of UWW and sludge contamination.

Finally, areas where there is a deficit of information concerning particular sources or contaminant groups, or their behaviour in the WWTS are identified and recommendations for further investigations to inform decisions about the direction of future 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 have been classified into three main categories:

· Domestic; · Commercial; · Urban runoff.

The information that was collected identified domestic inputs as the largest overall sources of Cu, Zn and Pb entering the UWW system, whereas commercial sources represent the major inputs of Hg and Cr (Table 7.2). Commercial discharges contribute moderately larger inputs of Cd (30 – 60 %) compared with domestic sources (20 – 40 %). Estimates of Ni loadings from domestic sources are highly variable and this contribution 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 significant proportion (>20 %) of the total metal load for Cd and Pb, but it is a relatively minor source of Cu and Hg. However, storm discharge was identified as an important source of Hg in a survey of metal inputs to the river Rhine (French region), contributing 15 % of the total load of this element to the river. Chromium, Ni and Zn represent an intermediate group and run-off typically contributes <20 % of the total input of these elements to UWW.

There is a degree of uncertainty in the mass balance, however, and a significant proportion of the load is from unidentified sources, which represent ³ 50 % of the total input of Cr, Ni and Zn. Unclassified inputs of Cu, Hg and Pb represent 20-40% of the total 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 to UWW is highly consistent with the estimated total loadings from these sources to major river systems in Europe (Table 7.3). Strategies to reduce discharges of metals to sewer

Pollutants in Urban Waste Water and Sewage Sludge 207 Section 7. Report Synopsis, Discussions and Conclusions can only focus on those sources that can be identified and quantified. Therefore, given the extent of the uncertainty in establishing the sources and inputs of certain elements, a priority area for research would be to determine these contributions and to develop a complete mass balance for potentially toxic elements entering to the UWW system in different European countries.

Cadmium and Hg have been proposed as priority hazardous substances by the European Commission and will be subject to the controls required by the WFD to end all releases of these elements to the water environment within a 20-year period. The status of Pb as a priority substance is currently under review. Nickel is not designated hazardous, but is classified as a dangerous substance and emission controls and quality standards will also apply to this element as required by the WFD.

Table 7.2 Provisional potentially toxic element load from different sources entering UWW in EU countries (% of total input)

PTE Domestic Commercial Urban Runoff Wastewater Wastewater Zn 30-50 5-35 10-20 Cu 30-75 3-20 4-6 Ni 10-50 30 10-20 Cd 20-40 30-60 3-40 Pb 30-80 2-20 30 Cr 2-20 35-60 2-20 Hg 4-5 50-60 1-5

The ranges in Table 7.2 are estimates from various sources in this study and may not add up to 100%. Quite a high proportion of the Zn, Ni and Cr are not identified and attributed to certain sources, and more research is needed in calculating the mass balances of these PTEs in the urban environment.

Table 7.3 Diffuse and point sources of potentially toxic elements entering the river Rhine reported in 1999 (% of total input)

PTE Point sources(1) Total diffuse sources(2) Domestic Industry Zn 30 10 60 Cu 20 15 65 Ni 20 20 60 Cd 10 10 80 Pb 15 10 75 Cr 15 25 60 Hg 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 evidence demonstrating the significant reduction in metal discharges to the WWTS, principally due to implementation of effective trade effluent controls, more efficient processes and change in industrial base. This is reflected in the declining concentrations of metals present in sewage sludge and in influent wastewaters (Section 2.1) and was reported in all countries where data could be abstracted on temporal trends (The Netherlands, France, UK and Sweden). Examples of the general declining trends in metal concentrations in sludge over the past 20 years in European countries are shown in Figure 7.1 for Zn and Cd for a major WWTP in the UK. In Sweden, Cd inputs declined by 60 % during the period 1992 – 1998 and Cr, Hg and Pb were reduced by 40 – 50 %. Zn

Pollutants in Urban Waste Water and Sewage Sludge 208 Section 7. Report Synopsis, Discussions and Conclusions

and Ni declined by 10 % during the same period, and there was no change in Cu generally reflecting the importance of the domestic contribution of these elements in UWW. Metal concentrations in rivers receiving treated UWW have also shown a marked decline. For example, concentrations in the river Thames have declined on average by 50 % in the period 1986 – 1995.

(a) Zinc (b) Cadmium 2000 2 1.8 1800

ds) 1.6 -1 ds) 1600 -1 1.4 1400 1.2 mg kg 1200 1 1000 0.8 0.6 800

Total Zn (mg kg 0.4 600 0.2 Total Cd (Log 400 0 1978 1983 1988 1993 1998 1978 1983 1988 1993 1998

Year Figure 7.1 Reduction in (a) zinc and (b) cadmium concentrations (untransformed and 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 to wastewater). Faeces typically contain 250 mg Zn kg-1, 70 mg Cu kg-1, 5 mg Ni kg-1, 2 mg Cd kg-1 and 11 mg Pb kg-1 dry solids. The residual sewage sludge from WWTS generally has significantly larger concentrations of metals than those present in human faeces because of the contributions from other sources.

For Cd, however, which is the most labile zootoxic element in UWW and sludge and its release is of particular concern in the environment, the weighted average content in sludge is comparable to the concentration in human faeces. Whilst other sources of this important element can be identified, solids inputs and secondary sludge production during WWT may have a significant dilution effect on the mass input and transfer to the sewage sludge. Consequently, a critical balance between input concentration in wastewater and dilution processes may control the final Cd content in sludge to within the normal range for faeces, which represents the minimum potential background concentration value. The precise nature of these mechanisms could be identified to better understand the behaviour of Cd in WWTS and in relation to effects of measures to reduce Cd inputs on sludge quality. Whilst the opportunities for further reducing Cd inputs would generally appear to be limited, removing phosphate from detergent formulations has successfully reduced Cd discharges from domestic sources in The Netherlands and Sweden (Section 2.1.2).

Pollutants in Urban Waste Water and Sewage Sludge 209 Section 7. Report Synopsis, Discussions and Conclusions

However, the principal sources of metals in domestic wastewater are body care products, pharmaceuticals, cleaning products and liquid wastes (eg paint). Plumbing is the main source of Cu in hard water areas, contributing >50 % of the Cu load. It may also be a principal source of Pb where this was used historically as a means of drinking water conveyance, accounting for 25 % of the input to WWTS. Lead solder in copper pipes may also be an important source.

One approach to reducing the domestic input of contaminants is for homeowners to be advised 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 the wastewater collector and regional authority responsible for municipal waste disposal to establish and publicise schemes and accessible facilities for the disposal of liquid wastes by homeowners. Metals (eg Zn) are also important active ingredients in other commonly used household products, but it may be impracticable to eliminate these although manufacturers minimise their excessive use, for example in packaging. An assessment and survey of the practical implementation of voluntary collection schemes for liquid waste and a programme of pilot studies is recommended to assess attitudes and the extent of homeowner participation.

Commercial sources and inputs of potentially toxic elements to UWW Regional studies of metal emissions from commercial premises indicate that further reductions in discharges of most elements of concern could be potentially achieved from this sector (Section 2.1.2). The main sources that may be the focus of further investigation include health establishments, manufacturing industry and hotel/catering. Approximately 30 % of medical establishments and 20 % of the other activities may be discharging significant amounts of potentially toxic elements in wastewater. In particular, Hg in UWW and sludge is largely attributable to dental and medical practices. Dental amalgam separators are effective at reducing Hg emissions where their use is compulsory and there is evidence, for example following the Danish ban, that alternatives to Hg used in thermometers and other products can reduce discharges to the WWTS.

Large industrial installations are required to treat wastewater to an appropriate standard before discharging to the WWTS and are subject to rigorous trade effluent control standards. However, small commercial, artisanal enterprises that are connected to WWTS may also be a potentially significant source of contaminants. One of the largest agglomerations of artisanal activities in Italy was investigated in Case Study (c) . This Case Study identified vehicle repair shops, metal processing and jewellery manufacturers as principal artisanal activities potentially responsible for discharging significant amounts of PTEs to the UWW system (table 7.4). However, many small enterprises pretreat wastewater before discharge and this was found to markedly reduce the 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 the total load entering the WWTS and Cr and Ni inputs were reduced to <2 and 0.5 % of the total input. Cadmium achieved the smallest overall reduction and approximately 40 % of the total load of this element to the WWTS, was identified as coming from artisanal and commercial activities, principally goldsmiths and car repairers. Further detailed assessment of wastewater management by these types of small enterprises is necessary in other European countries to determine the extent and effectiveness of wastewater treatment prior to discharge and whether specific measures and requirements are necessary and warranted to reduce potentially toxic element contamination from these sources.

Incomplete information regarding these sources could account for a major proportion of the unidentified inputs of metals to UWW (Table 7.2). The identification and control of

Pollutants in Urban Waste Water and Sewage Sludge 210 Section 7. Report Synopsis, Discussions and Conclusions these inputs is practiced in a number of regions in France by operating registration schemes and inventories of all discharging business premises/activities connected to the WWTS within specific catchment areas. Premises are inspected and the necessary levels of remedial action are agreed with the owner. Small business enterprises may be an 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 in other European countries to assess the effectiveness and practicability of targeting artisanal 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 major sources of potentially toxic elements in UWW. Oil separating devices may be fitted as a measure to reduce wastewater contamination from these activities. However, the formation of stable emulsions in the wastewater by detergent microemulsion formulations used for vehicle washing tend to limit the effectiveness of these systems at reducing pollutant emissions.

Pollutants in Urban Waste Water and Sewage Sludge 211 Section 7. Report Synopsis, Discussions and Conclusions

Table 7.4 Total metal loads (mean values) from artisanal activities (in discharged and segregated wastewater) in Northern Italy

Activity Zn Cu Ni Cd Pb Cr(III)(4) (number of % total load to % total load to % total load to % total load to % total load to % total load to enterprises(1)) WWTS (mean) WWTS (mean) WWTS (mean) WWTS (mean) WWTS (mean) WWTS (mean) Food (52) 0.28 0.09 5 % of the total input to the UWWS system are highlighted in bold and

Pollutants in Urban Waste Water and Sewage Sludge 212 Section 7. Report Synopsis, Discussions and Conclusions

Urban run-off Metal 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 of Pb from fuel additives has significantly reduced runoff inputs of this element. Precipitation represents a comparatively small proportion of the total bulk input of metals in urban run-off. Rainfall contributes approximately 10 % of the Cd, Pb and Cu and only about 1 % of the Zn load in run-off. The majority of elemental inputs are derived from dry deposition onto paved surfaces, surface degradation processes as well as from vehicle emissions.

Roads and vehicles have been investigated as potential sources of PTEs entering WWTS. Lead, Cu, Cd and Zn are present in high concentrations in brake linings and amounts 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 runoff water. Whilst these metals may be released during brake-lining wear, the amounts of Pb and Cu entering the drainage system in runoff is relatively small. Run-off is a minor source of Cu entering the UWW compared with the domestic load from plumbing. Nickel is the main element released from the abrasion of road surfaces. Roofing materials influence concentrations in runoff and Zn is significantly increased by galvanised surfaces. Lead in window frames and roof sheeting and also Pb painted surfaces can contribute significant amounts of this element to runoff water. Interestingly, the concentrations of Cd, Cu, Pb and Zn in roof runoff may be comparable to, or greater than, those in road run-off suggesting that atmospheric deposition onto paved surfaces may be more important as a source of metals in run- off than vehicle or road related inputs. Concrete structural materials transfer negligible amounts of potentially toxic elements to run-off. The relative size of the inputs of different potentially toxic elements to UWW in run-off can be ranked in increasing order as: Cd<

The available mass balance information suggests that the overall contribution of urban run-off to the metal flux in UWW is generally comparatively small, and intermittent depending on rainfall, relative to other sources (Table 7.2). However, the variable nature and uncertainty about the extent of these inputs could partly explain the apparent discrepancy between the identified sources of PTEs and the actual load entering the WWTS. The significance of potentially toxic elements in urban run-off could be quantitatively examined by relating rainfall frequency, duration and intensity with temporal trends in the metal content of sewage sludge sampled from WWTS serving different catchments types (industrial, domestic, mixed). There are a number of possible measures that can potentially reduce contaminant entry into UWW from run-off. For industrial and other paved areas, for example, where it is practicable, interception systems to remove contaminants bound to suspended solids by

Pollutants in Urban Waste Water and Sewage Sludge 214 Section 7. Report Synopsis, Discussions and Conclusions sedimentation or filtration can reduce inputs to the WWTS. However, these are only moderately effective because the smallest particles containing the largest concentrations of metals do not settle efficiently. The design and management of these systems are described in Case Study (b). In general, however, it is difficult to practically control or minimise run-off contamination due to the diffuse nature of the sources of metals, although targeting Pb painted surfaces for source preventive action would further curtail inputs of this element into the environment.

Metal transfers to sewage sludge during WWT The majority of the polluting load in UWW, including potentially toxic elements, is transferred to the sewage sludge during WWT and a relatively small soluble or suspended fraction may be discharged to the environment in the treated effluent.

Empirical and mechanistic models of metal transfer (TOXCHEM) have been developed. Both modelling approaches can be used to predict the metal concentration in wastewater on the basis of the content in sludge. Metal removal during primary sedimentation is a physico-chemical process, dependent on the settlement of precipitated, insoluble metal or the association of metals with settleable particulate matter. The removal of metals during secondary wastewater treatment is dependent upon the uptake of metals by the microbial biomass and the separation of the biomass during secondary sedimentation. Approximately 60 – 80 % of most elements in raw sewage are removed and transferred to the sludge. Nickel is the most soluble element and removals of 40 % are more typical for this element. The arithmetic mean metal concentrations in sewage sludge in European countries are listed in Table 7.5.

Mass balance of metals in the WWTS must take into account metal contents and their fates in sewage sludge. Detailed environmental, toxicological and ecotoxicological assessments of metals provide assurance that the concentrations of Cd, Cu, Ni, Cr, Hg and Pb in sewage sludge are well below values that constitute a potential 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 from phytotoxicity to affecting soil microbial processes. However, it is also an important trace element and the environment can be protected from the potentially toxic effects of Zn in sludge by technically based, but precautionary soil limits.

Table 7.5 Arithmetic mean metal content in sludge and maximum permissible limits in soil and sludge (mg kg-1ds) in the EU(1) Element Mean 86/278/EEC Proposed EU 86/278/EEC Proposed maximum in maximum in maximum in EU sludge sludge soil (range) maximum in (range) soil (pH 6-7) Zn 863(2) 2500-4000 2500 150-300 150 Cu 337 1000-1750 1000 50-140 50 Ni 37 300-400 300 30-75 50 Cd 2.2(3) 20-40 10 1-3 1 Pb 124 750-1200 750 50-300 70 Cr 79(4) 1000 60 Hg 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 in Polish sludge and sludge from Athens WWTP is 3641 and 2752 mg kg-1 ds, respectively. The mean European value including Poland and Greece is 1222 mg Zn kg-1.

Pollutants in Urban Waste Water and Sewage Sludge 215 Section 7. Report Synopsis, Discussions and Conclusions

(3)Excludes Poland; the mean Cd content in Polish sludge is 9.9 mg kg-1 value. The mean European 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. The mean European value including Greece is 141 mg Cr kg-1.

Variations in metal concentrations in sewage sludges observed between different countries are difficult to reconcile because all European governments, local environmental agencies, water utilities and municipal authorities actively enforce trade effluent control. Metal concentration data are most widely reported for sludge that is used in agriculture and lower values may be apparent where national regulations enforce stringent metal limits for sludge, although this approach tends to diminish the opportunities for utilising sludge on agricultural land. These data are therefore specific to particular sludge sources and do not provide an overall assessment of sludge contamination with potentially toxic elements disposed of by alternative routes such as incineration. However, certain countries report lower metal contents in sludge than in other states, even when sludge limit values do not appear to be generally restrictive of sludge recycling to land.

A possible explanation may be related to the statistical characteristics of metal concentrations and whether data are reported as arithmetic or weighted averages. The data are usually skewed and the median values are usually lower than the means. This would give rise to large discrepancies in amalgamated data from different member states. Variations could also be related to the efficiency of different analytical methods at extracting metals from the sludge matrix. The size of WWTP may also be important since sludges from large works contain more metals than from smaller works. This could be interpreted as the result of higher industrial inputs of metals to large works, although it may also be explained by the greater interception of atmospheric deposition of metals by paved areas in urban centres served by the largest sewage treatment works. A more detailed assessment and examination of the sludge qualities and quantities from different types of treatment centre in relation to disposal outlet in European Member States is necessary, and a consistent statistical format should be developed, for the comparative analysis of concentration data from different countries.

Enhancing metal partitioning between treated wastewater and sewage sludge The chemical treatment of wastewater to remove phosphorus is increasingly practiced to control P discharges in the treated effluent as a measure to reduce eutrophication of sensitive water courses. Chemical precipitation with Al or Fe salts can 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)). For example, Cu and Zn removal from UWW can be raised by 50 % compared to conventional sedimentation without chemical enhancement; Pb removal may increase by 80 % and a 3-fold increase in Cr (III) removal by chemical precipitation has been reported. However, potentially the most important effect of chemical precipitants on metal concentrations in sewage sludge is associated with the actual quality and metal content of commercially available precipitant formulations themselves. For example, Fe-based precipitants marketed for use in wastewater treatment can be industrial by-products from titanium oxide production and may contain 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 in sludge may increase by 130 % compared with national weighted average values and Zn and Pb contents may be typically raised by 10 %. Indeed, the use of low-grade precipitants could erode the significant progress that has been achieved in reducing

Pollutants in Urban Waste Water and Sewage Sludge 216 Section 7. Report Synopsis, Discussions and Conclusions metal emissions to sewer through controlling trade effluent discharge. In response to these concerns, certain European countries (e.g. Germany) have introduced controls on the composition of Fe and Al-based coagulants used for UWW treatment and potable water grade Fe salts are recommended to avoid potential problems associated with potentially toxic elements. However, relatively pure sources of Fe and 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 P precipitation during UWW treatment is intuitively attractive and also alleviates other potential environmental problems associated with the disposal of those wastes. However, the metal content of waste-derived products should be established, and the potential consequences of their use on sludge quality determined, before a particular product 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) and the current status of the sources and knowledge of the fate and environmental consequences of PGMs are described in Case Study (a). The commercial use of Pt and Pd in particular has expanded significantly in recent years in the manufacture of catalytic converters to reduce atmospheric emissions of carbon monoxide, hydrocarbons and from internal combustion engines. The discharge of Pt from excreted anti-neoplastic drugs used in the treatment of cancers is another identified source entering the WWTS. However, reliable quantitative estimates of major Pt and Pd sources are available and show that hospitals contribute a relatively minor input, equivalent to 6 – 12 % of Pt discharged to sewer, compared with vehicle exhaust catalysts. Glass, electronics and jewelery manufacturing are other potential sources of PGMs and can represent important local inputs of these elements entering particular WWTS. The rate of removal of PGMs from UWW by sewage treatment processes is generally within the range of most other potentially toxic element species and approximately 70 % of the Pt in wastewater is transferred to the sewage sludge. Reported concentrations of Pt in sludge from 2 WWTP in Munich were in the range 86 – 266 mg kg-1 ds.

Platinum group metals emitted as autocatalyst particles behave inertly and have limited mobility in soil so there would appear to be negligible risk to health, groundwater and the environment. However, it is possible for transformations to soluble, bioactive forms to occur and as the commercial use of PGMs continues to rise, there is a case for a limited investigation of their environmental significance. These studies should focus on the factors controlling the solubility and bioavailability of PGMs and on the behaviour of PGMs in surface waters receiving treated sewage effluents. Information is also needed on the uptake of PGMs by crops from the agricultural use of sludge and direct deposition onto urban garden soils to quantify the potential transfer by multiple exposure routes to the human foodchain.

7.1.3 Organic compounds

General Compared with the small number of PTEs of concern in wastewater and sludge which are routinely monitored and controlled, the range of organic contaminants present in these media, with the potential to exert a health or environmental hazard, are significantly more diverse. For example, approximately 140 organic compounds have been identified in UWW in Sweden and more than 330 organic substances have been determined in German sewage sludge. Forty two organic compounds are regularly detected in sludge.

Pollutants in Urban Waste Water and Sewage Sludge 217 Section 7. Report Synopsis, Discussions and Conclusions

Persistent organic contaminants Source and emission controls on persistent organic contaminants were introduced between 1980-90 to curb the extent of environmental releases as concern increased about the extent of their occurrence and potential toxicity. This legislation has been relatively effective and there are several reliable examples in the literature illustrating significant reductions in the primary sources of PAHs, PCBs and PCDD/Fs and this has lowered inputs to the UWW system and reduced concentrations in sewage sludge (Table 7.6). For example, discharges of PCBs to UWW declined by >99 % in the Rhine region of France between 1985 to 1996 due to stricter industrial source control and PAHs were also reduced by >90 % over the same time period. The principal inputs of these contaminants to UWW are from atmospheric deposition onto paved surfaces and run-off.

Controls on combustion and incineration emissions and the production of certain potentially contaminated chlorinated pesticides have also markedly reduced the release of PCDD/Fs to the environment. A review of PCDD emissions in Austria in 1998 show small consumers including household, trade and administration activities are the principal sources of release contributing almost 60 % of the total contemporary load to the environment (Figure 7.2). In the UK, incineration of municipal waste is the largest emitter of PCDD/Fs representing about 40 % of total emissions to the environment. The emission controls have had a positive effect on the quality of sewage and PCDD/F concentrations have declined significantly. For example, the average PCDD/F concentration in sludge sampled from Spanish WWTS declined from an average value of 620 ng kg-1 TEQ reported for the period 1979 – 1987 to 55 ng kg-1 TEQ in 1999. In the case of sludge from a major London WWTS (Figure 7.3), PCDD/F concentrations decreased by >97 % in the past 40 years from 166 ng kg-1 TEQ in 1960 to 4.2 ng kg-1 TEQ in 1998. Typical PCDD/F concentrations reported in sewage sludge are significantly below the highly precautionary standard for agricultural utilisation (100 ng kg-1 TEQ) proposed by the EU (Table 7.6). Indeed the numerical limit is unlikely to restrict land application yet the cost of PCDD/F analysis is high and routine monitoring of sludge for its PCDD/F content is impractical. The implementation of quality standards for PCDD/Fs in sewage sludge should be reviewed in terms of environmental effects, analysis frequency and cost.

Historic pollution levels are important for organic pollutants as well as potentially toxic elements. For example PCBs, which are no longer used, largely result from volatilisation from soil and are found in similar concentrations in all the WWTS. This suggests either that sources are diffuse and spread evenly or that this is due to general background concentrations. Soil is an effective scavenger and sorptive medium for organic pollutants and acts as a long-term and major repository, although biodegradation also takes place. Contemporary remobilisation by volatilisation from soil and redeposition onto surfaces and consequential collection by UWW systems is a major source of these compounds entering sewage sludge.

Controls on the use and emissions of persistent organic contaminants have significantly reduced industrial inputs of these substances to sewer. Therefore, the concentrations 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 (eg PCBs, PCDD/Fs and PAHs); · Atmospheric deposition from waste incineration (eg PCDD/Fs);

Pollutants in Urban Waste Water and Sewage Sludge 218 Section 7. Report Synopsis, Discussions and Conclusions

· 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 in sludge.

Polycyclic aromatic hydrocarbons are proposed as priority hazardous substances (COM(2001) 17 final) within the European Water Framework Directive and the aim will be to achieve cessation of emissions, discharges and losses of these compounds by 2020. However, curbing the emissions of PAHs and PCDD/Fs from domestic coal burning would be technically difficult and incinerators are already subject to stringent air quality emission standards. Consequently, there is probably little opportunity to further reduce the inputs and concentrations of PAHs, PCDD/Fs as well as PCBs in UWW and sewage sludge. Furthermore, the compounds are effectively removed by wastewater treatment processes because they strongly bind to the sludge solids minimising the discharge in treated effluent. The increasing amount of scientific investigation also shows there are no significant environmental consequences from PAHs, PCBs or PCDD/Fs when sludge is used on farmland as a fertiliser. In the light of such developments and their physico-chemical behaviour, it may be argued that the importance of these substances as major pollutants of UWW and sludge has been significantly diminished. Soil is a major repository for persistent organic contaminants and further investigations are necessary to improve understanding of the remobilisation and cycling processes in the environment that control diffuse inputs of these organic compounds to UWW.

Table 7.6 Concentrations and proposed limit values of selected organic contaminants in sewage sludge (mg kg-1 ds)

Organic compounds Mean content Proposed EU mg kg-1 ds maximum in sludge mg kg-1 ds Halogenated organics (AOX) 200(1) 500 Linear alkylbenzene sulphonates (LAS) 6500 2600 Di(2-ethylhexyl)phthalate (DEHP) 20 – 60 100 Nonylphenol and ethoxylates (NPE) 26 (UK: 330 – 640) 50 Total polycyclic aromatic hydrocarbons 0.5 – 27.8 6 (PAH) Total polychlorinated biphenyls (PCB) 0.09 0.8 Polychlorinated dibenzo-dioxins 36(2) 100(2) and –furans (PCDD/Fs) (1)German sludge only (2)Units ng kg-1 TEQ

Pollutants in Urban Waste Water and Sewage Sludge 219 Section 7. Report Synopsis, Discussions and Conclusions

(a) Austria (b) United Kingdom

Municipal waste incineration Small consumer Industrial

Industry - combustion Domestic

Clinical waste Industry - processes incineration Volatilisation from Waste and landfills chlorophenol Vehicle

Figure 7.2 Sources of dioxin emissions (% of total emitted) in (a) Austria in 1998 and (b) UK in 1991

450

ds) 400 350 US EPA proposed limit 300 250 200 150 German 100 limit 50 0

Dioxin concentration (ng TEQ kg 1944 1949 1953 1956 1958 1960 1998

Year Figure 7.3 Dioxin content of archived samples of sewage sludge from a WWTS in West London, UK

Other organic contaminants

Emissions of other organic contaminants and entry to the WWTS are associated with direct or indirect discharges resulting from their use in commercial and domestic activities. For example, the principal emissions of DEHP occur from the use of finished products and major domestic inputs to UWW are floor and wall coverings and textiles with PVC prints. In Sweden, for example the domestic contribution was equivalent to approximately 70 % of the total load of phthalates to Gothenburg WWTS. Numerically, detergent surfactants and residues (LAS and NPE) are the most significant contaminants in sewage sludge and the LAS content is typically more 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 detergent formulations, the proposed standard will prevent use of sludge on agricultural land. The UK also appears to have a particular problem with NP and NPEs as concentrations of these compounds in UK sludge are an order of magnitude larger than those reported in average sludges in other European countries Non-

Pollutants in Urban Waste Water and Sewage Sludge 220 Section 7. Report Synopsis, Discussions and Conclusions

Governmental organisations in Sweden and Denmark have been successful in independently introducing eco-labelling schemes persuading consumers against detergents containing LAS or NPE.

Linear alkylbenzene sulphonate and NPE can be substituted in detergent formulations and alternatives may also need to be sought for DEHP in plastics manufacture if emissions of this substance to water are to be controlled. Nonylphenol is a proposed priority hazardous substance within the Water Framework Directive and the use of NPEs in detergents is therefore likely to end with the implementation of European legislation to phase out discharges, losses and emissions of alkylphenolic compounds to the aquatic environment. A comprehensive independent review, also involving the detergent and plastics manufacturers, regarding the fate, behaviour, degradability, toxicity and environmental consequences of these and the alternatives compounds would provide information on the advantages and disadvantages of product substitution in detergent formulations and plastics manufacture.

A group of emerging compounds of potential significance in UWW were identified by the review due to their persistence during sewage or sludge treatment, persistence in soil or toxicity in the environment:

1. Little is known about the fate and behaviour in UWW of the large number (>200) of commercial chlorinated paraffin formulations in use as plasticisers in PVC and other plastics, extreme pressure additives, flame retardants, sealants and paints.

2. The brominated diphenyl ethers (PBDEs) are a group of compounds used for flame retardation in furnishings, textiles and electrical insulation and their use has expanded due to fire regulation requirements and the increased use of plastic material and synthetic fibres. A survey in Sweden indicated concentrations of PBDE congeners in sludge were between 15 and 19 mg kg-1, respectively. PBDEs are on the list of proposed priority hazardous substances.

3. Polychlorinated naphthalenes (PCNs) are released into the environment by waste incineration and landfill disposal of items containing PCNs. Concentrations in sewage sludge may be in a similar range to individual PCB congeners. A survey of Swedish 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 TCDD toxic equivalent values similar to those for coplanar PCBs and so have toxicological interest.

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 has low water solubility and half-lives in soil are reported to be in the range 5 – 10 months.

5. Polydimethylsiloxanes (PDMS) are nonvolatile silicone polymers used in industrial and consumer products including lubricants, electrical insulators and antifoams. They are hydrophobic and partition onto the sludge solids and concentrations 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 perfumed products (Case Study (e)). In Europe, current consumption is estimated to be 124 t y-

Pollutants in Urban Waste Water and Sewage Sludge 221 Section 7. Report Synopsis, Discussions and Conclusions

1 for musk ketone and 75 t y-1 for musk xylene and the release of these compounds to the environment is dominated by domestic discharges to WWTS.

Further studies on the biodegradation and transformations of these compounds during secondary wastewater and sludge treatment are required as a preliminary stage in understanding whether the release of musk compounds to the environment is significant from WWT.

7. Endogenous oestrogens (17b-oestradiol and oestrone) and synthetic steroids such as ethinyloestradiol, which is the active oestrogenic component in oral contraceptives, are discharged in trace amounts in the effluent from WWTS and are primarily a concern due to the possible impact on the aquatic ecosystem. These compounds also partition onto particulates and may be associated with sewage sludge, but this is unlikely to have significant environmental implications for use of sludge in agriculture. Recent investigations show that approximately 90 % of potential oestrogenic activity (based on 17b-oestradiol equivalency) in UWW is reduced by sewage treatment and that <3 % may be transferred to the sewage sludge.

8. Pharmaceutical compounds, designed for specific biological effects in medical and veterinary practice, can enter the wastewater system by excretion or residues and metabolites in urine and from intentional disposal (Case Study (d)). Disposal into the UWW system is common practice, although the rationale for justifying this on the grounds of the dilution received within the sewer system should be reviewed in context 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 the wider environment are generally unknown, but, because of their biological function, they are generally designed to be rapidly metabolised and degraded, although they are often lipophilic and potentially bioaccumulate. Removal efficiencies during WWT of >60 % are typical for most types of drug, although a wide range of removal rates is reported (7 – 96 %) and drug compounds are routinely found in treated effluents and surface waters. Many commonly used analgesic drugs are rapidly biodegraded during sewage treatment including aspirin, ibuprofen and paracetamol. Most of these are soluble and exist primarily in the aqueous phase and transfer to sewage sludge is probably of only minor concern, although it is not possible to predict partitioning and fate during sewage treatment due to the absence of physico-chemical data for many of the compounds. The occurrence of drugs in soil is widespread from veterinary administration to livestock. A risk assessment on the fate and biodegradation of drug compounds could form an integrated part of the approval procedure for pharmaceutical products. This would improve understanding of their behaviour and fate to balance the benefits to health with the potential consequences of their release into the environment. Collection systems for unwanted drugs should be encouraged to reduce disposal into the wastewater system.

Polyelectrolytes and sludge treatment Polyelectrolytes based on polyacrylamide and cationic copolymers are used extensively in sludge treatment to aid dewatering, see Case Study (g). The polyelectrolyte concentration in mechanically dewatered cakes is relatively large typically in the range 2500 – 5000 mg kg-1 ds and they only degrade relatively slowly by abiotic processes in cultivated soil at a rate of 10 % per year. Acrylamide is a common monomer associated with polyelectrolytes and is potentially toxic to humans and is a reported carcinogen. Concern that residual monomers in polyelectrolytes used in drinking water treatment may have implications for human health has

Pollutants in Urban Waste Water and Sewage Sludge 222 Section 7. Report Synopsis, Discussions and Conclusions resulted in their withdrawal from this application in Japan and Sweden, and stringent controls on their use are also in place in Germany and France. These factors have drawn attention to the possible environmental implications of sludge dewatering practices with polyelectrolytes and long-term accumulation in sludge-treated soil.

Attenuation and transformations during wastewater and sludge treatment Sewage sludge is treated to reduce its fermentability, nuisance and vector attraction and to aid its management and acceptability for use on agricultural land. These processes include physical, chemical and microbiological treatment of the sludge that may 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 by physico-chemical processes and can be predicted for individual compounds by the octanol-water partition coefficient (Kow). During primary sedimentation, hydrophobic contaminants may partition onto settled primary sludge solids and compounds can be grouped according to their sorption behaviour based on the Kow value as follows:

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

Many sludge organics are lipophilic compounds that adsorb to the sludge matrix and this 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 and the dichlorobenzenes, may be lost by volatilisation during wastewater and sludge treatment at thickening, particularly if the sludge is aerated or agitated, and by dewatering. 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 losses of specific organic compounds during sewage treatment can be predicted based on Henry’s constant (Hc) and Kow:

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

Biodegradation may be more important than air-stripping in removing volatile organic contaminants during secondary biological WWT.

Mesophilic anaerobic digestion is the principal sludge stabilisation process adopted in most European countries and many organic contaminants are biodegraded under anaerobic conditions and this is enhanced by increasing retention time and digestion temperature. Biodegradation during anaerobic digestion may eliminate certain organic contaminants from sewage sludge, but in general the destruction achieved is typically in the range of 15 – 35 %. Aromatic surfactants including LAS and 4- nonylphenol polyethoxylate (NPnEO) are not fully degraded during sewage treatment and there is significant accumulation in digested sludge. For example, mass balance calculations suggest that approximately 80 % of LAS is biodegraded during the activated sludge process and 15-20 % is transferred to the raw sludge, although near complete biodegradation (97-99 %) is also reported in some cases. Approximately 20

Pollutants in Urban Waste Water and Sewage Sludge 223 Section 7. Report Synopsis, Discussions and Conclusions

% of LAS in raw sludge may be destroyed by mesophilic anaerobic digestion. The compounds, nonylphenol monoethoxylate (NP1EO) and nonylphenol diethoxylate (NP2EO) are formed during sewage treatment from the microbial degradation of NPnEO. These metabolites are relatively lipophilic and accumulate in sludge and are also discharged in the treated effluent from WWT. Approximately 50 % of the NPnEO in raw sewage is transformed to NP by sludge digestion.

Lower molecular weight phthalate esters and butyl benzyl phthalate are completely degraded in 7 days by anaerobic digestion at 35°C and should be removed by most municipal anaerobic digesters. The extent and rate of biodegradation of organic compounds during anaerobic digestion is apparently related to the size of alkyl side chains and compounds with larger C-8 groups are much more resistant to microbial attack. Therefore, di-n-octyl and DEHP are more persistent and are not removed by conventional anaerobic treatment of sludge. Phthalate esters are rapidly destroyed under aerobic conditions, however, and biological WWT (eg activated sludge process) can usually achieve >90% removal in 24 h. In soil, the reported half-life of DEHP is <50 d.

Composting is a thermophilic aerobic stabilisation process and has the potential to biodegrade relatively persistent organic compounds in sludge. Thermophilic aerobic digestion processes and sludge storage for three months can achieve similar overall removal rates for organic contaminants as those obtained with mesophilic anaerobic digestion. Thermal hydrolysis conditioning of sludge prior to conventional anaerobic stabilisation may have a significant influence on the removal of organic contaminants from sludge, but this is a comparatively new enhanced treatment process and effects on the destruction of organic contaminants have yet to be investigated.

Pollutants in Urban Waste Water and Sewage Sludge 224 Section 7. Report Synopsis, Discussions and Conclusions

Table 7.7 Assessment of the significance of PTEs entering UWW and SS

PTE Trade (1)Priority (2)Environmental (3)Significance of input sources (L, Low; Moderate, M; High, H) effluent hazardous significance in (4)Commercial Run-off Domestic control substance (Yes, sludge-treated Relative Opportunity Relative (5)Opportunit Relative (6)Opportunit (Yes, Y; Y; soil (short importance to reduce importance y to reduce importanc y to reduce No, N) No, N) –term) e (Yes, Y; No, N) Zn Y N N* M M L - M L H M Cu Y N N L - M M L L H L Ni Y N N M M L – M L M L Cd Y (7)Y N M - H M L – M L (8)M L Pb Y (Y) N L – M M M L H L - M Cr Y N N M - H M L – M L L L Hg Y Y N H H L L L L Pt N N N M L H L L L Pd 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 Council establishing 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 maximum limit. (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 as sources 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 reduce pollutant inputs bound to suspended solids. (6)Mainly voluntarily through product ecolabelling and education concerning appropriate use and disposal of liquid wastes, body care products, cleaning agents and detergents by homeowners. A reduction in Pb may be possible in the long-term by replacing historical leaded pipework used for water conveyancing. (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 roofing materials. (8)Human faeces are the main domestic source of Cd representing the background concentration in UWW and sewage sludge.

Pollutants in Urban Waste Water and Sewage Sludge 225 Section 7. Report Synopsis, Discussions and Conclusions

Table 7.8 Assessment of the significance of organic contaminants entering urban wastewater and sewage sludge

Contaminant (1)Content in (2)Priority (3)Destruction in treatment Accumulation Background (4)Overall WW/sludge hazardous (Yes, Y; No, N) inputs significance substance (Yes, Y; No, N) Wastewater Sludge Biological Soil (Yes, Y; No, N) Wastewater (5)Sludge LAS H N H L Anaerobic N N N H L H Aerobic NPE M - H Y M L Anaerobic N N N H L H Aerobic DEHP M (Y) M L Anaerobic N N N H L M Aerobic PAHs L - M Y L L Y Y Y L L PCBs L N L L Y Y Y L L PCDD/Fs L N L L Y Y Y L L Pharmaceuticals L N M M N N N M L Oestrogenic: Endogenous L N M - H M - H Y N Y H L Synthetic 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 highly variable, 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 of water 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, H LAS Linear alkylbenzene sulphonates NPE Nonylphenolethoxylates DEHP Di-(2-ethylhexyl)phthalate PAHs Polycyclic aromatic hydrocarbons PCBs Polychlorinated biphenyls PCDDs Polychlorinated dibenzo-p-dioxins PCDFs Polychlorinated dibenzo-p-furans

Pollutants in Urban Waste Water and Sewage Sludge 226 Section 7. Report Synopsis, Discussions and Conclusions

Table 7.9 Sources and control of organic contaminants entering urban waste water and sewage sludge

Contaminant Commercial Run-off Domestic Relative Opportunity Relative Opportunity Relative Opportunity importance to reduce importance to reduce importance to reduce LAS H M L L H M NPE H M L L H L DEHP H M L L M M PAHs L L H L L L PCBs L L H L L L PCDD/Fs L L H L L L Pharmaceutical H M L L H M LAS Linear alkylbenzene sulphonates NPE Nonylphenolethoxylates DEHP Di-(2-ethylhexyl)phthalate PAHs Polycyclic aromatic hydrocarbons PCBs Polychlorinated biphenyls PCDDs Polychlorinated dibenzo-p-dioxins PCDFs Polychlorinated dibenzo-p-furans

Pollutants in Urban Waste Water and Sewage Sludge 227 Section 7 - Discussion and Conclusions

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 been reviewed and opportunities for further reducing metal inputs are considered. This may be possible through, for example, controlling discharges from a number of specified commercial sectors and the introduction of remedial measures to reduce inputs from certain domestic and diffuse sources. Therefore, a feasibility study of the most effective approaches to further reducing metal inputs is recommended so that appropriate remedial strategies can be implemented.

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 is more widely implemented and tightened, while the percentage contribution from domestic sources is expected to increase. This is important as it is less easy to regulate domestic discharges to UWW. However this can be achieved to some extent by controlling the products used in homes; for example cadmium emissions from households could be reduced by a ban on the use of phosphates in washing powders. Increasing consumer awareness of pollution in UWW, for example through eco labelling schemes, can lead to reductions in load to UWW from some sources such as the domestic use of LAS, however this requires much wider adoption by manufacturers of eco labelling schemes. Industry and stakeholder involvement need to be encouraged. Mechanisms to provide an incentive for reductions in pollution and alternative product development should be developed.

There is probably little scope for further reductions in the inputs and concentrations of persistent organic contaminant types in sewage sludge. For example, it is unlikely that much can be done to curb emissions of PAHs and PCDD/Fs from domestic coal burning, where this is widely practiced, and incinerators are already subjected to stringent air quality emission standards. Losses of PCBs from electrical transformers is rapidly declining as these are phased out of use. However, the importance of soil as a repository for persistent organic contaminants is emphasised and the focus of further investigations to minimise inputs to sludge could be directed towards better understanding of the remobilisation and cycling 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 necessary and should be implemented at a European level as soon as feasible.

Feasibility of of treated wastewater for non-household purposes, such as for irrigation of agricultural crops, parkland, and golf courses needs to be considered but must involve assessment of risks to human , environmental and economic strategic issues. Reuse of domestic grey waters in the EU needs to be re-evaluated. Development of common standards are necessary.

Return of unused pharmaceuticals to dispensaries/hospitals needs to be encouraged and made more widespread throughout Europe. Hospitals and medical centres effluents should be monitored closely for discharge to the urban wastewater network. Possible segregation and pre-treatment of hospital, medical centre and laboratory effluents, for the reduction of PGMs, pharmaceuticals and potentially toxic elements is recommended.

Sustainable urban drainage, discussed in detail, Case Study (b), may have great potential in terms of reduction of urban runoff pollution. However it is recommended that each case should be assessed individually, and an incremental approach containing both high tech and low-tech solutions is the most likely development scenario (Butler and Parkinson, 1997).

Financial incentives/grants to encourage and facilitate removal of lead piping from households would reduce lead emissions to UWW and should initially target areas with soft water.

Pollutants in Urban Waste Water and Sewage Sludge 228 Section 7 - Discussion and Conclusions

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 to agriculture. Urban use of pesticides and herbicides in gardens and allotments is often indiscriminate and can result in major uncontrolled inputs to the UWWS.

For mercury, dentists are important sources and improved dental practices may reduce this source, 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 as a replacement for mercury. Mercury recycling schemes have also been a success in many regions and could be extended to other potential pollutants used in domestic and commercial settings.

Adjusting the pH of tap water may be useful in reducing corrosion, for example, from domestic and commercial heating systems, but may be limited by practical and economic factors and it can still be a problem in drinking water in some regions. For example, lead passing into drinking water supplies from old lead piping and then into the sewage system can be controlled by adjusting the pH and hardness of the .

This report benefits from the presentation of several specific case histories, which provide detailed information in support of the above conclusions. Some specific potential problem areas have been reviewed. For example, the differences in cadmium concentrations reported in sewage sludge in the United Kingdom and Germany have been shown to be due to different approaches to data treatment and presentation rather than real variations in composition. (see Section 2.3)

In the WWTS the fate of compounds often differs, for example, Ni is removed less well from treated wastewater than other metals. This means that more Ni will pass out in the effluent from the WWTS and less will be concentrated in the SS. The behaviour of compounds should 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 and copper, are essential trace elements to plants and animals in low concentrations. It is important that levels in treated UWW and SS are not set so low by regulation that mineral deficiency could arise due to removal of these sources where UWW and SS might have been used, for example, on nutrient deficient land.

Domestic and industrial product formulation can have an important role in determining the nature and amount of pollutants that are likely to enter the WWTS. Risk assessment of domestic products should be based on their input of pollutants to the urban wastewater and the biodegradability of the pollutants. Health and environmental effects should be considered in the risk assessment process. Efforts to remove phosphate in detergents have had a knock on effect in lowering the cadmium input to the WWTS. NPE bans and regional LAS reduction through eco-labelling schemes have also been proven effective. It would also be possible to pre-treat high LAS concentration in UWW from commercial sources, which would then reduce levels reaching the WWTP.

Advice on household refurbishment (e.g. for old lead paint, piping) and advice on disposal of potential pollutants (e.g. down sinks) and eco-labels and public education should be provided and extended where possible to raise awareness of ecological impacts of various processes and products in urban wastewater.

The inclusion of Cu, Zn and LAS on the proposed list of priority substances is recommended because they are the most limiting substances to utilising sludge in agriculture within the limits proposed in the revision of Directive 86/278/EEC and they are ubiquitous in wastewater. Therefore, these substances should also be the focus of measures to reduce

Pollutants in Urban Waste Water and Sewage Sludge 229 Section 7 - Discussion and Conclusions discharges and emissions to further improve sludge quality and support the recycling route for sewage sludge.

In order to reduce inputs to UWW, it is important to target the sources contributing most to the system, especially the diffuse sources. Trade effluent discharge controls have been successful in lowering emissions but are still necessary for certain industries. Reductions of many pollutants including lead and dioxins have been seen in sewage sludge since the 1960s. Increasing control on emissions, not only to water but also to air and land, are likely to continue this reduction. This study concludes that small users, hospitals, garages and car washes, dental and medical practices need a close monitoring and control for connection to the urban wastewater systems.

Pollutants in Urban Waste Water and Sewage Sludge 230 Section 7 - Discussion and Conclusions

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, in some 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 and this is identified as a priority area for further data gathering.

There are very few studies available on the mass balance of potentially toxic elements and organic pollutants through WWTS. More work is needed to understand the partitioning and speciation of pollutants, particularly for compounds, such as NPE, which may become more toxic through wastewater treatment. Some pollutant sources still need further identification. In fact there are also many instances where the majority of pollutants sources remain unidentified, for example the ADEME study in France estimated that over 50% of some of the metals came from unidentified sources.

Identification and monitoring of trace organics is still sparse and more detailed work, though costly, is urgently required. Better source inventory data is essential. In some cases specific industrial sources are difficult to locate. For example, platinum uses and losses from industry 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 is needed for estimates and quantification of diffuse sources of organic pollutants especially in Iberia, Italy and Greece. Some data especially from Iberia is not readily available and an integrated data collection system is needed across the EU 15 for PTEs and organic pollutants.

Information is still needed in order to asses health and environmental effects. UWW effluent and sewage sludge use on land usually involves exposure to very small quantities of mixed pollutants over a prolonged period of time, and there is little health data about chronic exposure, particularly to organic pollutants. A particular area of current concern is the possible impact of cocktail effects where several contaminants are present at the same time. Interactions between metals, organics, and metals and organics may be synergistic or antagonistic, are complex and far from clearly understood.

Another area where knowledge is lacking is that concerning the effects of both potentially toxic elements and trace organics on components of the ecosystem, both in soils and in surface and subsurface waters. The prime effects to be considered are those on sensitive receptors, including micro-organisms, invertebrates and plants. At present there is a lack of base line information in this area. Ecotoxocity tests that are applied are undertaken in laboratory conditions assuming 100% bioavailability of the pollutant, and are not appropriate for field conditions. Research into the transfer of organic pollutants through uptake into pasture plants and into crops and thus into the food chain is limited and should be considered.

Pollutants in Urban Waste Water and Sewage Sludge 231 Appendix A

APPENDIX A URBAN 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 and also 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 the admission of surface water to a sewer or sub-drain. It has a sediment sump at its base to trap grit and detritus below the point of overflow.]

Pollutants in Urban Waste Water and Sewage Sludge 232 Appendix A

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 by wastewater in solution or suspension in order to render water fit and safe for discharge and intended uses. Stringent water quality and effluent standards have been developed that require reduction in suspended solids, biochemical oxygen demand (BOD, related to biodegradable organic compounds), COD () and to some extent coliform organisms (indicators of faecal pollution), control of pH as well as the concentrations of certain organic compounds, together with some potentially toxic elements and non-metals.

Sewage sludge consists of residues originating from mechanical, biological, chemical and physical treatment of wastewater in sewage plants. The quantity and nature of the arising sewage sludge are subjected to strong fluctuations depending on the wastewater composition, the kind of wastewater purification process and the purification degree. Two different sewage sludge types can mainly be distinguished:

· Primary sludge: becomes physically or chemically separated from wastewater in primary treatment · Secondary sludge: arises from the biological step (surplus activated sludge, sewage sludge from trickling filters) and tertiary treatment (often nutrient removal).

Primary and secondary sludges are usually combined to create a composite sludge, which often goes for further treatment in sludge digestion and dewatering. Residues resulting from screening in preliminary treatment are not considered as sludge, consisting mainly of coarse solid particles, grits, and grease [Magoarou, 1998].

Although the ‘nitrate’ and ‘phosphate’ ions in sewage are beyond the realm of this project, it has to be noted that pollutants such as potentially toxic elements, organics, sulphides and residuals, can form insoluble phosphates in the course of treatment processes. These have great propensity for sedimentation in all stages of the sewage treatment process, thus becoming part of the sewage sludge that has to be managed.

Figure A.2 shows the schematics of urban wastewater and sewage sludge treatment.

Preliminary

Primary Advanced

Secondary Treatment Secondary

Treatment

Treatment Treament

Secondary

Nutrient Removal, Nutrient

Sceening, Physical/Chemical Sludge Activated

Settlement

Colour Removal Colour

Girt Removal Girt Solid and Liquid

Basin

& Grease Trap Grease & Separation

SECONDARY TERTIARY

PRIMARY

RESIDUES

SLUDGE SLUDGE

SLUDGE

Sludge Digestion Sludge

Anaerobic/Aerobic

Sludge

Dewatering

Mesophilic/Thermophilic

Conditioning

Composting

Stabilisation ponds Stabilisation

Advanced Treatments Advanced Land Application Land

Figure A.2 Schematic Urban Wastewater and Sewage Sludge Treatment

Pollutants in Urban Waste Water and Sewage Sludge 233 Appendix A

REGIONAL ASPECTS:

Protection of the receiving waters from pollution by harmful effluent is the primary goal for the treatment of urban wastewaters at WWTS. Urban wastewater is defined by the Council Directive 91/271/EEC of 21 May 1991 concerning urban wastewater treatment (and as amended by Commission Directive 98/15/EC of 27 February 1998) as, domestic wastewater or 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 across the European Union. The water supply (litres/inhabitant/day) fluctuates greatly accross the European urban areas (EUROSTAT data,1998-2000, Ginés, 1997). Not all collected wastewater is treated; the percentage of urban wastewater not receiving treatment ranges from 3% in Germany, up to 77% in Greece. This may include unplanned wastewater collection, such as septic systems or leakage of UWW collection systems. Cities such as Milan and Brussels do not yet have a centralised WWTS.

Table A.1 Population and household access to sewerage and wastewater treatment facilities Country Population Percentage of population with access to sewerage and public WWT (1998) facilities thousands Access to No treatment P P + S P + S + Year sewerage T (%) (%) (%) (%) (%) Austria 8,075 76 1 1 39 35 1995 Belgium 10,192 78 37 1997 Denmark 5,295 87 18 1996 Finland 5,147 78 0 7 71 1997 France 58,727 81 4 0 0 77 1994 Germany 82,057 92 >31 >47 1995 Greece 10,511 70 16 0 0 54 1996 Ireland 3,694 68 32 23 12 1 1995 Italy 57,563 84 16 3 38 26 1996 Luxembourg 424 88 0 19 57 11 1995 Netherlands 15,654 98 2 0 68 28 1994 Portugal 9,957 55 34 9 11 0 1990 Spain 39,348 62 13 11 34 3 1995 Sweden 8,848 86 0 0 5 81 1995 UK 59,090 96 9 9 64 14 1996 - England & - 96 10 0 0 86 1995 Wales -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=tertiary treatment

Table A.1 shows the data in EU15 for the total population, population access to UWW collecting systems and to urban wastewater treatment facilities. The EU15 average shows that 84 % of the households have access to UWW collecting systems, whereas only 48 % of the urban wastewater is treated in primary + secondary + tertiary treatment facilities.

According to the European Waste Water Group (1997) report on urban wastewater treatment in the EU and accession countries, there are marked differences between various European regions in terms of primary, secondary and tertiary treatment as shown in Figure A.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, , Hungary, Lithuania, Latvia, Poland, Romania, Slovenia, Slovakia).

Pollutants in Urban Waste Water and Sewage Sludge 234 Appendix A

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%)

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 need for greater investment in the urban wastewater treatment in the southern region and the accession countries.

Figure A.3 Urban wastewater treatment in EU and accession countries (minus Cyprus) [after EEA, 1999, chapter 3.5 Water Stress]

Pollutants in Urban Waste Water and Sewage Sludge 235 Appendix A

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 the environmental damage experienced in the past decades in this region from trans-boundary acid deposition due to the burning of high sulphur fossil fuels. A number of national pollution events, such as of fish due to discharges from chloralkali plants also increased awareness of the damage emissions can cause.

This led to increased recognition of the need for national and international agreements to limit local and global environmental perturbations and has led Sweden and Norway in particular to be among the most pro-active nations in terms of setting environmental standards and education amongst the general population.

Finland, Norway and Sweden have low population densities and there are similarities in the commercial activities in these regions, for example the oil production, metals and engineering and paper manufacturing industries. Denmark has a higher population density than other countries in the Northern region and therefore more of an urban environment but also has a high level of environmental concern about water pollution, although all the drinking water in this region is from groundwater sources. This compares with the UK, for example, where approximately one third of potable drinking water is abstracted from surface water 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 and for Central and Eastern European countries (Bulgaria, Czech Republic, Estonia, Hungary, Latvia, Lithuania, Poland, Slovakia, Slovenia), all associated EU countries. Environmental awareness in the Central Region is high in Germany, Austria and Switzerland whereas the CEE countries experience a high level of environmental damage due to the long years of neglect while part of the Soviet bloc. Transboundary pollution issues are very important in the accession CEE countries and their neighbours. Very little data is available for the CEE countries regarding the pollutants in UWW and SS. An inventory of organic pollutants in the environment in the CEE countries was done by a team at Brno University, the Czech Republic. (Persistent, Bioaccumulative and Toxic Chemicals in Central and Eastern European Countries - State-of-the-art Report-TOCOEN REPORT No. 150, 1999). The main organic pollutants investigated are the PAHs, PCBs, PCDD/PCDFs and their ocurrence in the urban environments in the CEE countries. Most data available is from the Czech Republic, 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. Water resources are limited in the Southern Region and therefore recycling of treated waste water and pollution reduction at source are important issues. Data regarding sources of pollutants is less abundant in the 'Southern Region' especially regarding specific organic pollutants. The region has among the lowest percentage of population access to UWW collecting systems in the EU (70% for Greece, 84% for Italy, 55% for Portugal and 62% for Spain). The percentage of non-treated wastewater is among the highest in the EU (16% for Greece and Italy, 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 of domestic origin, with negligible contribution from urban and industrial wastewater. Data for Southern Italy were gathered in a two years pilot campaign [Braguglia et al., 2000; Marani et al., 1998; Mininni et al., 1999].

Pollutants in Urban Waste Water and Sewage Sludge 236 Appendix A

Metal Concentration range Southern Italy Literature As 1.1-1.8 0.3-20 Cd 3-23 1-50 Co 1-4 5-30 Cr 227-535 40-1500 Cu 258-373 160-1600 Hg 1.1-3.1 1-12 Mn 80-109 240-600 Ni 34-57 20-240 Pb 95-137 80-850 Zn 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 following requirements 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 to groundwater) out of which 240,000 are to UWW collecting systems [Ministerio de Medio Ambiente, 1997]. However, the quantification of the pollutants discharged is very limited. Routine controls are generally limited to those established in Royal Decree 509/1996 and which must be published every two years: BOD5, COD, suspended solids, total phosphorus and total nitrogen in the case of treatment plants located in sensitive areas. However, these controls are usually employed only for the discharges (not for the wastewater coming into the treatment plant). As the discharges into the UWW collecting systems are under the control of the local governments (municipalities), the limited information on levels of pollutants in wastewater that exists comes from the enforcement inspection controls and is not 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 urban wastewater and sewage sludge in Iberia [Palerm and Singer, 2000]. Water management in Spain is completely de-centralised and the discharges into the UWW collecting system are under the competence of the municipal authorities, which must meet the established parameters for discharge into the public waterways.

There is very little information sources of pollutants and levels of pollutants in urban wastewater and sewage sludge in Portugal [Palerm and Singer, 2000]. Most of the information that exists was obtained prior to designing the wastewater treatment plants, but this 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, and more limited for pollutants in urban wastewater. The Hellenic Ministry for Environment, Physical Planning and Public works co-ordinates the data management in terms of pollutants load in urban areas. The research studies tend to centre on the cities of Thessaloniki or Athens.

Pollutants in Urban Waste Water and Sewage Sludge 237 Appendix A

The Western Region (UK, Ireland, France and the Benelux countries, Belgium Netherlands and Luxembourg):

In this region, wastewater treatment is among the most advanced in the EU. France has over 160,000 km of sewers and 11,300 treatment plants. The WWTS are operated by public or private bodies. The ownership of WWTS and sewers are always public, owned by municipalities 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 the sewer systems and often, private companies run the plants [EWWG, 1997].

In the Netherlands, the municipalities are responsible for collection of sewage and disposal of sludge from sewers. The treatment of wastewater and disposal of effluent and sludge is the responsibility of 27 waterboards. Sludge disposal is partially privatised. In Belgium, the municipalities are responsible for the sewerage systems in both Vlanderen and Wallonie. Similarly, in Luxembourg, the 118 municipalities are responsible for the collection and treatment of urban wastewater and sludge. The management of sludge is partially shared with operators of solid waste [EWWG, 1997].

In UK there are over 300,000 km sewers and 7,600 WWTS. The collection of sewage, its treatment and disposal of effluent and sludge are the responsibility of privately-financed water service companies in England and Wales (10), public water and sewerage authorities in 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 is the responsibility of local councils. Private-Public Partnerships (PPP) schemes are in place for modernising and upgrading the WWTS [Dept. of Environment and Local Govt, Ireland, 1999]

Pollutants in Urban Waste Water and Sewage Sludge 238 Appendix B

APPENDIX B Physical and chemical properties of selected pollutants

In order to understand the behaviour of organic pollutants in the urban environment the knowledge of their physical and chemical properties is very important. Vapour pressure, boiling-point, water solubility and distribution coefficients describe the distribution between solid, liquid and gaseous phase. The adsorption coefficient KOC is important for transitions between soil or sewage sludge particles and water. The transition water/air is governed by the Henry coefficient KAW. The distribution coefficient octane/water KOW is a measure for the lipophilic or lipophobic qualities of a chemical compound. For substances with similar physical 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 in detergents, washing-up liquids and cleaning agents. The most important LAS qualities are represented in Table II.22 Figure B.1 shows the LAS structure formula.

Table B.1 LAS properties [Bürgermann 1988]. Abbreviation Molecular Molar mass Colour Solubility Vapor Density formula pressure at 20°C [g/mol] [g/l] [g/cm3]

LAS C18H30O3S ca. 326 colourles 1.1 extremely low 1 s

CH3-CH2-CH-(CH2)8-CH3

+ SO3-Na Figure B.1: LAS structure formula.

A2. NPEs or nonylphenol polyethoxylates and APEs alkylphenol ethoxylates

Polyethoxylated nonylphenols are important surfactants used comercially and in some household products for many years, also as emulsifiers and solubilisers in industrial processing, 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 comparable physical, chemical and biological qualities. They differ from each other in the position and number of chlorine atoms and the symmetry of the basic structure. Figure B.2 shows the structure formulas of the polychlorinated dibenzo-p-dioxins and polychlorinated dibenzo furans (PCDD/F).

239 Appendix B

9 1

8 2

7 3

Clx Cly 6 4

PCDD 9 1

8 2

7 3

Clx Cly 6 4

PCDF Figure B.2 PCDD/F structure formulas. PCDD/F are extremely heavy-volatile components (high melting and boiling-points). The vapour pressure drops with chlorine substitution, so that low-chlorinated dioxins and furans are quite volatile. Highly-chlorinated compounds are found in solid state adsorbed on particles. The greatest mobility is in the air. In water, PCDD/F are almost completely bound to particles (high octane/water distribution coefficient). The affinity to organic carbon compounds 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 thermally stable at temperatures up to 600-800 °C. The extraordinary stability and the low photolytic reduction render them to be very persistent in the environment. Because of the long life time and the lipophilic qualities PCDD/F can accumulate in organisms: they accumulate predominantly in animal fatty substances [Mahnke, 1997].

C. Polychlorinated Biphenyls (PCBs)

The polychlorinated biphenyls (PCBs) form a group of over 209 chlorinated, aromatic compounds with the same structural features. They differ with the degree of chlorine substitution 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 high lipophilic character, so they accumulate in the food-chain. Their biological degradability in the environment depends on the complexity and chlorination of each particular compound. Figure B.3 shows the PCBs structure formula. The physical-chemical qualities of selected PCBs compounds are specified in Table B.2.

3 2 2' 3'

4 4'

Clx 5 6 6' 5' Cly Figure B.3 PCBs structure formula.

240 Appendix B

Table B.2 Physical-chemical data of selected PCBs. Substance Henry Solubility Log KOW constant [g/m3] 2,4,4´-Trichlorbiphenyl n.a. 8.5E-02 5.74/5.69 2,5,2´,5´-Tetrachlorbiphenyl 4.9E-03 4.6E-02 6.26/6.09 3,4,3´,4´-Tetrachlorbiphenyl 3.1E-04 1.8E-01 6.52/5.62 2,4,5,2´,5´-Pentachlorbiphenyl 1.2E-0.3 3.1E-0.2 6.85/7.07 2,3,4,2´,4´,5´-Hexachlorbiphenyl 5.3E-04 n.a. n.a./7.44 2,4,5,2´,4´,5´-Hexachlorbiphenyl 6.9E-04 8.8E-03 7.44/7.75 2,3,4,5,2´,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 their ubiquitous occurrence and high carcinogenic potential. PAHs are multi-core aromatic ring systems with 5 and 6 rings. They represent benzene condensation products. PAHs are solid, mostly colourless compounds. They have a strong lipophilic character and their water solubility decreases with the increase of ring numbers. Low-molecular PAHs are relatively volatile. PAHs with a boiling-point below 400°C exist in the air in gaseous state. The higher boiling compounds are adsorbed to particles. The physico-chemical features of some PAHs are 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 Vapour Solubility KOW KOC Henry constant formula pressure [Pa] [mg/l] [cm3/g]

Benzo(k)fluoranthene C10H12 0.1E-07 0.68E-06 6.84 2843420 0.2E-05 Benzo(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)

241 Appendix B

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 Solubility in Kow Carcinogeni pressure water (mg.l-1) c potency (Torr at 20°C) IARC/USEP A* classification Acenaphthene, Ace 10-3-10-2 3.4 at 25°C 21000 Acenaphthylene, Acy 10-3-10-2 3.93 12000 Fluorene, Flu 10-3-10-2 1.9 15000 Naphtalene, Np 0.0492 32 2300 Anthracene, An 2.10-4 0.05 – 0.07 at 28000 3 25°C Fluoranthene, Fl 10-6 to 10-4 0.26 at 25°C 340000 3 Phenanthrene 6.8.10-4 1.0 to 1.3 at 29000 3 25°C Benzo[a ] anthracene, B[a ]An 5.10-9 0.01 at 25°C 4.105 2A/B2 Benzo[ß]fluoranthene, B[ß]Fl 10-11 to 10-6 - 4.106 2B/B2 Benzo[k]fluoranthene, B[k]Fl 9.6.10-7 - 7.106 2B Chrysene, Chry 10-11 to 10-6 0.002 at 25°C 4.105 3/B2 Pyrene 6.9.10-9 0.14 at 25°C 2.105 3 Benzo[ghi]perylene, B[ghi]Pe ~10-10 0.00026 at 107 3 25°C Benzo[a ]pyrene, B[a ]Py 5.10-9 0.0038 at 25°C 106 2A/B2 Dibenzo[a ,h]anthracene, dB[a ,h]An ~10-10 0.0005 at 25°C 106 2A/B2 Indeno[1,2,3-cd]pyrene, ~10-10 - 5.107 2B/B2 I[1,2,3-cd]Py 2A/B2:Probably carcinogenic to humans/Probable human carcinogen; 2B:Possibly carcinogenic to humans; 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.

242 Appendix B

Di-(2-ethyhexyl)phthalate (DEHP)

Di-(2-ethyhexyl)phthalate (DEHP) appears at 25°C as a colourless, almost odourless, oily liquid and is fat-soluble (lipophilic). It is transported almost exclusively with fatty substances and accumulates in sediments. DEHP forms water-soluble complexes with humic and fulvic acids. Figure II.10 shows the DEHP structure formula. The more important physico-chemical qualities of DEHP are represented in Table II.30.

C4H9

HC-C2H5

CH2

O

C=O

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 Vapor Solubility KOW KOC Henry formula pressure constant [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 compounds Three representatives of the polycyclic musk compounds with abbreviation name, trade name, and molecular weight are shown in Table B.6.

Table B.6 Polyclyclic musk compounds. Abbreviation Trade name Chemical Molecular formula weight HHCB Galaxolide C18H26O 258.40 AHTN Tonalide C18H26O 258.40 ADBI Celestolide C17H24O 244.38

243 Appendix B

Nitro-musk Compounds

Properties. Nitro-musk compounds are nitro aromatic bonds with a high stability to chemical and biochemical reduction, high persistancy and lipophillic behaviour. Musk ambrette, musk xylene, musk ketone, musk tibetene and musk moskene belong to the nitro-musk compounds. Trade names and formulas of the nitro-musk compounds are listed in Table B7.

Table B7 Trade name and formula of the nitro-musk compounds. Trade name Formula Musk ambrette 1-tert.-butyl-2-methoxy-4-methyl-3,5- dinitrobenzene Musk xylene 1-tert.-butyl-3,5-dimethyl-2,4,6-trinitrobenzene Musk ketone 1-tert.-butyl-3,5-dimethyl-2,6-dinitro-4- acetylbenzene Musk tibetene 1-tert.-butyl-3,4,5-trimethyl-2,6-dinitrobenzene Musk 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 are persistent, lipophile and accumulate in the food chain. Musk xylene has a biological accumulation factor (concentration in fatty tissues / concentration in the environment) of 4.1, Musk ketone of 1.1.

244 Section Eight - References

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.uk Aqualine academically-networked ASFA-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.html Ecology Abstracts http://www.csa1.co.uk/ (CSA) EIS-Environmental Impacts Statements http://www.csa1.co.uk/ (CSA) EMBASE (Medicine and ) http://www.bids.ac.uk/ Environment Abstracts academically-networked Environmental Engineering Abstracts http://www.csa1.co.uk/ (CSA) Environmental Fate Database http://esc.syrres.com/efdb.htm ESPM-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/CEdb Index to Scientific and Technical Proceedings http://wos.mimas.ac.uk/istpcgi/login.cgi Ingenta 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.cgi Pollution 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-networked Water Resources Abstracts http://www.csa1.co.uk/ (CSA)

i Section Eight - References

8.2 REFERENCES (PAPERS)

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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 de toiture, Numéro 9693028,. (4) · Agence de l'eau Rhin-Meuse (1985) Essai d'évaluation de l'apport global de métaux lourds par les eaux usées domestiques, Pages 1-10, April 1985. · Agences de l'Eau (1997a) La mesure des micropolluants dans le cadre du réseau national de bassins, Issue 54, 1997 · Agences de l'Eau (1997b) Seuils de qualité pour les micropolluants organiques et minéraux dans les eaux superficielles, Synthese 53, 1997 · Agences de l'Eau, (1992) Etude qualitative et quantitative des sources diffuse de mercure, Pages 1-47, November 1992 · Agences de l'Eau: (1993) Etude qualitative et quantitative des sources diffuses de solvants chlorés, 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 verschmutzte Abwässer sonstiger Industriegruppen“, Dritte, überarbeitete Auflage. · ATV (1999): "ATV-Handbuch. Industrieabwasser. Grundlagen“, 4. Auflage. · ATV-A 115 (1994): 'Einleiten von nicht häuslichem Abwasser in eine öffentliche Abwasseranlage“, ATV-Regelwerk Abwasser/Abfall. · Berücksichtigung von Schadstoffen im Klärschlamm, Berichte aus Wassergütewirtschaft und Gesundheitsingenieurwesen, Technische Universität München, Heft 38, Eigenverlag München · Centre d'Information sur l'Eau (1998) Le bon usage de l'eau à la maison, Leaflet,. · Chalmers tekniska högskola. Institutionen för vattenförsörjnings- och avloppsteknik. B ; (1974) 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 nach dem Chemikaliengesetz“, (ChemVerbotsV). · Chemikalien Verbotsverordnung vom 19. Juli (1996): 'Verordnung über Verbote und Beschränkungen des Inverkehrbringens gefährlicher Stoffe, Zubereitung und Erzeugnisse nach dem Chemikaliengesetz“, (ChemVerbotsV). · Chemikaliengesetzt vom 25. Juni (1994), Bundesrepublik Deutschland: “Gesetz zum Schutz vor gefährlichen Stoffen”, (ChemG). · Chemikaliengesetzt vom 25. Juni (1994), Bundesrepublik Deutschland: “Gesetz zum Schutz vor gefährlichen Stoffen”, (ChemG). · Commission Internationale pour la Protection du Rhin (1999): Inventaire des apports de substances prioritaires dans le Rhin 1996, Pages 1-87, December 1999. (10) · Commission Internationale pour la Protection du Rhin contre la Pollution: (1987). Programme d'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 im Lockergesteinsbereich der ehemaligen DDR“, Band 22, Verlagsunion Agrar Frankfurt. · Department of the Environment, UK. (1991) The use of herbicides in non-agricultural situations in England and Wales, HMSO, London.

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· Department of Environment and Local Government, Ireland (1999), water and Sewerage Systems Annual Report. · EBAV (1996) “Scarichi da attività artigianali-impatto sugli impianti municipali di trattamento dei liquami” · 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 clash over 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: the environment agency's strategy · Environment Agency-UK, (1998), Environmental Issues Series, Endocrine-disrupting substances in 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 und Klärschlämmen, 1991-1996. · Indirekteinleiterverordnung vom April (1999), Baden-Wüttemberg: " Verordnung des Ministeriums für 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 umweltverträglichen Beseitigung 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 Beseitigung von Abfällen“, (Kreislaufwirtschafts- und Abfallgesetz-KrW-/AbfG). · Law Lecture Notes (1999) and Policy: Key Regulation Issues, Imperial College, UK. · LIFE Report 96ENV/F/410, (1999), Garantir la qualite des boues par la maitrise globale du systeme 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 the Environment in Spain 1997), Madrid. · Ministerio de Medio Ambiente (1998) Libro Blanco del Agua en España (White Paper for Water in Spain), Madrid. · Ministerio de Medio Ambiente (2000) Actuaciones Públicas en Materia de Medio Ambiente (21 de Febrero de 2000) (Public Activities regarding the Environment). · Ministry of the Environment and Energy (1999)- Denmark, Action Plan for reducing and phasing out phthalates in soft plastics June 1999 · Niedersächsisches Landesamt für Ökologie, (2000), Landwirtschaftliche Klärschlammverwertung in Niedersachsen

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· OECD (1994) Applying Economic Instruments to Environmental Policy in OECD and Dynamic Non-member Economies, OECD, Paris. · OECD (1999a) Economic Instruments for Pollution Control and Management in OECD Countries: a survey, Working Party on Economic and environmental Policy Integration. · OECD (1999b) Household water pricing in OECD countries, Report ENV/EPOC/GEEI(98)12/FINAL. · Office fédéral de l'environnement des forêts et du paysage: Informations concernant la protection des eaux: les prescriptions de la Confédération en matière de boues d'épuration, Issue 14, Pages 47, 1994. · Öko-Institute. V, (1999) Waste Prevention and Minimisation – Final Report, Institute for Applied Ecology, Berlin, Germany · Rapport / Naturvårdsverket (1999), Lead party report on combined municipal and industrial discharges (Swedish), 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 report number 773003001 · Statistika centralbyrån Sweden (1998), Discharges to water and sludge production in 1998. Mun icipal 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 and Behavior“, 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 1 und 2. · USEPA (2000), REPORT OF THE MEETING TO PEER REVIEW “THE INVENTORY OF SOURCES OF DIOXIN IN THE UNITED STATES”—Final Report—Prepared for the National Center 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 surface contaminants. · 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 zur Ordnung des Wasserhaushalts“, (Wasserhaushaltsgesetz-WHG). · Wasserhausgesetzt vom 18. November (1996), Bundesrepublik Deutschland: "Gesetz zur Ordnung des Wasserhaushalts“, (Wasserhaushaltsgesetz-WHG). · World Wildlife Federation (WWF) Canada Report (1997), Nonylphenol Ethoxylates-Reduction and Phase-Out Initiatives, Ottawa 1997. · WRc (1993). Surface water , 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 the Environment DoE 3624, June 1994 · WRc; Water Research Centre (1994) The Destruction of Organic Micropollutants in Sludge Digestion Processes. WRc Report No. UM 1441. WRc, Swindon.

xxv Section Eight - References

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

xxvi Section Eight - References

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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