Understanding the Risks to Drinking Water

Volatile Organic Compounds - Understanding the Risks to Drinking Water

Report Reference: DWI9611.04 October 2014

Volatile Organic Compounds - Understanding the Risks to Drinking Water

Report Reference: DWI9611.04

Date: October 2014

Authors: Rockett, L., Aldous, A., Benson, V., Brandon, Y., Briere, B., Dee, T., Gee, R.H., Rumsby, P., Shepherd, D., Turrell, J., and Watts, C.

Project Manager: Leon Rockett

Project No.: 16004-0

Client: Drinking Water Inspectorate

Client Manager: Peter Marsden

DWI Ref: DWI 70/2/292

RESTRICTION: This report has the following limited distribution:

External: Drinking Water Inspectorate

© Drinking Water Inspectorate 2014 The contents of this document are subject to copyright and all rights are reserved. No part of this document may be reproduced, stored in a retrieval system or transmitted, in any form or by any means electronic, mechanical, photocopying, recording or otherwise, without the prior written consent of Drinking Water Inspectorate.

This document has been produced by WRc plc.

Any enquiries relating to this report should be referred to the Project Manager at the following address:

WRc plc, Telephone: + 44 (0) 1793 865000 Frankland Road, Blagrove, Fax: + 44 (0) 1793 865001

Swindon, Wiltshire, SN5 8YF Website: www.wrcplc.co.uk

Version Control Table

Version Quality Checks Purpose Issued by Date number Approved by

DWI9611.01 Interim report issued to DWI for comment. Leon Rockett Paul Rumsby 29/05/13 Project Manager Simon Blake

DWI9611.02 Draft report of VOC modelling outputs Leon Rockett Paul Rumsby 19/11/13 Project Manager Chris Watts

DWI9611.03 Draft final report issued to DWI for Leon Rockett Paul Rumsby 07/03/14 comments Project Manager Chris Watts

DWI9611.04 Final report issued to DWI Leon Rockett Paul Rumsby 14/10/14 Project Manager

The research was funded by the Drinking Water Inspectorate, Defra under project DWI 70/2/292. The views expressed here are those of the authors and not necessarily those of the Department.

Contents

Summary ...... 1

1. Introduction ...... 6

2. 1,2-Dichloropropane ...... 7 2.1 Physico-Chemical Properties ...... 7 2.2 Toxicological Data Review ...... 7 2.3 Derivation of Tolerable Daily Intakes ...... 10 2.4 Review of Current and Historical Usage ...... 11 2.5 Review of Occurrence Data ...... 12 2.6 Review of Literature Data on Removal during Sewage Treatment ...... 12 2.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 14

3. Dichloromethane ...... 15 3.1 Physico-Chemical Properties ...... 15 3.2 Toxicological Data Review ...... 15 3.3 Derivation of Tolerable Daily Intakes ...... 18 3.4 Review of Current and Historical Usage ...... 19 3.5 Review of Occurrence Data ...... 20 3.6 Review of Literature Data on Removal during Sewage Treatment ...... 20 3.7 Review of Literature Data on Removal During Drinking Water Treatment ...... 23

4. Aniline ...... 25 4.1 Physico-Chemical Properties ...... 25 4.2 Toxicological Data Review ...... 25 4.3 Derivation of Tolerable Daily Intakes ...... 27 4.4 Review of Current and Historical Usage ...... 28 4.5 Review of Occurrence Data ...... 29 4.6 Review of Literature Data on Removal During Sewage Treatment ...... 29 4.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 33

5. Benzylchloride ...... 35 5.1 Physico-Chemical Properties ...... 35 5.2 Toxicological Data Review ...... 35 5.3 Derivation of Tolerable Daily Intakes ...... 37 5.4 Review of Current and Historical Usage ...... 38 5.5 Review of Occurrence Data ...... 39

5.6 Review of Literature Data on Removal During Sewage Treatment ...... 39 5.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 39

6. 1,3-Butadiene ...... 40 6.1 Physico-Chemical Properties ...... 40 6.2 Toxicological Data Review ...... 40 6.3 Derivation of Tolerable Daily Intakes ...... 42 6.4 Review of Current and Historical Usage ...... 44 6.5 Review of Occurrence Data ...... 47 6.6 Review of Literature Data on Removal during Sewage Treatment ...... 48 6.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 48

7. 1,1-Dichloroethane ...... 49 7.1 Physico-Chemical Properties ...... 49 7.2 Toxicological Data Review ...... 49 7.3 Derivation of Tolerable Daily Intakes ...... 51 7.4 Review of Current and Historical Usage ...... 52 7.5 Review of Occurrence Data ...... 52 7.6 Review of Literature Data on Removal during Sewage Treatment ...... 52 7.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 53

8. Nitrobenzene ...... 55 8.1 Physico-Chemical Properties ...... 55 8.2 Toxicological Data Review ...... 55 8.3 Derivation of Tolerable Daily Intakes ...... 58 8.4 Review of Current and Historical Usage ...... 59 8.5 Review of Occurrence Data ...... 60 8.6 Review of Literature Data on Removal during Sewage Treatment ...... 61 8.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 65

9. Oxirane methyl ...... 67 9.1 Physico-Chemical Properties ...... 67 9.2 Toxicological Data Review ...... 67 9.3 Derivation of Tolerable Daily Intakes ...... 69 9.4 Review of Current and Historical Usage ...... 70 9.5 Review of Occurrence Data ...... 72 9.6 Review of Literature Data on Removal during Sewage Treatment ...... 72 9.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 74

10. 1,2,3-Trichloropropane ...... 75

10.1 Physico-Chemical Properties ...... 75 10.2 Toxicological Data Review ...... 75 10.3 Derivation of Tolerable Daily Intakes ...... 77 10.4 Review of Current and Historical Usage ...... 78 10.5 Review of Occurrence Data ...... 79 10.6 Review of Literature Data on Removal during Sewage Treatment ...... 79 10.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 79

11. Urethane ...... 81 11.1 Physico-Chemical Properties ...... 81 11.2 Toxicological Data Review ...... 81 11.3 Derivation of Tolerable Daily Intakes ...... 83 11.4 Review of Current and Historical Usage ...... 84 11.5 Review of Occurrence Data ...... 85 11.6 Review of Literature Data on Removal during Sewage Treatment ...... 85 11.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 85

12. Ethylene oxide ...... 86 12.1 Physico-Chemical Properties ...... 86 12.2 Toxicological Data Review ...... 87 12.3 Derivation of Tolerable Daily Intakes ...... 89 12.4 Review of Current and Historical Usage ...... 90 12.5 Review of Occurrence Data ...... 92 12.6 Review of Literature Data on Removal during Sewage Treatment ...... 92 12.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 93

13. Formaldehyde ...... 95 13.1 Physico-Chemical Properties ...... 95 13.2 Toxicological Data Review ...... 95 13.3 Derivation of Tolerable Daily Intakes ...... 98 13.4 Review of Current and Historical Usage ...... 99 13.5 Review of Occurrence Data ...... 103 13.6 Review of Literature Data on Removal during Sewage Treatment ...... 103 13.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 107

14. o-Toluidine ...... 109 14.1 Physico-Chemical Properties ...... 109 14.2 Toxicological Data Review ...... 109 14.3 Derivation of Tolerable Daily Intakes ...... 111 14.4 Review of Current and Historical Usage ...... 112

14.5 Review of Occurrence Data ...... 114 14.6 Review of Literature Data on Removal during Sewage Treatment ...... 114 14.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 117

15. Quinoline ...... 118 15.1 Physico-Chemical Properties ...... 118 15.2 Toxicological Data Review ...... 118 15.3 Derivation of Tolerable Daily Intakes ...... 120 15.4 Review of Current and Historical Usage ...... 121 15.5 Review of Occurrence Data ...... 121 15.6 Review of Literature Data on Removal during Sewage Treatment ...... 122 15.7 Review of Literature Data on Removal during Drinking Water Treatment ...... 123

16. Modelling of the Routes of VOCs to Drinking Water ...... 125 16.1 Introduction ...... 125 16.2 Input Volumes ...... 126 16.3 Surface Water Model ...... 135 16.4 Groundwater Model ...... 148 16.5 Surface Water-Groundwater Interactions ...... 159 16.6 Limitations within the Estimations ...... 161

17. Exposure Modelling ...... 163 17.1 Dermal Exposure Model ...... 164 17.2 Inhalation Exposure Model ...... 166 17.3 Multi-route Exposure ...... 169 17.4 Comparison of Exposure Data with Tolerable Daily Intakes ...... 169

18. Conclusions ...... 173

19. Suggestions for Further Research ...... 177

20. References ...... 178 20.1 1,2-Dichloropropane ...... 178 20.2 Dichloromethane ...... 184 20.3 Aniline ...... 192 20.4 Benzylchloride ...... 203 20.5 1,3-Butadiene ...... 207 20.6 1,1-Dichloroethane ...... 213 20.7 Nitrobenzene ...... 218 20.8 Oxirane methyl ...... 227

20.9 1,2,3-Trichloropropane ...... 234 20.10 Urethane ...... 238 20.11 Ethylene oxide ...... 246 20.12 Formaldehyde ...... 252 20.13 o-Toluidine ...... 269 20.14 Quinoline ...... 279 20.15 Surface and Groundwater Model Development ...... 284 20.16 Exposure Modelling ...... 291

List of Tables

Table 2.1 Physico-chemical properties of 1,2-dichloropropane ...... 7 Table 2.2 Summary of key toxicological data for 1,2-dichloropropane ...... 8 Table 2.3 Summary of studies reporting 1,2-dichloropropane removal during sewage treatment ...... 12 Table 2.4 Summary of studies reporting toxicity of 1,2-dichloropropane to sewage treatment organisms ...... 13 Table 2.5 Removal of 1.25 mg 1,2-dichloropropane/l by nanofiltration (Agenson et al., 2003) ...... 14 Table 3.1 Physico-chemical properties of dichloromethane ...... 15 Table 3.2 Summary of key toxicological data for dichloromethane ...... 16 Table 3.3 Import and export tonnages of dichloromethane in the UK (ONS, 2013) ...... 19 Table 3.4 Summary of studies reporting dichloromethane removal during sewage treatment ...... 20 Table 3.5 Summary of studies reporting toxicity of 1,2-dichloropropane to sewage treatment organisms ...... 23 Table 4.1 Physico-chemical properties of aniline ...... 25 Table 4.2 Summary of key toxicological data for aniline ...... 26 Table 4.3 Summary production and usage data reported by the European Chemical Bureau (EU, 2004) ...... 28 Table 4.4 Summary of studies reporting aniline removal during sewage treatment ...... 29 Table 4.5 Summary of studies reporting toxicity of aniline to sewage treatment organisms ...... 33 Table 5.1 Physico-chemical properties of benzylchloride ...... 35 Table 5.2 Summary of key toxicological data for benzylchloride ...... 36 Table 5.3 Summary of studies reporting benzylchloride removal during sewage treatment ...... 39 Table 6.1 Physico-chemical properties of 1,3-butadiene ...... 40 Table 6.2 Summary of key toxicological data for 1,3-butadiene ...... 41 Table 6.3 Summary of Western European production data (EU, 2002) ...... 44 Table 6.4 Estimated percentage of total 1,3-Butadiene used in Europe per application type (EU, 2002) ...... 45 Table 6.5 Production capacity of synthetic rubber products by country in 1994 (EU, 2002) ...... 45 Table 6.6 Estimated consumption of synthetic rubber products in Western Europe (EU, 2002) ...... 46 Table 6.7 Estimated breakdown of usage of 1,3-butadiene in the US (ATSDR, 2012) ...... 46 Table 6.8 Estimated releases of 1,3-butadiene into the Canadian environment in 1994 (WHO, 2001) ...... 47 Table 7.1 Physico-chemical properties of 1,1-dichloroethane ...... 49

Table 7.2 Summary of key toxicological data for 1,1-dichloroethane...... 50 Table 7.3 Summary of studies reporting 1,1-dichloroethane removal during sewage treatment ...... 53 Table 7.4 Summary of studies reporting toxicity of 1,1-dichloroethane to sewage treatment organisms ...... 53 Table 8.1 Physico-chemical properties of nitrobenzene ...... 55 Table 8.2 Summary of key toxicological data for nitrobenzene ...... 56 Table 8.3 Estimated usage per application type in the EU (EU, 2007) ...... 59 Table 8.4 Production capacities for European countries in 1985 (WHO, 2003) ...... 60 Table 8.5 Summary of studies reporting nitrobenzene removal during sewage treatment ...... 61 Table 8.6 Summary of studies reporting toxicity of nitrobenzene to sewage treatment organisms ...... 64 Table 9.1 Physico-chemical properties of oxirane methyl ...... 67 Table 9.2 Summary of key toxicological data for oxirane methyl ...... 68 Table 9.3 Usage of oxirane methyl by product type (EU, 2002) ...... 71 Table 9.4 Summary of historical global oxirane methyl production data (OECD, 2001 and EU, 2002) ...... 72 Table 9.5 Import and export tonnages of oxirane methyl in the UK (ONS, 2013) ...... 72 Table 9.6 Summary of studies reporting oxirane methyl removal during sewage treatment ...... 73 Table 10.1 Physico-chemical properties of 1,2,3-trichloropropane ...... 75 Table 10.2 Summary of key toxicological data for 1,2,3-trichloropropane ...... 76 Table 10.3 Summary of studies reporting toxicity of 1,2,3- trichloropropane to sewage treatment organisms ...... 79 Table 11.1 Physico-chemical properties of urethane ...... 81 Table 11.2 Summary of key toxicological data for urethane ...... 82 Table 12.1 Physico-chemical properties of ethylene oxide ...... 86 Table 12.2 Summary of key toxicological data for ethylene oxide ...... 87 Table 12.3 Annual data for imports and exports into the UK ...... 91 Table 12.4 Summary of studies reporting ethylene oxide removal during sewage treatment ...... 92 Table 12.5 Summary of studies reporting toxicity of ethylene oxide to sewage treatment organisms ...... 93 Table 13.1 Physico-chemical properties of formaldehyde ...... 95 Table 13.2 Summary of key toxicological data for formaldehyde ...... 96 Table 13.3 Global production of formaldehyde ...... 99 Table 13.4 European production data from 1982 to 1990 (IPCS, 2002) ...... 99 Table 13.5 Annual production volumes for the USA (ATSDR, 1999 and 2010) ...... 100

Table 13.6 Approximate percentages of global formaldehyde use (OECD, 2002; ATSDR, 2010) ...... 101 Table 13.7 Swiss products register approximate concentrations in commercial and domestic products (OECD, 2002) ...... 102 Table 13.8 Summary of studies reporting formaldehyde removal during sewage treatment ...... 103 Table 13.9 Summary of studies reporting toxicity of formaldehyde to sewage treatment organisms ...... 106 Table 13.10 Removal of formaldehyde by sand/coal-based GAC, sand/wood-based GAC and sand/anthracite coal media ...... 107 Table 13.11 Removal of formaldehyde by ozone, UV and ozone/UV ...... 108 Table 14.1 Physico-chemical properties of o-toluidine ...... 109 Table 14.2 Summary of key toxicological data for o-toluidine ...... 110 Table 14.3 Estimated production volumes in 2001 (OECD, 2004) ...... 113 Table 14.4 Summary of studies reporting o-toluidine removal during sewage treatment ...... 114 Table 14.5 Summary of studies reporting toxicity of o-toluidine to sewage treatment organisms ...... 116 Table 15.1 Physico-chemical properties of quinoline ...... 118 Table 15.2 Summary of key toxicological data for quinoline ...... 119 Table 15.3 Summary of studies reporting quinoline removal during sewage treatment ...... 122 Table 15.4 Summary of studies reporting toxicity of quinoline to sewage treatment organisms ...... 123 Table 16.1 Input volumes used in the surface and groundwater models ...... 135 Table 16.2 Removal of VOCs during sewage treatment ...... 136 Table 16.3 Fugacity distribution of VOCs ...... 138 Table 16.4 Indicative magnitude of removal ...... 139 Table 16.5 Summary of results of the surface water model ...... 143 Table 16.6 Comparison of the surface water model output with literature data ...... 144 Table 16.7 Typical physical soil parameters ...... 150 Table 16.8 Remedial targets methodology parameters ...... 151 Table 16.9 Diffuse source methodology results ...... 153 Table 16.10 Point source methodology results ...... 154 Table 16.11 Estimated concentrations in drinking water using the groundwater diffuse pollution model ...... 157 Table 16.12 Estimated concentrations in drinking water using the groundwater point source pollution model (at 50 metres) ...... 158 Table 16.13 Estimated concentrations in drinking water using the groundwater point source pollution model (at 100 metres) ...... 159 Table 16.14 Estimated concentrations in drinking water from a combined surface water-groundwater (diffuse source) model ...... 160

Table 17.1 Determination of skin permeability co-efficients for the fourteen VOCs...... 165 Table 17.2 Dermal exposure to VOCs as a daily litre-equivalent ...... 165 Table 17.3 Determination of the significance of volatilisation of VOCs during bathing and showering ...... 168 Table 17.4 Inhalation exposure to VOCs as a daily litre-equivalent...... 168 Table 17.5 Equivalent total daily intake of the fourteen VOCs following oral consumption of drinking water and one 30-minute bath or shower ...... 169 Table 17.6 Risk characterisation ratios for the fourteen VOCs using the combined surface water-ground water (diffuse source) model and assuming the poorest removal of the VOC during drinking water treatment ...... 171 Table 18.1 Tolerable daily intakes for the fourteen VOCs ...... 173 Table 18.2 Equivalent total daily intake of the fourteen VOCs following oral consumption of drinking water and one 30-minute bath or shower ...... 174 Table 18.3 Risk characterisation ratios for the fourteen VOCs using the combined surface water-ground water (diffuse source) model ...... 175

List of Figures

Figure 16.1 Diagrammatic representation of models predicting the occurrence of VOCs in drinking water ...... 126 Figure 16.2 Groundwater risk assessment results ...... 155

Drinking Water Inspectorate

Summary i Reasons

Volatile Organic Compounds (VOCs) are compounds that contain, at least, the element carbon and one or more element from hydrogen, oxygen, sulphur, phosphorus, silicon, nitrogen or halogens, and have a high vapour pressure at room temperature. VOCs have a wide range of uses in many industries as solvents, chemical intermediates and components of paints and varnishes. Additionally, many have a diverse range of historical uses that may have resulted in significant releases to the environment. Fourteen VOCs on the US Environmental Protection Agency‘s (EPA‘s) 3rd Chemical Contamination List (CCL3) have been identified for further investigation; 1,2-dichloropropane, dichloromethane, aniline, benzylchloride, 1,3-butadiene, 1,1-dichloroethane, nitrobenzene, oxirane methyl, 1,2,3-trichloropropane, urethane, ethylene oxide, formaldehyde, o-toluidine and quinonline. ii Objectives

1. Review existing literature on VOCs, and summarise data on occurrence in the environment including water sources and treated water, toxicity, physico-chemical properties and removal in water treatment;

2. Identify the quantities used and manufactured in England and Wales and the industries that use them;

3. Determine the possible routes into the water system;

4. Estimate the likely concentrations found in raw and treated water;

5. Compare estimate of potential exposure via drinking water with appropriate toxicological end points; and

6. Compare potential drinking water intakes of VOCs with other routes of exposure. iii Benefits

This project provides a better understanding of the potential risks of contamination of drinking water supplies by these VOCs, and allows determination of whether exposure to these substances via drinking water poses a risk to human health. It also provides guidance to water companies on the factors that may need to be considered in their risk assessments of VOCs.

iv Conclusions

Detailed toxicological reviews considering oral, dermal and inhalation exposure have been conducted on fourteen VOCs within this report. These reviews identified a number of gaps within the toxicological databases for these chemicals, particularly in regards to dermal toxicity data. However, Tolerable Daily Intakes (TDIs) via the oral and inhalation routes have been derived for all fourteen VOCs, and dermal TDIs have been derived where appropriate.

VOC Oral TDI Dermal TDI Inhalation TDI

1,2-Dichloropropane 14 µg/kg bw/day 14 µg/kg bw/day 66 µg/m³

Dichloromethane 6 µg/kg bw/day 30 µg/kg bw/day 124 µg/m³

Aniline 7 µg/kg bw/day No TDI derived 6 µg/m³

Benzylchloride 6 µg/kg bw/day No TDI derived 27 µg/m³

1,3-Butadiene 5 µg/kg bw/day No TDI derived 2 µg/m³

1,1-Dichloroethane 475 µg/kg bw/day 475 µg/kg bw/day 3620 µg/m³

Nitrobenzene 5 µg/kg bw/day 50 µg/kg bw/day 0.7 µg/m³

Oxirane methyl 17 µg/kg bw/day No TDI derived 71 µg/m³

1,2,3-Trichloropropane 2 µg/kg bw/day 2 µg/kg bw/day 1.7 µg/m³

Urethane 0.07 µg/kg bw/day 0.07 µg/kg bw/day 250 µg/m³

Ethylene oxide 15 µg/kg bw/day No TDI derived 60 µg/m³

Tolerable Concentration: Formaldehyde 1380 µg/kg bw/day 0.06 µg/m³ 2600 µg/l

o-Toluidine 150 µg/kg bw/day 8 µg/kg bw/day 67 µg/m³

Quinoline 25 µg/kg bw/day 25 µg/kg bw/day 19 µg/m³

VOC Total Intake (l-eq/day)

1,2-Dichloropropane 4.47

Dichloromethane 4.39

Aniline 2.37

Benzyl chloride 3.86

1,3-Butadiene 8.72

1,1-Dichloroethane 4.85

Nitrobenzene 2.70

Oxirane methyl 2.25

1,2,3-Trichloropropane 3.28

Urethane 2.00

Ethylene oxide 2.54

Formaldehyde 2.78

o-Toluidine 2.54

Quinoline 2.63

Three types of models have been developed to estimate the concentrations in drinking water; a surface water model; a diffuse source groundwater model; and a point source groundwater model. The results of the surface water model and the diffuse source groundwater model were combined to provide an extreme worst-case scenario of a drinking water treatment works that is supplied by a river that also receives a significant proportion of flow from a groundwater source.

These models produced estimated concentrations in final drinking water that were used to calculate estimated daily exposures which were compared to the oral TDIs to determine the Risk Characterisation Ratios (RCRs) for each of the VOCs.

 An RCR of less than 1 would indicate that there is unlikely to be a concern following exposure to that VOC in drinking water at the concentrations predicted by the model.

 An RCR of greater than 1 indicates that it is not possible to preclude adverse health effects following exposure to a VOC in drinking water at the concentrations predicted in the model.

All of the VOCs considered in this study had RCRs significantly below 1. Therefore, none are anticipated to be of concern to human health.

VOC Risk Characterisation Ratio

1,2-Dichloropropane 0.0000759

Dichloromethane 0.00175

Aniline 0.0015

Benzyl chloride 0.0000264

1,3-Butadiene 0.00416

1,1-Dichloroethane 0.00000566

Nitrobenzene 0.00495

Oxirane methyl 0.000616

1,2,3-Trichloropropane 0.000118

Urethane 0.00138

Ethylene oxide 0.000384

Formaldehyde 0.000225

o-Toluidine 0.00000304

Quinoline 0.0000582

However, it should be noted that these TDIs are based on the assumption of a threshold level for an adverse health effect; the weight of evidence indicates that 1,3-butadiene and ethylene oxide are both genotoxic in vivo, and are both considered to be Group 1 carcinogens (i.e. carcinogenic to humans) by the International Agency for Research on Cancer (IARC). Therefore, there is theoretically no ‗safe‘ level for these chemicals, and it would be appropriate to ensure that their concentrations in drinking water are as low as reasonably practicable. v Suggestions

This project was commissioned to provide an understanding of the significance of exposure to VOCs in drinking water. The available literature data and the results of exposure modelling suggest that the concentrations of VOCs in drinking water are very low, and with regards to non-carcinogenic endpoints, are likely to be at concentrations below those of health concern.

However, there are several gaps within the data that, if addressed, may aid in the validation of the model and in supporting this conclusion. Therefore, the following suggestions and considerations are made based on this report:

 The most significant data gap within this report is a lack of information on the current volumes of VOCs manufactured and used within England and Wales. Additional information would significantly aid in the development of the model, and may enable the development of models to assess the risks to drinking water supplies in specific regions.

 The available data on the occurrence of these VOCs are also limited. As such, while the results from the model have generally been consistent with reported environmental and drinking water concentrations, it has not been possible to fully validate the results of the model. Further information on the occurrence of these VOCs in the environment would aid in the determination of the reliability of the model predictions.

It has not been possible to derive tolerable daily intakes via all three of the routes considered within this project (oral, dermal and inhalation) for a number of chemicals due to a lack of relevant toxicological data. To overcome this, route-to-route extrapolations have been applied to derive as many tolerable daily intakes as possible, and a multi-route exposure approach developed. This is considered to be a conservative approach; however, if additional toxicological data were to become available, it would be appropriate to reconsider this approach in the light of these additional data.

1. Introduction

DWI identified ten VOCs that are being considered for group regulation in the US for investigation; 1,2-dichloropropane, dichloromethane, aniline, benzylchloride, 1,3-butadiene, 1,1-dichloroethane, nitrobenzene, oxirane methyl, 1,2,3-trichloropropane and urethane. During the early stages of the project a further four compounds on the US Environmental Protection Agency‘s (EPA‘s) 3rd Chemical Contamination List (CCL3) were identified: ethylene oxide, formaldehyde, o-toluidine and quinoline.

The aims of this project are to provide a better understanding of the potential risks of contamination of drinking water supplies by these fourteen VOCs, to determine if exposure to these substances via drinking water poses a risk to human health and to provide guidance to water companies on the factors that may need to be considered in their risk assessments of VOCs.

Therefore, the objectives of this project are to:

2. Identify the quantities used and manufactured in England and Wales and the industries that use them;

3. Determine the possible routes into the water system;

4. Estimate the likely concentrations found in raw and treated water;

5. Compare estimate of potential exposure via drinking water with appropriate toxicological end points; and

6. Compare potential drinking water intakes of VOCs with other routes of exposure.

2. 1,2-Dichloropropane

2.1 Physico-Chemical Properties

Data on the physico-chemical properties of 1,2-dichloropropane are provided in Table 2.1.

Table 2.1 Physico-chemical properties of 1,2-dichloropropane

CAS Number 78-87-5

Chemical Formula C3H6Cl2

Structure

Molecular weight 112.986 (ChemID, 2013)

Physical state Liquid (OECD, 2003)

Melting point 96.8°C (WHO, 2003)

Boiling point -100°C (WHO, 2003)

Water solubility 2700 mg/l at 20°C (WHO, 2003)

Log Kow 1.99 (WHO, 2003)

Koc 68 (OECD, 2003)

Vapour pressure 55.3 mmHg at 25°C (SRC, 2013a)

Henry’s Law constant 0.00282 atm.m³/mole at 25°C (SRC, 2013a)

Dissociation constant No data

Density 1.155 g/cm² at 20°C (OECD, 2003)

2.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with 1,2-dichloropropane is provided in Appendix B1, and only an overall summary is provided here (Table 2.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 2.2 Summary of key toxicological data for 1,2-dichloropropane

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 860 mg/kg bw (mice) to 2200 mg/kg bw (rats)

Dermal Low acute toxicity LD50: 10 100 mg/kg bw (rabbits)

Inhalation Low acute toxicity LC50 range: 2256 mg/m3 (mice) to 14 000 mg/m3 (rats)

Irritation and - Irritating effects have been observed in both the eyes and the skin of Sensitisation experimental animals. Data on sensitisation are equivocal. No data are available on the respiratory irritation.

Genotoxicity In vitro and The in vitro genotoxicity studies have given mixed results, although there In vivo appears to be genotoxic activity in mammalian cell assays. The majority of in vivo studies located have been negative, indicating that 1,2-dichloropropane is unlikely to be genotoxic in vivo.

Repeat Dose Oral Sub-chronic Study: Key study used for assessment: Toxicity and Sprague-Dawley male rats (15-16/dose) were administered Carcinogenicity 1,2-dichloropropane in corn oil at doses of 0, 100, 250, 500 or 750 mg/kg bw/day, once daily, 5 days/week for 13 weeks.

A LOAEL of 100 mg/kg bw/day was identified (71.4 mg/kg bw/day following correction for the 5 days/week dosing regimen) based on haematological changes.

Dermal No data were located

Inhalation Chronic Study: Key study used for assessment: Fischer 344/DuCrj rats (50/sex/dose) were administered 1,2-dichloropropane via inhalation at doses of 0, 80, 200 or 500 ppm (0, 370.61, 926.52 and 2316.30 mg/m³, respectively) for 6 hours/day, 5 days/week for 2 years.

A LOAEC of 80 ppm (370.61 mg/m3) was identified based on induction of esthesioneuroepitheliomas in the naval cavity.

Reproductive and Oral Reproductive Study: Developmental Sprague-Dawley rats (30/sex/dose) were administered 1,2-dichloropropane Toxicity in drinking water at doses of 20-30, 70-130, 130-250 mg/kg bw/day. Female rats were administered higher doses of 60, 200 and 450-500 mg/kg bw/day during lactation.

Parental NOAEL (decreased bodyweight): 20-30 mg/kg bw/day Offspring NOAEL (decreased post-natal survival and neonatal bodyweight):

Route or Endpoint Summary study type 70-130 mg/kg bw/day Reproductive NOAEL: 130-250 mg/kg bw/day (highest dose tested)

Developmental Study: Pregnant Sprague-Dawley rats (30/dose) and pregnant New Zealand white rabbits (18/dose) were administered 1,2-dichloropropane by gavage in corn oil at concentrations of 0, 10, 30 or 125 mg/kg bw/day in rats (days 6-15 of gestation) and 0, 15, 50 or 150 mg/kg bw/day in rabbits (days 7-19 of gestation).

Rat maternal NOAEL (decreased body weight): 30 mg/kg bw/day Rat foetal NOAEL (delayed ossification of skull bones): 30 mg/kg bw/day Rat teratogenic NOAEL: 125 mg/kg bw/day (highest dose tested)

Rabbit maternal NOAEL (decreased body weight and haematological effects): 50 mg/kg bw/day Rabbit foetal NOAEL (delayed ossification of skull bones): 50 mg/kg bw/day Rabbit teratogenic NOAEL: 150 mg/kg bw/day (highest dose tested)

Dermal No data were located

Inhalation No data were located

2.2.1 Evaluations by Authoritative Bodies

In 1991 The United States Environmental Protection Agency (US EPA) concluded that there was insufficient data to derive an Oral Reference Dose (RfD) for 1,2-dichloropropane, however an Inhalation Reference Concentration (RfC) of 0.004 mg/m3 was derived based on a 13-week inhalation study (Nitschke et al., 1988, cited in OECD, 2003).

The International Agency for Research on Cancer (IARC) evaluated 1,2-dichloropropane in 1986 and 1999. IARC classified 1,2-dichloropropane as Group 3 (i.e. not classifiable as to its carcinogenicity to humans), based on a lack of relevant epidemiological data and limited evidence for carcinogenicity in experimental animals (IARC, 1986).

In 1998 the World Health Organisation (WHO) concluded that based on the IARC evaluation above, the use of a threshold approach for the toxicological evaluation of 1,2-DCP was appropriate. WHO considered the 13-week toxicity study in which male rats were administered 1,2-DCP by gavage in corn oil for 5 days/week (Bruckner et al., 1989) to be the most appropriate study for the derivation of a drinking water guideline value. The LOAEL was reported to be 71.4 mg/kg bw/day (following correction for the 5 days/week dosing regimen) based on changes in haematological parameters. WHO applied an uncertainty factor of 5000 (100 for inter- and intra-species variation, 10 for the use of a LOAEL and 5 for limitations of

the database, including the limited data on in vivo genotoxicity and use of a subchronic study) to derive a Tolerable Daily Intake (TDI) of 14 µg/kg bw/day. 10% of this TDI was allocated to exposure via drinking-water and a 60 kg adult drinking 2 litres of water a day was used as the basis to calculate the provisional guideline value of 40 µg/l (rounded) WHO states that the guideline value is considered to be provisional owing to the magnitude of the uncertainty factor and the fact that the database has not changed since the previous guideline value was derived. This evaluation was still current in the 4th edition of the Guidelines for Drinking-water Quality (GDWQ) published in 2011 (WHO, 2011).

2.2.2 Toxicity Summary

1,2-Dichloropropane is of low acute oral, dermal and inhalation toxicity in experimental animals and the available data on the irritation and sensitisation of 1,2-dichloropropane to experimental animals indicate the presence of irritating effects to both eyes and skin. Although in vitro genotoxicity studies have given mixed results, the majority of in vivo studies located have been negative, indicating that 1,2-dichloropropane is unlikely to be genotoxic in vivo. Studies in experimental animals in which 1,2-dichloropropane was administered orally indicate toxicity to the liver and a low incidence of hepatic adenocarcinomas, in addition to haemotoxicity. An increased incidence in mammary gland adenocarcinomas was also observed. 1,2-Dichloropropane has been classified as Group 3 by IARC (i.e. not classifiable as to its carcinogenicity to humans) In reproductive and developmental toxicity studies limited evidence of malformations during foetus development was observed, however the concentrations at which reproductive toxicity occurred were greater than those for maternal toxicity.

2.3 Derivation of Tolerable Daily Intakes

2.3.1 Oral

In 1998 (and current for the 4th edition of the guidelines published in 2011), in the derivation of their provisional Guideline for Drinking-water Quality (GDWQ), the World Health Organization (WHO) concluded that the use of a threshold approach for the toxicological evaluation of 1,2-dichloropropane was appropriate. WHO considered the 13-week toxicity study in which male rats were administered 1,2-dichloropropane by gavage in corn oil for 5 days/week to be the most appropriate study for the derivation of a guideline value. A LOAEL of 71.4 mg/kg bw/day (following correction for the 5 days/week dosing regimen) was identified based on changes in haematological parameters. An uncertainty factor of 5000 (100 for inter- and intra-species variation, 10 for the use of a LOAEL and 5 for limitations of the database, including the limited data on in vivo genotoxicity and use of a subchronic study) to derive an oral Tolerable Daily Intake (TDI) of 14 µg/kg bw/day (rounded) (WHO, 2003; WHO, 2011).

2.3.2 Dermal

No dermal toxicity studies were located upon which a dermal TDI can be derived. However, as there is no evidence to suggest that dermal exposure will significantly impair the integrity of

skin (i.e. 1,2-dichloropropane is not corrosive), a dermal TDI can be derived by extrapolation from oral toxicity data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

1,2-Dichloropropane is reported to be well absorbed via both the oral and the dermal route, however, no quantitative data on absorption via the dermal route were located. Therefore, equivalent bioavailability for the oral and dermal routes can be assumed. Therefore a dermal TDI of 14 µg/kg bw/day can be derived.

However, it should be noted that this is likely to be a highly conservative value, as in reality, very few substances are absorbed as readily via the dermal route as they are via the oral route.

2.3.3 Inhalation

In a 2-year inhalation study, rats were exposed to 1,2-dichloropropane for 6 hours/day, 5 days/week at concentrations of 0, 80, 200 or 500 ppm. A LOAEC of 80 ppm (approximately 370 mg/m³) was identified based on neoplasms. Adjusting to a continuous exposure, this LOAEC would be equivalent to a concentration of 66 mg/m³. Following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEC), an inhalation Tolerable Daily Concentration of 0.066 mg/m³ (66 µg/m³) is identified.

2.4 Review of Current and Historical Usage

In 2003 the World Health Organization listed the primary use of 1,2-dichloropropane as an intermediate in the production of perchloroethylene (WHO, 2003). It was also used as a solvent in fats, oils, resins and waxes (WHO, 2003). 1,2-Dichloropropane historically has been used as a soil and peach tree insecticide fumigant and as a dry cleaning fluid, paint remover, metal degreasing agent and lead scavenger for antiknock fluids (IARC, 1986).

The Organization for Economic Co-operation and Development quoted an annual global production of 1,2-dichloropropane in 2001 of 350 kilotonnes (OECD, 2003). In Switzerland, 46 industrial products were listed in 2003 (including paints, lacquers, varnishes, solvent, degreasers, diluters and strippers) as containing 1,2-dichloropropane and no product applications in Denmark and France. Use in agriculture in North America and Europe has been banned. The OECD highlighted that 1,2-dichloropropane is used primarily in site-limited or limited-transported as a co-product/raw material for the manufacture of many chlorinated products (OECD, 2003).

Production of 1,2-dichloropropane declined in the United States of America (USA) between 1979 and 1999 as its uses as a soil fumigant, industrial solvent, paint stripper and varnish ingredient declined. The main use in the USA is as an intermediate in the production of perchloroethylene (ATSDR, 1999). Use as a soil fumigant was halted in the USA prior to

1989; (ATSDR, 1989, is the earliest reference located that notes 1,2-dichloropropane is no longer used as a pesticide in the USA).

In the UK, 1,2-dichloropropane was used in agriculture as a soil fumigant prior to 1980 (PPDB, 2013) it is no longer approved for use. It is no longer approved for use in pesticides by the European Union under Regulation EC 1107/2009 (EU, 2013).

2.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with 1,2-dichloropropane is provided in Appendix C1, and only an overall summary is provided here.

Data for 1,2-dichloropropane are available for drinking water, groundwater, fresh and marine surface water and effluents. Concentrations of 1,2-dichloropropane in drinking water range from 0.01 to 21 µg/l, in ground water from 0.01 to 1200 µg/l, in fresh surface water from 0.007 to 300 µg/l, in marine surface water <0.1 µg/l and in sewage effluents from 0 to 210 µg/l.

2.6 Review of Literature Data on Removal during Sewage Treatment

There are studies available reporting removal of 1,2-dichlopropropane during sewage treatment indicating that it is biodegradable after a period of adaptation, and is of low toxicity to sewage treatment organisms. The available data are summarised in Table 2.3 and Table 2.4.

Table 2.3 Summary of studies reporting 1,2-dichloropropane removal during sewage treatment

Concentration Method/type of study Inoculum Degradation Reference (mg/l) Aerobic

Kawasaki (1980) activated 14 day Japanese MITI test 100 29% theoretical BOD cited in SRC sludge (2013b)

42%-89% depending Tabak et al. 7 day serial shake flask sewage seed 5 on acclimation (1981) cited in period SRC (2013b) 36%-81% depending Tabak et al. 7 day serial shake flask sewage seed 10 on acclimation (1981) cited in period SRC (2013b) 28 day OECD 301D activated 0% (not inherently 1000 OECD (2003) (closed bottle test) sludge biodegradable) 28 day OECD 302B activated 0% (not inherently 150 OECD (2003) (modified Zahn-Wellens test) sludge biodegradable)

Concentration Method/type of study Inoculum Degradation Reference (mg/l)

activated Biodegradation test NR 0% after 7 days OECD (2003) sludge Flask screening method of domestic waste Bunch and Chambers, non- 10 42% after 7 days OECD (2003) water adapted Flask screening method of domestic waste 10 81% after 7 days OECD (2003) Bunch and Chambers, adapted water Flask screening method of domestic waste 5 89% after 7 days OECD (2003) Bunch and Chambers, adapted water

Influent and Biodegradation test effluent of 182 98.9 – 99.2% OECD (2003) chemical waste OECD Guide-line 302 B Activated "Inherent biodegradability: NR 96% after 3 hours OECD (2003) sludge Modified Zahn-Wellens Test" Biodegradation test sludge 100 0% after 14 days OECD (2003) Semi-continuous activated Shell (1984) activated sludge process over ten weeks NR No degradation cited in ATSDR sludge and retention time of 25 hours (1989)

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound

MITI: Ministry of International Trade and Industry, Japan

NR: Not Reported

Table 2.4 Summary of studies reporting toxicity of 1,2-dichloropropane to sewage treatment organisms

Method/type of Concentration. Inoculum Results Reference study (mg/l)

Biodegradation test Nitrosolobus 75% degradation after OECD 0.048 (not specified) multiformis 1 hour (2003) Biodegradation test Pseudomonas 3% degradation after OECD 100 (not specified) fluorescens 24 hours (2003) IUCLID NR activated sludge 630 EC50 (effect not stated) (2000)

EC50: Median Effect Concentration i.e. the concentration that is estimated to cause a specified toxic effect in 50% of exposed organisms

NR: Not Reported

2.7 Review of Literature Data on Removal during Drinking Water Treatment

Speth and Miltner (1990) have reported carbon binding capacities for 1,2-dichloropropane of 0.7 mg/g and 7.3 mg/g for concentrations of 4.1 µg/l and 196 µg/l, respectively, at equilibrium. In laboratory tests at equilibrium concentrations of 0.5 mg/l, activated carbons gave adsorption capacities between 3 and 10 mg/g (Bembnowska et al., 2003). Najm et al. (1991) has estimated that a Powdered Activated Carbon (PAC) dose of 30 mg/l is required to reduce a concentration of 10 µg 1,2-dichloropropane/l to 1 µg/l.

Ozonation is not effective at reducing the concentration of 1,2-dichloropropane during drinking water treatment. Ozonation of an unspecified (but in the range 50 to 384 µg/l) concentration 1,2-dichloropropane gave removals of 0%, 0% and 5% for applied ozone doses of 2, 6 and 20 mg/l, respectively (Fronk, 1987).

Gauntlett and Packham (1974) have conducted studies using reverse osmosis, in which they reported removals of 10%, 61% and 90% using cellulose acetate, polyamide and thin film composite membranes, respectively.

Agenson et al. (2003) conducted a study on the effectiveness of a range of nanofiltration membranes at removing a 1.25 mg/l solution of 1,2-dichloropropane during drinking water treatment. The results of this study are presented in Table 2.5, and indicate a wide range of removal efficiencies.

Table 2.5 Removal of 1.25 mg 1,2-dichloropropane/l by nanofiltration (Agenson et al., 2003)

Membrane Type Removal (%)

UTC60 Aromatic polyamide 10

NTR 729HF Polyvinyl alcohol/polyamide 42

ES10C Polyamide 68

UTC70 Aromatic polyamide 68

LF10 Polyvinyl alcohol/ polyamide 64

According to the US Environmental Protection Agency (EPA), 1,2-dichloropropane has ―average strippability‖ in air. Studies examining the effectiveness of air stripping have reported removal efficiencies of 50-90% (US EPA, 2009).

3. Dichloromethane

3.1 Physico-Chemical Properties

Data on the physico-chemical properties of dichloromethane are provided in Table 3.1.

Table 3.1 Physico-chemical properties of dichloromethane

CAS Number 75-09-2

Chemical Formula CH2Cl2

Structure

Molecular weight 84.9328 (ChemID, 2013)

Physical state Clear, colourless liquid (IPCS, 1996)

Melting point -95.1°C (WHO, 2003)

Boiling point 40°C (WHO, 2003)

Water solubility 20 000 mg/l at 20°C (WHO, 2003)

Log Kow 1.3 (WHO, 2003)

Koc Log Koc: 0.89 (IPCS, 1996)

Vapour pressure 435 mm Hg at 25°C (SRC, 2013a)

Henry’s Law constant 0.00325 atm.m³/mole at 25°C (SRC, 2013a)

Dissociation constant No data

Density 1.3255 g/cm³ at 20°C (WHO, 2003)

3.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with dichloromethane is provided in Appendix B2, and only an overall summary is provided here (Table 3.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 3.2 Summary of key toxicological data for dichloromethane

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 1410-3000 mg/kg bw (rats)

Dermal Low acute toxicity LD50: >2000 mg/kg bw (rats)

Inhalation Low acute toxicity LC50 range: 49 118 mg/m3 (rats) to 51 500 mg/m3 (mice)

Irritation and - Irritating effects have been observed in both the eyes and the skin of Sensitisation experimental animals. Not a skin sensitiser. Irritating to the respiratory tract in humans.

Genotoxicity In vitro and Tests for the genotoxicity of dichloromethane in bacterial assays are generally in vivo positive, although the majority of mammalian in vitro and in vivo tests indicate that dichloromethane is not genotoxic. The exceptions to these are the in vitro and in vivo studies using B6C3F1 mice. The positive results observed in this strain of mice may be a result of the high rate of metabolism by glutathione-S- transferase enzymes. Overall, the effect of dichloromethane on mammalian cells has only indicated weak genotoxicity.

Repeat Dose Oral Chronic Study: Key study used for assessment: Toxicity and Fischer 344 rats (85/sex/dose) were administered dichloromethane in drinking Carcinogenicity water at doses of approximately 0, 5, 50, 125 or 250 mg/kg bw/day (reported as 0, 6, 52, 125 and 235 mg/kg bw/day in males, respectively, and 0, 6, 58, 136 and 263 mg/kg bw/day in females, respectively) for 104 weeks

A NOAEL of 6 mg/kg bw/day was identified for this study based on liver lesions.

Dermal No data were located

Inhalation Chronic Study: Key study used for assessment: Sprague-Dawley rats (90 males/dose, 108 females/dose) were exposed to dichloromethane via inhalation at doses of 0, 50, 200 or 500 ppm (0, 174.11, 696.43 and 1741.07 mg/m³, respectively) for 6 hours/day, 5 days/week for 20 months for males and 24 months for females

A NOAEC of 200 ppm (696.43 mg/m³) was identified based on histopathological changes in the liver.

Reproductive and Oral Developmental Study (study not considered reliable): Developmental Fischer 344 rats were administered dichloromethane in corn oil by gavage at Toxicity concentrations of 0 mg/kg bw/day (21 treated, 15 pregnant), 337.5 mg/kg bw/day (16 treated, 13 pregnant) or 450 mg/kg bw/day (17 treated, 14 pregnant) on gestational days 6 to 19.

Route or Endpoint Summary study type

Maternal NOAEL (decreased weight gain and increased extrauterine weight gain, rales, nasal congestion and vocalisation): 337.5 mg/kg bw/day Developmental NOAEL: 450 mg/kg bw/day (highest dose tested)

Dermal No data were located

Inhalation Reproductive Study: Fischer 344 rats (30/sex/dose/generation) were administered dichloromethane via inhalation at doses of 0, 100, 500 or 1500 ppm (0, 348.21, 1741.07 and 5223.30 mg/m³) for 6 hours/day, 5 days/week, beginning 13-14 weeks before mating and continuing until the end of the study.

Reproductive NOAEL: 1500 ppm (5223.3 mg/m³; the highest dose tested)

3.2.1 Evaluation by Authoritative Bodies

In 2011, the US Environmental Protection Agency (EPA) derived an oral Reference Dose (RfD) for dichloromethane of 0.006 mg/kg bw/day (6 µg/kg bw/day; rounded). This oral RfD was based on a NOAEL of 6 mg/kg bw/day identified in a 2-year study in rats. Using a physiologically-based pharmacokinetic (PBPK) model, a 1st percentile human equivalent dose (HEQ) of 0.19 mg/kg bw/day was identified. An uncertainty factor of 30 (rounded) was applied to the HEQ (3 to account for inter-species differences, 3 to account for intra-species sensitivity and 3 to account for the limited database) was applied to the HEQ to derive the oral RfD (US EPA, 2012). In 2011, the US EPA also derived an Inhalation Reference Concentration (RfC) of 0.6 mg/m3 based on a 2-year inhalation study (Nitschke et al., 1988a).

In 1999, the International Agency for Research on Cancer (IARC) evaluated dichloromethane and classified it as Group 2B, (i.e. possibly carcinogenic to humans), on the basis that there is sufficient evidence for carcinogenicity in animals, but insufficient evidence for carcinogenicity in humans (IARC, 1999).

In 1993 (and re-affirmed in 2004 and maintained in the 4th edition of the guidelines published in 2011), the World Health Organization (WHO) concluded that the balance of evidence suggests that dichloromethane is a non-genotoxic carcinogen (WHO, 2011).

In 1993 (and retained in 2004 and 2011), the World Health Organization (WHO) derived a Guideline for Drinking-water Quality (GDWQ) for dichloromethane of 0.02 mg/l (20 µg/l; rounded). The GDWQ was based on a NOAEL of 6 mg/kg bw/day identified in a 2-year study in rats. An uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for possible carcinogenicity) was applied to the NOAEL to derive a Tolerable Daily Intake (TDI) of 0.006 mg/kg bw/day (6 µg/kg bw/day). The GDWQ was derived based on

a 60 kg adult drinking 2 litres of water per day with an allocation of 10% of the TDI to water (WHO, 2011).

3.2.2 Toxicity Summary

Dichloromethane is of low acute oral, dermal and inhalation toxicity in experimental animals. The available data on the irritation and sensitisation of dichloromethane to experimental animals indicate the presence of irritating effects to both eyes and skin, and human exposure data indicate that dichloromethane may also be irritating to the respiratory tract. Tests for the genotoxicity of dichloromethane in bacterial assays are generally positive, although the majority of mammalian in vitro and in vivo tests indicate dichloromethane is not genotoxic. In repeated dose studies dichloromethane was observed to induce liver toxicity and the development mammary gland tumours was observed. In carcinogenicity studies in humans no significant evidence of carcinogenicity was observed following exposure to dichloromethane. Limited data are available on the reproductive and developmental effects of dichloromethane, however no strong reproductive effects were reported.

3.3 Derivation of Tolerable Daily Intakes

3.3.1 Oral

In 1993 (and retained in the 4th edition of the guidelines published in 2011), the World Health Organization (WHO) derived a Guideline for Drinking-water Quality (GDWQ) for dichloromethane of 0.02 mg/l (20 µg/l; rounded). The GDWQ was based on a NOAEL of 6 mg/kg bw/day identified in a 2-year study in rats. An uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for possible carcinogenicity) was applied to the NOAEL to derive an oral Tolerable Daily Intake (TDI) of 0.006 mg/kg bw/day (6 µg/kg bw/day) (WHO, 2003; WHO, 2011).

3.3.2 Dermal

No dermal toxicity studies were located upon which a dermal TDI can be derived. However, as there is no evidence to suggest that dermal exposure will significantly impair the integrity of skin (i.e. dichloromethane is not corrosive), a dermal TDI can be derived by extrapolation from oral toxicity data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

Dichloromethane is reported to be absorbed via the dermal route, but at a slower rate than either oral or inhalation exposure. Therefore, using the IGHRC guideline 50% bioavailability via the oral route and 10% bioavailability via the dermal route can be assumed. Therefore a dermal TDI of 30 µg/kg bw/day can be derived.

3.3.3 Inhalation

In a 20-month study, a NOAEC of 200 ppm (approximately 696 mg/m³) was identified in rats administered dichloromethane for 6 hours/day, 5 days/week based on histopathological changes in the liver in females. Adjusting to a continuous exposure, this continuous NOAEC would be equivalent to a concentration of 124 mg/m³. Following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for possible carcinogenicity), an inhalation Tolerable Daily Concentration of 0.124 mg/m³ (124 µg/m³) is identified.

3.4 Review of Current and Historical Usage

Major global industrial applications of dichloromethane include use in paint and varnish strippers, solvent degreasers, aerosols, cleaning fluids, refrigerants, dewaxing solutions, metal cleaners, paints, photo-resistent stripping operations, as a blowing agent in foams and it is used in film processing. It is also used as a solvent for a number of extraction processes and is sometimes mixed with methanol, petroleum naphtha or tetrachloroethylene to increase its effect (WHO, 1996; IPSC, 1996; Government of Canada, 1993; OECD, 2011). Dichloromethane is also used in insecticides (WHO, 1996). It is also reported that dichloromethane has been used historically as an extraction solvent for spice oleoresins, hops, caffeine from coffee and other extraction processes in the food industry (Government of Canada, 1993; OECD, 2011).

In 1984 it was reported that dichloromethane was increasingly used as a replacement for fluorocarbons in aerosols and in fire extinguishing products and as a grain insecticide fumigant (WHO, 1984).

In 2009 100 000 tonnes of dichloromethane was sold within Europe. Dichloromethane is produced alongside other chloromethanes such as methyl chloride, chloroform and carbon tetrachloride. In 2011 the major uses of dichloromethane were: used as a solvent in the pharmaceutical and chemical industry for chemical reactions; and purification and isolation of intermediates (OECD, 2011).

The UK office for National statistics reports import and export tonnages for the UK as presented in Table 3.3.

Table 3.3 Import and export tonnages of dichloromethane in the UK (ONS, 2013)

Year Unit Intra EU Imports Extra EU Imports UK NET supply

2008 tonnes/year 5641 0.75 5642 2009 tonnes/year 7828 0.001 7828 2010 tonnes/year 12740 0.16 12740 2011 tonnes/year 13568 0.03 13568

3.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with dichloromethane is provided in Appendix C2, and only an overall summary is provided here.

Data for dichloromethane are available for drinking water, groundwater, fresh and marine surface water and effluents. Concentrations of dichloromethane in drinking water range from 0.008 to 3600 µg/l, in ground water from 0.0082 to <200 000 µg/l, in fresh surface water from 0 to 743 µg/l, in marine surface water from 0.0022 to 2.6 µg/l and in waste water effluents from 0 to <100 µg/l.

3.6 Review of Literature Data on Removal during Sewage Treatment

There are studies available reporting removal of dichloromethane during sewage treatment indicating that is inherently biodegradable and of moderate toxicity to a range of sewage treatment organisms. The available data are summarised in Table 3.4 and Table 3.5.

Table 3.4 Summary of studies reporting dichloromethane removal during sewage treatment

Concentration Method/type of study Inoculum Results Reference (mg/l)

Aerobic

28 day OECD 301C (Ready biodegradability NR 100 5-26% IUCLID (2000) – Modified MITI test I) Kawasaki (1980) 14 day Japanese MITI activated sludge 100 29% theoretical BOD cited in SRC test (2013b) Klecka (1982)

Closed bottle activated sludge 10 50% CO2 cited in SRC (2013b) Tabak et al. 0-21% depending on 7 day serial shake flask sewage seed 5 - 10 (1981) cited in acclimation period SRC (2013b) sewage seed or Tabak et al. activated sludge Completely NR NR (1981) cited in between 6 hours to biodegradable HSDB (2013) 7 days Kincanon and Biological treatment Stover (1981) activated sewage NR 94.5% simulation - TOC cited in SRC (2013b)

Concentration Method/type of study Inoculum Results Reference (mg/l)

Klecka (1982) Biological treatment activated sewage 1 1.3 mg/l/hour cited in SRC simulation, 14 day (2013b) Klecka (1982) Biological treatment activated sewage 10 3.1 mg/l/hour cited in SRC simulation, 14 day (2013b) Klecka (1982) Biological treatment activated sewage 100 3.6 mg/l/hour cited in SRC simulation, 14 day (2013b)

Static subcultures taken domestic 5-10 100 transformation IPCS (1996) at 14 and 21 days wastewater Respirometer test, activated sludge 10 49% after 50 hours IUCLID (2000) adapted, (domestic) Respirometer test, activated sludge 660 100% after 24 hours IUCLID (2000) adapted, (domestic) Respirometer test, activated sludge 396 100% after 12 hours IUCLID (2000) adapted, (domestic) Respirometer test, activated sludge 264 100% after 12 hours IUCLID (2000) adapted, (domestic)

Respirometer test, activated sludge 132 100% after 12 hours IUCLID (2000) adapted, (domestic) 100% degradation of 1 mg/l with ammonia Biodegradation test (not Sewage organism: 4 67% degradation of IUCLID (2000) specified) Nitrosomonas sp. 1 mg/l without ammonia Sewage organisms: Methylobacteria Methylosinus trichosporium OB3b, Hyphomicrobium, Methylopila, B- Proteobacteria, Methylorhabdus, aerobically degrade Trotsenko and Biodegradation study - Albibacter, NR dichloromethane (no Torgonskaya dehalogenation Methylobacterium values) (2009) dichloromethanicum, Methylopila helvetica, Methylorhabdus multivorans, Methylophilus leisingeri,

Concentration Method/type of study Inoculum Results Reference (mg/l) Paracoccus methylutens, Albibacter methylovorans

Warburg respirometer activated sludge IUCLID (2000) 50 92% after 6 hours test, adapted, (industrial) and SRC (2013b) activated sludge NR except adapted, 25 100% after 20 hours IUCLID (2000) (domestic) activated sludge >99% degradation NR from a continuous- 180 IPCS (1996) after 2-6 days flow reactor municipal activated 49-66% NR sludge (9-11 days 1-100 mineralisation after IPCS (1996) acclimation) 50 hours 100% after 4 hours activated sludge NR NR (ready IUCLID (2000) (domestic) biodegradable) 100% after 4 hours activated sludge NR NR (ready IUCLID (2000) (domestic) biodegradable)

Conventional activated 96-96.3% after activated sludge 0.150 IUCLID (2000) sludge plant 6 hours enriched primary almost complete Static, closed 25 IPCS (1996) sewage effluent transformation biodegradation began after 6 days with Up-flow fix bed reactors overall removal in inoculated with activated excess of 80% sludge for 162 days. between days 62-99. Osuna et al. activated sludge 1 mmol/l/day Simultaneous feeding Sequential alternating (2008) and biodegradation pollutant feeding did monitored. not affect performance of reactor. Anaerobic

Anaerobic digestion with 86-92% Gossett (2013) acclimation, conversion wastewater NR mineralisation after cited in HSDB to carbon dioxide acclimation (2013)

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound

MITI: Ministry of International Trade and Industry, Japan

NR: not reported

TOC: Total Organic Carbon. The total amount of organic carbon in an aqueous solution/suspension

Table 3.5 Summary of studies reporting toxicity of 1,2-dichloropropane to sewage treatment organisms

Method/type of Inoculum Duration Conc. (mg/l) Reference study

100% degradation of 1 mg/l with Biodegradation test IUCLID Nitrosomonas sp. 4 ammonia 67% degradation of 1 mg/l (not specified) (2000) without ammonia

EC50 (effect not IUCLID Nitrosomonas sp. 24 1 stated) (2000) EC50 (effect not IUCLID Pseudomonas putida. 16 500 stated) (2000) NOEL (effect not IUCLID Aeromonas hydrophila NR >19600 stated) (2000) NOEL (effect not IUCLID Bacillus subtilis NR >19600 stated) (2000) EC10 (effect not IUCLID Escherichia coli 16 37.2 stated) (2000)

IC10 (effect not IUCLID Escherichia coli NR 1049 stated) (2000) Methylobacteria IC10 (effect not IUCLID Methylosinus. NR 1468 stated) (2000) smegmatis

EC10: Median Effect Concentration i.e. the concentration that is estimated to cause a specified toxic effect in 10% of exposed organisms

EC50: Median Effect Concentration i.e. the concentration that is estimated to cause a specified toxic effect in 50% of exposed organisms

IC10: Inhibition Concentration. A point estimate of the toxicant concentration that would cause 10% reduction in a non-lethal biological measurement of the test organism e.g. reproduction or growth

NOEL: No Observed Effect Level

NR: Not reported

3.7 Review of Literature Data on Removal During Drinking Water Treatment

Shouli et al. (1992) have determined that primary water treatment (coagulation and flocculation) is not effective at removing dichloromethane during drinking water treatment, with only 16% removal being reported.

The US Environmental Protection Agency (EPA) has reported that Filtrasorb 300 exhibits adsorptive capacities of 1.3 mg and 0.09 mg dichloromethane/g carbon at equilibrium concentrations of 1000 and 100 mg/l, respectively (US EPA, 1985). Speth and Miltner (1990) reported carbon capacities of 0.1 mg/g and 1.2 mg/g for concentrations of 1.2 and 715 µg/l, respectively, at equilibrium using F400.

Fronk (1987) has reported that ozone is ineffective at removing dichloromethane; ozone doses of 2, 6 and 20 mg/l gave dichloromethane removals of 4, 6 and 1%, respectively.

Air stripping is effective at removing dichloromethane during drinking water treatment. In a packed tower, an influent concentration of 770 µg dichloromethane/l was 98% removed by air- stripping using an air to water ratio of 51. Experimental results have shown cascade cross- flow air stripping to be effective in removing tetrachloromethane (Pekin and Moore, 1982; Wood et al., 1990). In a pilot plant utilising air stripping, coagulation and adsorption, an influent groundwater concentration of 48 mg dichloromethane/l was 99.9% removed (Kelly et al., 1981).

Huang et al. (2012) recently demonstrated that zero-valent copper nanoparticles (50 nm, specific surface area of 19 m²/g) may be an effective catalyst for removing dichloromethane under sodium borohydride reducing conditions. A concentration of 26.4 mg dichloromethane/l was 90% reduced after one hour using a copper nanoparticle concentration of 2.5 g/l and a sodium borohydride concentration of 1 g/l.

4. Aniline

4.1 Physico-Chemical Properties

Data on the physico-chemical properties of aniline are provided in Table 4.1.

Table 4.1 Physico-chemical properties of aniline

CAS Number 62-53-3

Chemical Formula C6H7N

Structure

Molecular weight 93.1283 (ChemID, 2013)

Physical state Colourless oily liquid (EU, 2004)

Melting point -6°C (SRC, 2013a)

Boiling point 184.1°C (SRC, 2013a)

Water solubility 36 000 mg/l at 25°C (SRC, 2013a)

Log Kow 0.9 (SRC, 2013a)

Koc No data

Vapour pressure 0.49 mm Hg at 25°C (SRC, 2013a)

Henry’s Law constant 2.02 x10-6 atm.m³/mole at 25°C (SRC, 2013a)

Dissociation constant No data

Density 1.022 g/cm³ at 20°C (EU, 2004)

4.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with aniline is provided in Appendix B3, and only an overall summary is provided here (Table 4.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 4.2 Summary of key toxicological data for aniline

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 195 mg/kg bw (dogs) to 2000 mg/kg bw (rats)

Dermal Low acute toxicity LD50 range: 1290 mg/kg bw (guinea pigs) to 1540 mg/kg bw (rabbits)

Inhalation Low acute toxicity LC50 range: 175 ppm (668 mg/m3; mice) to 950 mg/m3 (rats)

Irritation and - Irritating effects have been observed in both the eyes and the skin of Sensitisation experimental animals. Mild to moderate skin sensitiser. No data are available on the respiratory irritation.

Genotoxicity In vitro and Tests for the genotoxicity of aniline in bacterial assays are generally in vivo negative, although the majority of other in vitro and in vivo tests indicate a weak genotoxicity of aniline. The EU has classified aniline as a Category 3 mutagen (positive in vivo in somatic cells)

Repeat Dose Toxicity Oral Chronic Study: Key study used for assessment: and Carcinogenicity Fischer 344 rats (130/sex/dose) were administered aniline in the diet, at concentrations of 7, 22 and 72 mg/kg bw/day for 104 weeks. A LOAEL of 7 mg/kg bw/day was identified for this study based on haematological and splenic effects.

Dermal No data were located

Inhalation Sub-acute Study: Key study used for assessment: Male Wistar rats (30/dose) were exposed to aniline via nose-only inhalation at concentrations of 0, 9.2, 32.4, 96.5 and 274.9 mg/m3 for 6 hours/day, 5 days/week for 2 weeks, followed by a 2 week recovery period. A NOAEC of 32.4 mg/m³ was identified was identified based on incidence of cyanosis and haematological and splenic effects.

Reproductive and Oral Developmental Study: Developmental Pregnant Fischer 344 rats (21-24/dose) were administered aniline Toxicity hydrochloride by gavage at concentrations of 0, 10, 30 or 100 mg/kg bw/day during gestational days 7-20.

Maternal LOAEL (increased relative spleen weight): 10 mg/kg bw/day (as aniline hydrochloride; 7 mg/kg bw/day as aniline) Developmental NOAEL (pup weight): 30 mg/kg bw/day (as aniline hydrochloride; 21 mg/kg bw/day as aniline)

Dermal No data were located

Inhalation No data were located

4.2.1 Evaluation by Authoritative Bodies

In 1987, the International Agency for Research on Cancer (IARC) evaluated the available data and classified aniline in Group 3, (i.e. it is not classifiable as to its carcinogenicity to humans), stating that 'there are inadequate data relating to its potential carcinogenicity in humans and limited data as to its carcinogenicity in experimental animals' (IARC, 1987).

In 1994, the US Environmental Protection Agency (EPA) and Health Canada evaluated the carcinogenicity data for aniline, classifying it as Group 2B (probable human carcinogen) and Group III (possibly carcinogenic), respectively.

The European Union (EU) have classified aniline as a Category 3 mutagen (positive in vivo in somatic cells) (EU, 2004).

4.2.2 Toxicity Summary

Aniline is of low acute oral, dermal and inhalation toxicity in experimental animals. The available data on the irritation and sensitisation of aniline to experimental animals indicate the presence of irritating effects to both eyes and skin, with longer lasting effects manifesting in eye irritation. The available data on sensitisation indicate that aniline is a mild to moderate skin sensitiser in both experimental animals and humans. Tests for the genotoxicity of aniline in bacterial assays are generally negative, although the majority of other in vitro and in vivo tests indicate a weak genotoxicity of aniline. In repeat dose toxicity studies in experimental animals evidence of spleen toxicity and increases in methaemoglobin levels were observed, however there was little evidence of carcinogenicity. Limited studies are available on the reproductive and developmental toxicity of aniline, and there is no evidence of effects at doses below those causing maternal toxicity.

4.3 Derivation of Tolerable Daily Intakes

4.3.1 Oral

In a 2-year study in rats, a LOAEL of 7 mg/kg bw/day was identified, based on haematological effects at all doses. Using this LOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEL), a Tolerable Daily Intake (TDI) of 0.007 mg/kg bw/day (7 µg/kg bw/day) is derived.

4.3.2 Dermal

No dermal toxicity studies were located upon which a dermal TDI can be derived. Consideration has been made to extrapolate a dermal TDI from oral data following the guidelines developed by the Interdepartmental Group on Health Risks from Chemicals (IGHRC) (IGHRC, 2006). However, the available data on skin and eye irritation indicate that aniline is severely irritating, and may therefore, disrupt the integrity of the skin. Therefore, it is considered inappropriate to extrapolate a dermal TDI.

4.3.3 Inhalation

In a 2-week study in rats, a NOAEC of 32.4 mg/m³ was identified following exposure to aniline for 6 hours/day, 5 days/week based on haematological effects. Adjusting to a continuous exposure, this continuous NOAEC would be equivalent to a concentration of 5.8 mg/m³ (rounded) Following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the short study duration), an inhalation Tolerable Daily Concentration of 0.006 mg/m³ (6 µg/m³; rounded) is derived.

4.4 Review of Current and Historical Usage

Aniline is mainly used as a chemical intermediate, with approximately 71-76% of all manufactured aniline being processed to 4,4'-methylenedianiline (MDA), the starting material for polyurethane plastics (CSTEE, 2003; EU, 2004; OECD, 2000). Aniline is also used in the manufacture of dyestuffs, rubber accelerators, antioxidants, and in solvents, vulcanising agents and resins (CSTEE, 2003; EU, 2004; OECD, 2000). Aniline is also used in the manufacture of varnishes, perfumes, pharmaceuticals and agricultural chemicals, although it is not approved for use in the European Union as a pesticide (EU, 2013; CSTEE, 2003; EU, 2004; OECD, 2000). Small amounts of aniline may occur in coal-tar (RSC, 1991).

The latest figures located for production and usage in Western of aniline are summarised in Table 4.3 (EU, 2004). It is reported that in 1989 there was an annual production capacity in Western Europe of 649 000 tonnes/year with an annual growth rate of 3.5% (EU, 2004; OECD, 2000). In 2003 a production volume for Europe was reported as 650 000 tonnes/year (CSTEE, 2003) and the European Chemicals Bureau reported that actual production volumes exceeded those predicted for 1998 although confidentiality meant actual volumes were not reported (EU, 2004).

Table 4.3 Summary production and usage data reported by the European Chemical Bureau (EU, 2004)

Tonnes/year (% of total European usage) Use 1989 1990 1993 1998 (Predicted)

Production Volume 649 000 500 000 530 000 605 000 Import - 65 000 30 000 50 000 Export - 5 000 3 000 - Usage - - 555 000 652 000 Processing to MDA - - 410 000 (74%) 498 000 (76%) Processing to rubber chemicals - - 88 000 (16%) 90 000 (14%) Processing to dyes - - 38 000 (6.8%) 42 000 (6.4%) Processing to pesticides - - 10 000 (1.8%) 10 000 (1.5%) Processing to others - - 9 000 (1.6%) 12 000 (1.8%)

4.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with aniline is provided in Appendix C3, and only an overall summary is provided here.

Data for aniline is available for drinking water, groundwater, fresh surface water and effluents. Concentrations of aniline in drinking water range from 0 to 0.18 µg/l, in ground water from 0 to 720 µg/l, in fresh surface water from 0 to 12 µg/l and in effluents is identified but not quantifiable. No data were located for aniline in marine surface waters.

4.6 Review of Literature Data on Removal During Sewage Treatment

There are studies available reporting removal of aniline during sewage and data available for the toxicity of aniline to sewage treatment organisms indicate that it is readily biodegradable by a range of organisms. The available data are summarised in Table 4.4 and Table 4.5.

Table 4.4 Summary of studies reporting aniline removal during sewage treatment

Method/type of Concentration Inoculum Results Reference study (mg/l) Aerobic

Activated EC50 (respiration) NR 100 Yoshioka et al. (1986) sludge 14% inhibition of Activated Becarri et al. (1980) NR 0.5 nitrification sludge and Richardson (1985) 84% inhibition of Activated Becarri et al. (1980) NR 10 nitrification sludge and Richardson (1985) degradation Heterotrophs Almost complete degradation following in activated ≤400 MAFF (1997) after 24 hours acclimatisation sludge 91% showed values >40% ThOD after 7days, 96% Respirometer test at non-adapted showed values >60% ThOD 100 MAFF (1997) 20°C sludge after 14 days, and 100% showed values >70% ThOD after 28 days Warburg screening Helfgott et al. (1977) sewage NR 53% BODT depletion test, 30 days at 20˚C cited in SRC (2013b) Warburg screening Activated Malaney (1960) cited 500 56% BODT test, 6 days at 20˚C sludge in SRC (2013b) Warburg screening Activated Lutin et al. (1965) cited 500 25 - 42% BODT depletion test, 6 days at 20˚C sludge in SRC (2013b)

Method/type of Concentration Inoculum Results Reference study (mg/l)

Warburg screening Activated Baird et al. (1977) cited 20 54% depletion test, 4 days at 25˚C sludge in SRC (2013b) Warburg %BODT Activated Symons et al. (1961) screening test, 0.42 200 64% BODT sludge cited in SRC (2013b) days Warburg %BODT Marion and Malaney Activated screening test, 5-7 500 17, 42, 48% BODT (1964) cited in SRC sludge days at 20˚C (2013b) Malaney and Mckinney Warburg screening Activated 500 50 µg/l oxygen uptake (1966) cited in SRC test, 8 days at 25˚C sludge (2013b)

CO2 Screening test Activated 150,450 and Bilyk et al. (1971) cited 100% degradation over 7 days sludge 1000 in SRC (2013b)

CO2 Screening test Activated Lyons et al. (1984) 250 100% degradation in 4 days over 7 days sludge cited in SRC (2013b) BOD Screening test Branson (1978) cited in NR NR 68% degradation over 5 days SRC (2013b) COD % removal Activated 94.5% COD removal and 19.0 Pitter (1976) cited in screening test over 5 200 sludge mg COD/g/hr SRC (2013b) days

Schefer and Waelchli %DOC elimination Activated 50 (DOC) 80 - 100% DOC elimination (1980) cited in SRC screening test sludge (2013b) DOC Screening test Brown and Laboureur (modified OECD) sewage 20 >90% degradation (1983) cited in SRC over 28 days (2013b) %BODT Screening test (modified OECD) over Calamari et al. (1980) sewage 10, 50, 100 47.8. 57.8, 66.7% BODT 14 days, cited in SRC (2013b) acclimatisation of 15 days, 20˚C %BODT Screening Heukelekian and Rand test (modified sewage 1.5 - 3 61%, 73% BODT (1955) cited in SRC OECD) over 5 days, (2013b) 20˚C %BODT Screening test (modified Calamari et al. (1980) OECD) over sewage 10, 50, 100 101.8, 70.5, 83.2% BODT cited in SRC (2013b) 14 days, no acclimatisation, 20˚C

Method/type of Concentration Inoculum Results Reference study (mg/l)

Freitag et al. (1982), CO2 screening test, activated 0.05 37% CO2 Korte and Klein (1982) 5 days, 25˚C sludge cited in SRC (2013b) Zahn-Wellens Gerike and Fischer activated %DOC screening 400 100% DOC after 3 days (1979) cited in SRC sludge test, 14 days, 25˚C (2013b) Zahn-Wellens % Zahn and Wellens degradation activated 90% degradation in 4 days and NR (1980) cited in SRC screening test, 28 sludge 22%COD per day (2013b) days Zahn-Wellens activated Liu (1983) cited in SRC %DOC screening 100 - 400 4.2 days (half-life) sludge (2013b) test, 19 days, 22˚C Sewage Japanese MITI (microbial OECD (1979) cited in %BOD screening 50 36 – 47% degradation population SRC (2013b) test, 14 days, 25˚C 30 mg/l) Japanese MITI Gerike and Fischer activated %BODT screening 50 99% BODT (1979) cited in SRC sludge test, 14 days, 25˚C (2013b)

Japanese MITI Kawasaki (1980) and activated %BODT screening 100 30 - 100% BODT Kitano (1978) cited in sludge test, 14 days, 25˚C SRC (2013b) Japanese MITI %BODT screening King and Painter activated test, 28 days with NR 64 - 115% BODT (1983) cited in SRC sludge acclimatisation 2-4 (2013b) days STURM CO2 Gerike and Fischer 92% CO2 and 98% DOC evolution screening sewage 10 (1979) cited in SRC removal test, 28 days (2013b)

Gerike and Fischer OECD screening sewage 3 - 20 93% DOC (1979) cited in SRC test, 19 days (2013b) OECD screening activated Liu (1983) cited in SRC 3 - 5 7 day half life DOC test, 19 days sludge (2013b) Gerike and Fischer Closed bottle (1979) and Gerike %BODT screening sewage 2 90% BODT (1984) cited in SRC test, 30 days (2013b)

Method/type of Concentration Inoculum Results Reference study (mg/l)

BLOK %BODT Gerike (1984) cited in screening test, 42 sewage 50 92% BODT SRC (2013b) days OECD %DOC Activated OECD (1979) cited in removal screening 5 - 40 81 - 100% DOC removal sludge SRC (2013b) test, 19 days OECD %DOC Gerike (1984) cited in removal screening sewage 5 - 40 93% DOC removal SRC (2013b) test, 42 days

UV removal Hallas and Alexander 100% disappearance of initial screening test, sewage 10 (1983) cited in SRC UV absorbance 14 days (2013b) An aerobic activated sludge unit. Biomass from aerobic Removal efficiencies more activated Gheewala and activated sludge was 350 than 95% with insignificant sludge Annachhatre (1997) acclimatized to inhibition of nitrification synthetic wastewater containing aniline Anaerobic

53 day study 10% degradation forming Hallas and Alexander including 28 day lag sewage NR acetanilide and 2- (1983) cited in HSDB with inoculum methylquinoline as products (2013) anaerobic reactor in inoculum Chou et al. (1979) 110 days with a 2-10 maintained NR no degradation cited in HSDB (2013) day retention time on acetate anaerobic conditions phenol- Razo-Fores et al. with a lag time of Aniline was not biodegraded in adapted NR (1996) and (1997) greater than a sludge cited in HSDB (2013) 150 days

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound

BODT: Biological Oxygen Demand with regard to time. The amount of oxygen consumed by microorganisms when metabolising a compound in a specific time

COD: Chemical Oxygen Demand. The amount of oxygen consumed during oxidation of a compound with hot acid dichromate; it provides a measure of the oxidisable matter present

DOC: Dissolved Organic Carbon. The amount of carbon present in organic compounds in aqueous solution

MITI: Ministry of International Trade and Industry, Japan

NR: Not reported

OECD: The Organisation for Economic Co-operation and Development. Standardised methods of testing for assessing the potential effects of chemicals on human health and the environment.

UV: Ultraviolet

Table 4.5 Summary of studies reporting toxicity of aniline to sewage treatment organisms

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

100% degradation Pseudomonas sp. 10 150 Wu et al. (2012) EC0 (cell multiplication) Pseudomonas putida 16 130 MAFF (1997) 75% inhibition of ammonia Nitrifying bacteria from a Tomlinson et al. NR 7.7 oxidation pilot scale (1966) 75-100% inhibition of Richardson Nitrosomonas sp NR 100 ammonia oxidation (1985) 100% degradation of aniline Dietzia natronolimnaea 120 100 Jin et al. (2012) 80.5% degradation of Dietzia natronolimnaea 120 300 Jin et al. (2012) aniline 72% degradation of aniline Dietzia natronolimnaea 120 500 Jin et al. (2012)

EC0: Median Effect Concentration i.e. the concentration that is estimated to cause a specified toxic effect in 0% of exposed organisms

NR: Not reported

4.7 Review of Literature Data on Removal during Drinking Water Treatment

Aniline is reported to undergo removal during drinking water treatment by activated carbon. An initial concentration of 100 mg/l was reported to undergo 55% adsorption onto Norit Granular Activated Carbon (GAC) at a dose of 500 mg/l (Faria et al., 2007). Another study has reported that the adsorption capacity for activated carbon was 272 mg aniline/g activated carbon using an initial aniline concentration of 465 mg/l (Laszlo et al., 2007).

Fielding et al. (1989) have demonstrated that aniline will react rapidly with chlorine during drinking water treatment. Using an aniline concentration of 10 mg/l with 1 mg chlorine/l at pH 7, approximately 90% of the chlorine dose was consumed within 1 hour. Aniline initially forms N-chloroaniline and chloroaniline isomers, which may undergo oxidation or polymerisation processes in the presence of excess chlorine.

Aniline is reported to react extremely quickly with ozone; Chan and Larson (1991a) reported that an aniline concentration of 9 mg/l was decomposed by 10 mg ozone/l, however, in this study it was also noted that the reaction products that were produced were of greater toxicological concern than aniline. In another study, an initial aniline concentration of 100 mg/l was completely removed following application of an ozone dose of 10 mg/l/minute for

20 minutes at pH 7. The rate of removal was reported to be increased further in the presence of Norit GAC 1240 Plus (500 mg/l) (Faria et al., 2007). Complete removal of an aniline concentration of 100 mg/l was reported after 10 minutes of ozonation at a dose of 80 mg ozone/l at pH 3. The rate of aniline removal was reported to increase in the presence of ferrous iron and/or UV irradiation (Sualeda and Brillas, 2001). In wastewater, an aniline concentration of 2 mg/l was reduced to 0.2 mg/l with an ozone dose of 1156 mg/l at pH 10 and a contact time of 70 minutes (Sarasa et al., 1998).

Chan and Larson (1995) have reported that the products of a reaction between ozone and aniline that may have potential concern with regards to potential mutagenicity included azobenzene, azoxybenzene, benzidine, nitrosobenzene, nitrobenzene, 4-aminodiphenylamine and phenazine. As the aniline to ozone ratio decreases, the yield (relative to the amount of aniline) of the dimer azobenzene increases. However, the presence of fulvic acid appears to allow the removal of aniline without the formation of these breakdown products (Chan and Larson, 1995).

In another study, Chan and Larson (1991b) also reported that o-, m- and p-nitro-aniline were formed in the presence of nitrite ion. Yields of up to 8% were reported at an aniline concentration of 1 x10-4 M (approximately 9 mg/l) and an ozone dose of 2.1 x10-4 M (approximately 10 mg/l), with yields higher in acidic or neutral conditions than in alkaline conditions.

In an ozonation study using an aniline dose of 1 mmol/l (approximately 93 mg/l) and an ozone dose of 1 mg/l/minute at pH 6.3, 98% degradation was reported within 15 minutes, with 3-dioxopropanoic acid, 2-oxopropanedioic acid, oxalic acid and glyoxal identified as reaction products (Gilbert, 1983).

Aniline may undergo some removal by reverse osmosis, but it is poorly removed by nanofiltration. A concentration of 100 mg/l was 63% removed by an ESPA1 reverse osmosis, but 20% removed by an NF 200 nanofiltration membrane (Ben-David et al., 2006).

5. Benzylchloride

5.1 Physico-Chemical Properties

Data on the physico-chemical properties of benzylchloride are provided in Table 5.1.

Table 5.1 Physico-chemical properties of benzylchloride

CAS Number 100-44-7

Chemical Formula C7H7Cl

Structure

Molecular weight 126.5853 (ChemID, 2013)

Physical state Liquid (OECD, 1998)

Melting point -45°C (SRC, 2013a)

Boiling point 179°C (SRC, 2013a)

Water solubility 525 mg/l at 25°C (SRC, 2013a)

Log Kow 2.3 (SRC, 2013a)

Koc No data

Vapour pressure 1.23 mm Hg (SRC, 2013a)

Henry’s Law constant 0.000412 atm.m³/mole at 25°C (SRC, 2013a)

Dissociation constant No ionisable functional groups (OECD, 1998)

Density No data

5.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with benzylchloride is provided in Appendix B4, and only an overall summary is provided here (Table 5.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 5.2 Summary of key toxicological data for benzylchloride

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 340 mg/kg bw (rats) to 1620 mg/kg bw (mice)

Dermal Low to moderate acute toxicity LD50: >145 mg/kg bw (rabbits); LDL0: 10 ml/kg bw (guinea pigs)

Inhalation Moderate acute toxicity LC50 range: 1.79 mg/l (rats) to 2080 mg/m3 (cats)

Irritation and - Corrosive to the skin and irritating to the eyes and respiratory tract of Sensitisation experimental animals. Skin sensitiser.

Genotoxicity In vitro and Both negative and positive results have been reported for benzylchloride in in in vivo vitro genotoxicity assays. Overall the weight of evidence indicates that benzylchloride is genotoxic in vitro. However, the weight of evidence indicates that benzylchloride is non genotoxic in vivo.

Repeat Dose Oral Chronic Study: Key study used for assessment: Toxicity and F344 rats (52/sex/dose) were administered benzylchloride by oral gavage at 0, Carcinogenicity 15 or 30 mg/kg bw twice per week (reported to be 0, 6.4 and 12.9 mg/kg bw/day, respectively, adjusting to a 7 day/week dosing regimen) for 2-years. A NOAEL of 15 mg/kg bw/day (reported to be 6.4 mg/kg bw/day, adjusting to a 7 day/week dosing regimen) was identified for this study based on forestomach tumours.

Dermal None of the studies located were considered appropriate for use in this assessment

Inhalation Sub-acute Study: Key study used for assessment: Male Swiss OF1 mice (10 per dose and 5 as control) were exposed to benzylchloride at 0, 22 and 46 ppm (reported to be 0, 107 and 224 mg/m³, respectively) by whole-body inhalation for 6 hours per day for 4, 9 or 14 days. A NOEL of 22 ppm (reported to be 107 mg/m³) was identified based on lesions in respiratory and olfactory epithelia.

Reproductive and Oral Developmental Study: Developmental Pregnant female Sprague-Dawley rats (8 per dose) were orally administered Toxicity benzylchloride at 0, 50 and 100 mg/kg bw/day on days 6-15 of gestation.

Maternal NOAEL: 100 mg/kg bw/day (highest dose tested) Foetal NOAEL (reduced foetal length): 50 mg/kg bw/day Developmental NOAEL: 100 mg/kg bw/day (highest dose tested)

Dermal No data were located

Inhalation No data were located

5.2.1 Evaluations by Authoritative Bodies

The International Agency for Research on Cancer (IARC) has classified benzylchloride as a group 2A carcinogen (probably carcinogenic to humans). This is a group classification for combined exposure to α-chlorinated toluenes (benzal chloride, benzotrichloride, benzylchloride) and benzoyl chloride on the basis that there is limited evidence for the carcinogenicity of α-chlorinated toluenes and benzoyl chloride in humans and sufficient evidence for the carcinogenicity of benzylchloride in experimental animals (IARC, 1999). In 1994 the US Environmental Protection Agency (US EPA) classified benzylchloride as a group B2 carcinogen (probable human carcinogen) based on their being sufficient evidence of carcinogenicity in animals (US EPA, 1994).

5.2.2 Summary

Benzylchloride is of low acute oral toxicity, of low to moderate dermal toxicity and of moderate inhalation toxicity. It is corrosive, a slight eye irritant and a skin sensitiser in experimental animals. Overall, the data indicate that benzylchloride is genotoxic in vitro but not genotoxic in vivo. Benzylchloride has been shown to be carcinogenic in experimental animals, and there is limited evidence from epidemiological studies for the carcinogenicity of α-chlorinated toluenes and benzoyl chloride in humans; therefore, IARC has issued a group classification of 2A (probably carcinogenic to humans) for combined exposure to α-chlorinated toluenes (benzal chloride, benzotrichloride, benzylchloride) and benzoyl chloride. No reproductive studies were located for benzylchloride; however there is limited evidence of effects on the embryo and foetus from the two available oral developmental studies. There is evidence for neurotoxic and immunotoxic effects of benzylchloride in experimental animals, however further experiments are needed to confirm the significance of these findings.

5.3 Derivation of Tolerable Daily Intakes

5.3.1 Oral

In a 2-year study, rats were administered benzylchloride at doses of 0, 15 or 30 mg/kg bw, 2 days/week (reported to be 0, 6.2 and 12.9 mg/kg bw/day, respectively, adjusting to a 7 day/week dosing regimen). A NOAEL of 15 mg/kg bw (reported to be 6.2 mg/kg bw/day) was identified based on forestomach tumours. Using this NOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the limited oral toxicity database) a Tolerable Daily Intake (TDI) of 0.006 mg/kg bw/day (6 µg/kg bw/day; rounded) is derived.

It should be noted that as humans do not have a forestomach, the relevance of this effect as the basis of a TDI is questionable. However, it is probable that these tumours arose as secondary effect of irritation of the forestomach. (as there was no evidence of a significant increase at the low dose) Therefore, there is potential for high concentrations of benzylchloride to illicit irritancy in the human gastrointestinal system. Therefore, in the absence of additional robust oral toxicity data, this TDI is proposed.

5.3.2 Dermal

Although several dermal toxicity studies were located for benzylchloride, these studies were primarily focussed on the potential for tumour initiation, rather than the determination of a no- effect-level. As such, they are not considered appropriate for the derivation of a dermal TDI.

Consideration has been made to extrapolate a dermal TDI from oral data following the guidelines developed by the Interdepartmental Group on Health Risks from Chemicals (IGHRC) (IGHRC, 2006). However, the available data on skin irritation indicate that benzylchloride is corrosive. Therefore, it is considered inappropriate to extrapolate a dermal TDI.

5.3.3 Inhalation

In a 14-day study, mice were exposed to benzylchloride at concentrations of 0, 22 or 46 ppm (reported to be 0, 107 and 224 mg/m³, respectively) for 6 hours/day. A NOAEC of 22 ppm (reported to be 107 mg/m³) was identified, based on lesions in respiratory and olfactory epithelia. Adjusting to a continuous exposure, this continuous NOAEC would be equivalent to a concentration of 27 mg/m³ (rounded). Following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the short study duration and possible carcinogenicity), an inhalation Tolerable Daily Concentration of 0.027 mg/m³ (27 µg/m³; rounded) is derived.

5.4 Review of Current and Historical Usage

Benzylchloride is widely used in the production of other substances such as plastics, dyes, lubricants, petrol, photographic developer, flavour products and pharmaceuticals. It has previously been used as an irritant gas in chemical warfare. Benzylchloride can also be used in the manufacture of synthetic tannins (US EPA, 2013, HSDB, 2013).

It was reported in 1988 it was used in the manufacture of bactericides, fungicides, insecticides, flavours, odourants, plastics and plasticisers. It was reported that in 1991 more than two thirds of the benzyl chloride produced was used in the manufacture of benzyl butyl acetate and flexible plastics used for applications such as food packing.

Data of usage or production of benzylchloride in Europe or the UK is limited. In 1993 an annual production and import tonnage for Japan was 7759 tonnes/year. It was reported to be used as an intermediate in a closed system for synthesis organics such as dyes, benzyl alcohol and perfumes (OECD, 1998). Environment Canada and Health Canada (2009) reported in 2009 that in Canada benzyl chloride was only used as an intermediate chemical in the manufacture of benzalkonium chloride. In 2006 the annual usage of benzyl chloride in Canada was 100 to 1 000 tonnes (Environment Canada and Health Canada 2009).

5.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with benzylchloride is provided in Appendix C4, and only an overall summary is provided here.

Data for benzylchloride are available for fresh surface water and effluent. Concentrations of benzylchloride in fresh surface water and effluent were detected but not quantifiable. No data were located for drinking water, groundwater or marine surface water.

5.6 Review of Literature Data on Removal During Sewage Treatment

There are three studies available reporting removal of benzylchloride during sewage treatment indicating that it is readily biodegradable. There are no data available for the toxicity of benzylchloride to sewage treatment organisms. The available data are summarised in Table 5.3.

Table 5.3 Summary of studies reporting benzylchloride removal during sewage treatment

Concentration Method/type of study Inoculum Results Reference (mg/l) Aerobic

Kitano (1978), and Readily biodegradable Japanese MITI protocol NR NR Sasaki (1978) both (no value reported) cited in HSDB (2013) Japanese MITI protocol NR NR 70.9% after 4 weeks OECD (1998) OECD TG 301C 2-day incubation and Raw sewage Jacobson and Biodegraded significantly acclimated to non- and raw NR Alexander (1981) (no value reported) chlorinated compounds sewage cited in HSDB (2013)

MITI: Ministry of International Trade and Industry, Japan

NR: Not reported

5.7 Review of Literature Data on Removal during Drinking Water Treatment

No data were located on the removal of benzylchloride during drinking water treatment. However, benzylchloride undergoes rapid hydrolysis in water to benzyl alcohol; half-lives of 10.0, 9.48 and 9.64 hours have been reported at pH 4, 7 and 9, respectively (OECD, 1998).

6. 1,3-Butadiene

6.1 Physico-Chemical Properties

Data on the physico-chemical properties of 1,3-butadiene are provided in Table 6.1.

Table 6.1 Physico-chemical properties of 1,3-butadiene

CAS Number 106-99-0

Chemical Formula C4H6

Structure

Molecular weight 54.0914 (ChemID, 2013)

Physical state Colourless gas (EU, 2002)

Melting point -108.9°C (SRC, 2013)

Boiling point -4.4°C (SRC, 2013)

Water solubility 735 mg/l at 25°C (SRC, 2013)

Log Kow 1.99 (SRC, 2013)

Koc Log Koc: 1.86-2.36 (Environment Canada, 2000)

Vapour pressure 2110 mm Hg at 25°C (SRC, 2013)

Henry’s Law constant 0.0736 atm.m³/mole at 25°C (SRC, 2013)

Dissociation constant No data

Density 0.65 g/cm³ at -6°C (EU, 2002)

6.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with 1,3-butadiene is provided in Appendix B5, and only an overall summary is provided here (Table 6.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 6.2 Summary of key toxicological data for 1,3-butadiene

Route or Endpoint Summary study type

Acute toxicity Oral No data were located

Dermal No data were located

Inhalation Low acute toxicity LC50 range: 270 mg/m3 (mice) to 554 mg/m3 (rabbits)

Irritation and - No indications of injury to the eye were noted in dogs and rabbits following Sensitisation exposure to 6700 ppm 1,3-butadiene. Acute exposure to butadiene at the higher concentrations of 90 000–140 000 ppm was reported to cause conjunctivitis and respiratory tract irritation in mice, and exposure to 150 000- 250 000 ppm was reported to cause conjunctivitis and lacrimation in rabbits.

Genotoxicity In vitro and There is evidence of genotoxic potential in mammalian cells, however it in vivo should be noted that the EU has assessed the genotoxicity studies in mammalian cells as being of poor quality, particularly with regards to ensuring adequate exposure to the test material. It should also be noted that positive results have been reported with the three major metabolites of 1,3-butadiene (epoxybutene, diepoxybutane and butanediol) in a variety of in vitro test systems, where diepoxybutane is reported to be the most mutagenic.

A significant amount of data is available regarding the in vivo genotoxicity of 1,3-butadiene. Overall, these data indicate that 1,3-butadiene is genotoxic in vivo in mice, but not in the rat, hamster, monkey or drosophila. This is likely to be due to differences in 1,3-butadiene metabolism between these species.

Repeat Dose Oral No data were located Toxicity and Dermal No data were located Carcinogenicity Inhalation Chronic Study: Key study used for assessment: Male and female B6C3F1 mice (70/sex/group, apart from the top dose where 90 animals/sex/group were used) were exposed to 1,3-butadiene at 0, 6.25, 20, 62.5, 200 or 625 ppm (approximately 0, 13.9, 44.4, 139, 444 or 1390 mg/m³, respectively) for 6 hours/day, 5 days/week for 2 years (approximately 0, 4.46, 14.3, 44.6, 143 and 446 ppm, respectively, adjusted to a 7 day/ week dosing regimen). A LOAEC of 6.25 ppm (13.9 mg/m³) was identified in females was identified based on neoplastic lesions.

Reproductive and Oral No data were located Developmental Dermal No data were located Toxicity Inhalation Developmental Study: Female CD-1 mice, which had been mated with unexposed males, were exposed to 1,3-butadiene at 0, 40, 200 or 1000 ppm (approximately 0, 89, 444 and 2218 mg/m³, respectively) for 6 hours per day, on days 6-15 of gestation by whole-body inhalation.

Route or Endpoint Summary study type Maternal NOAEC (maternal body weight and gravid uterus weight): 40 ppm (approximately 89 mg/m³) Foetal LOAEC (reduce weight of male foetuses): 40 ppm (approximately 89 mg/m³) Developmental NOAEC: 1000 ppm (approximately 2218 mg/m³; the highest dose tested)

6.2.1 Evaluations by Authoritative Bodies

In 2008, the International Agency for Research on Cancer (IARC) classified 1,3-butadiene as a Group 1 carcinogen (carcinogenic to humans) on the basis of there being sufficient evidence for carcinogenicity in both experimental animals and humans (IARC, 2008). The US Environmental Protection Agency (US EPA) has classified 1,3-butadiene as carcinogenic to humans by inhalation based similarly on an assessment of their being sufficient evidence for the carcinogenicity of 1,3-butadiene (US EPA, 2002).

6.2.2 Summary

1,3-Butadiene is of low acute oral and inhalation toxicity, although the acute inhalation studies are considered to be of poor quality. No dermal studies were located. Due to the gaseous nature of 1,3-butadiene, it is difficult to carry out conventional skin and eye irritation tests; however, the data available indicate it to be irritating to the eyes and respiratory tract of both experimental animals and humans. Conflicting results have been obtained from in vitro genotoxicity studies, however 1,3-butadiene is clearly genotoxic in vivo in the mouse, but not the rat. There is also evidence for genotoxicity to humans, and for the genotoxicity of 1,3-butadiene metabolites both in vitro and in vivo. 1,3-Butadiene does not appear to be teratogenic in either the rat or the mouse via inhalation exposure, but there is evidence of slight foetotoxicity in the mouse. The observed different sensitivities between species are thought to be due to differences in 1,3-butadiene metabolism, where there is also evidence for inter-individual variation within a human population. 1,3-Butadiene has been classified as a Group 1 carcinogen (carcinogenic to humans) by IARC, on the basis of there being sufficient evidence for carcinogenicity in both experimental animals and humans.

6.3 Derivation of Tolerable Daily Intakes

6.3.1 Oral

No oral toxicity studies were located upon which an oral Tolerable Daily Intake (TDI) can be derived. However, a number of inhalation studies are available and an oral TDI can be derived by extrapolation from these data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

In 2002, using Benchmark Dosing, the US Environmental Protection Agency (EPA) derived a BMCL10 of 0.88 ppm (human equivalent concentration; HEC) (approximately 2 mg/m³) (US EPA, 2002). Assuming an inhalation volume of 0.043 m³ per day and a bodyweight of 0.038 kg (38 g), this would equate to a systemic BMD10 of 2.3 mg/kg bw/day.

Following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to extrapolate from an effect level to a no effect level) and multiplying by 100/50 (an arbitrary assumption within the IGHRC guidelines to account for the differences between oral and inhalation absorption), an oral Tolerable Daily Intake (TDI) of 0.005 mg/kg bw/day (5 µg/kg bw/day; rounded) can be derived.

It should be noted that this approach is likely to be conservative, as absorption via inhalation is likely to be at least as efficient, but may conceivably be more efficient, that oral exposure.

However, 1,3-butadiene is classified as a Group 1 carcinogen by the International Agency for Research on Cancer (IARC). Therefore, its concentration in water should be as low as reasonably practicable.

6.3.2 Dermal

No dermal toxicity studies were located upon which a dermal TDI can be derived, and it is not appropriate to extrapolate a dermal TDI from inhalation toxicity data, as there are no guidance available on conducting such extrapolations.

However, 1,3-butadiene is classified as a Group 1 carcinogen by IARC. Therefore, its concentration in water should be as low as reasonably practicable.

6.3.3 Inhalation

In 2002, the US Environmental Protection Agency (EPA) derived an inhalation Reference Concentration (RfC) of 0.0009 ppm (0.002 mg/m³). This RfC was based on a study in mice were exposed to concentrations of 1,3-butadiene of at 0, 6.25, 20, 62.5, 200 or 625 ppm (approximately 0, 13.9, 44.4, 139, 444 or 1390 mg/m3, respectively) for 6 hours/day, 5 days/week for 2-years and considered ovarian atrophy to be the critical endpoint. A Benchmark Dosing approach was adopted to identify a BMC10 of 1 ppm and a BMCL10 of 0.88 ppm (human equivalent concentration; HEC). A rounded uncertainty factor of 1000 (3 to account for inter-species variation, 10 to account for intra-species variation, 3 to account for a limited and 10 to extrapolate from an effect-level to a no-effect-level) was applied to the BMCL10 to derive the RfC (US EPA, 2002).

However, 1,3-butadiene is classified as a Group 1 carcinogen by IARC. Therefore, its concentration in water should be as low as reasonably practicable.

6.4 Review of Current and Historical Usage

6.4.1 Uses

The major global uses of 1,3-butadiene includes the manufacture of synthetic rubber (styrene- butadiene rubber (SBR) and polybutadiene rubber) and thermoplastic resins (acrylonitrile- butadiene styrene (ABS) and styrene-butadiene latex). Polybutadiene rubber is used in tyres, tyre products and car body sealants, and ABS is used in the production of oil resistant gaskets, business equipment and automotive parts (EU, 2002).

1,3-Butadiene is also used as an intermediate in the production of neoprene for automotive and industrial rubber goods, in the production of methylmethacrylate-butadiene-styrene (MBS) polymer, which is used as a PVC reinforcing agent, and for producing adiponitrile, a nylon precursor (EU, 2002).

1,3-Butadiene is also either used in the production of or is a component of plastic bottles, food wrap, epoxy resins, lubricating oils, hoses, drive belts, moulded rubber goods, adhesives, paint, latex foams for carpet backing or underpad, shoe soles, moulded toys/household goods, medical devices, and chewing gum (WHO, 2001).

6.4.2 Production and consumption

In 2002, the European Union Risk Assessment report (EU RAR) stated that there are 22 producers of 1,3-butadiene within the European Union (EU), with a total production capacity between 1,202 000 and 4,960 000 tonnes/year (EU, 2002). Table 6.3 summarises the production data reported in the EU RAR for Western Europe.

Table 6.3 Summary of Western European production data (EU, 2002)

Year Production tonnes/year

1991 1 778 000 1992 1 853 000 1993 1 742 000 to 1 752 000 1994 1 892 000

It is reported that 96% of 1,3-butadiene used is used in the manufacture of synthetic rubber and plastic products with the remainder being used as an intermediate in other compounds (OECD, 1996). Table 6.4 shows the estimated percentage of the total usage in Europe per application type.

Table 6.4 Estimated percentage of total 1,3-Butadiene used in Europe per application type (EU, 2002)

Estimated amount used Application Percentage of total use (tonnes/year)

Styrene-butadiene rubber/latex 56 1 059 520 Polybutadiene rubber 22 416 240 Chloroprene rubber 6 113 520 Nitrile-butadiene rubber/latex 4 75 680 Acrylonitrile-butadiene-styrene resin 4 75 680

Hexamethylenediamine 4 75 680 Other uses 4 75 680

The production capacity of synthetic rubber products in Europe by type is summarised in Table 6.5 and the estimated consumption of each product type in Western Europe is summarised in Table 6.6.

Table 6.5 Production capacity of synthetic rubber products by country in 1994 (EU, 2002)

SBR XSBR & NBR NBR Country SBR solid BR Chloroprene Latex PSBR solid latex

Austria - - 6 000 - - - - Belgium 20 000 - 15 000 20 000 - - - Finland - - 89 400 - - - - France 139 000 48 000 87 000 155 000 44 000 2 000 40 000 Italy 175 000 20 000 65 000 80 000 30 000 10 000 - Netherlands 150 000 23 000 75 000 - - 15 000 - Spain 50 000 - 21 000 20 000 - - - Sweden - - 32 000 - - - -

United Kingdom 150 000 35 000 92 500 80 000 10 000 5 000 33 000 Germany 255 000 62 000 202 000 101 000 40 000 20 000 60 000 Total 939 000 188 000 684 900 456 000 124 000 52 000 133 000

BR: Polymerised polybutadiene rubber

CR: Polymerised polychloropropene rubber

SBR: Styrene-butadiene rubber

NBR: Nitrile-butadiene rubber

PSBR: Pyridine vinyl styrene-butadiene rubber

XSBR: Carboxylated styrene-butadiene rubber

Table 6.6 Estimated consumption of synthetic rubber products in Western Europe (EU, 2002)

Estimated consumption (tonnes/year) Product 1993 1994 1998

Styrene-butadiene rubber solid 542 000 551 000 596 000 Styrene-butadiene rubber latex 123 000 121 000 126 000 Carboxylated styrene-butadiene rubber 510 000 520 000 560 000

Polybutadiene rubber 265 000 270 000 292 000 Polychloropropene rubber 61 000 62 000 65 000 Nitrile-butadiene rubber 80 000 81 000 87 000

In the United States (US) it is reported that large scale 1,3-butadiene production commenced during World War II. By 1965 the annual production of 1,3-butadiene had risen to 1 224 699 tonnes/year (2.7 billion pounds) and in 1974 the annual production was 1 678 292 tonnes/year (3.7 billion pounds). During the 1980s the market fluctuated but increased again by the 1990s to 2 041 166 tonnes/year (4.5 billion pounds). By 2012 it was reported that the estimated annual production in the US was 2 721 554 tonnes/year (6 billion pounds) (ATSDR, 2012). In the US approximately 60% of the total consumption of 1,3-butadiene is used in the manufacture of synthetic rubber in the automotive industry (ATSDR, 2012). Table 6.7 shows the typical usage of 1,3-butadiene in the US in 2012.

Table 6.7 Estimated breakdown of usage of 1,3-butadiene in the US (ATSDR, 2012)

Usage type Estimated proportion used (%)

Styrene-butadiene rubber 30

Synthetic rubbers Polybutadiene 25 (61% of total usage) Polychloroprene 4

Nitrile elastomer 2 Adiponitrile/hexamethylene diamine (HMDA) 11 Styrene-butadiene latex 12 ABS resins 5 Other uses 11

6.4.3 Other sources

1,3-Butadiene is also released from combustion of biomass especially from forest fires but also from internal combustion engines as a result of incomplete combustion of fossil fuels, natural gas, oil, wood, cigarettes and waste incineration. Total global emissions of

1,3-butadiene from forest fires has been estimated to be approximately 770 000 tonnes/year (WHO, 2001). Table 6.8 summarises the estimated releases of 1,3-butadiene into the Canadian environment during 1994.

Table 6.8 Estimated releases of 1,3-butadiene into the Canadian environment in 1994 (WHO, 2001)

Estimated release of 1,3-butadiene Source (tonnes/year)

Combustion engines (including on/off road vehicles, 3751-9489 lawnmowers, marine sector and rail sector

Chemical and chemical production 270.4 Plastics products industries 17.5 Refined petroleum and coal products industry 22.3 Sent for waste disposal (incineration, landfill, sewage 128.7 treatment) Industrial on-site uses 225.8 Fuel distribution 24 Forest burning 1191 Wood space heating 3706

Natural oil/gas space heating 11 Cigarettes 1-9

ATSDR (2012) reported estimations of the amount of 1,3-butadiene emissions to the US environment in 2005 which showed that petrol driven on-road vehicles and non-vehicular engines had the highest contribution to the environment out of all the potential sources at 45.59% of the total. The second highest contributing source was classed as ―miscellaneous sources‖ which are source that could not be otherwise classified at 33.39%. The remaining 21.02% includes sources such as wildfires, industrial processes, petroleum refineries, waste disposal, fuel combustion and logging/slash burning.

6.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with 1,3-butadiene is provided in Appendix C5, and only an overall summary is provided here.

Data for 1,3-butadiene is available for drinking water, groundwater, fresh surface water and effluents. Concentrations of 1,3-butadiene in drinking water and groundwater are detected but not quantifiable, in fresh surface water from 0 to 2 µg/l and in effluents from 2 to 130 µg/l. No data were located for marine surface water.

6.6 Review of Literature Data on Removal during Sewage Treatment

There are no studies available reporting removal of 1,3-butadiene during sewage treatment and toxicity to sewage treatment organisms. However, when 1,3-butadiene is inoculated with methane utilizing organisms, it was found to biodegrade to 1,3-butadiene monoepoxide (or epoxybutene). Therefore it is concluded that a biological sewage system should be able to degrade 1,3-butadiene, given suitable acclimation conditions (Toronto, 2013).

6.7 Review of Literature Data on Removal during Drinking Water Treatment

No data were located on the removal of 1,3-butadiene during drinking water treatment, however, some predictions on the fate of this chemical in drinking water treatment can be made based on its physico-chemical properties.

A log Koc of 1.86-2.36 (Environment Canada, 2000) has been reported for 1,3-butadiene, which would suggest that it has high mobility in the water column and therefore is unlikely to be amenable to removal by GAC.

1,3-Butadiene is highly volatile; a vapour pressure of 2110 mm Hg and a Henry‘s Law constant of 0.0736 atm.m³/mole have been reported at 25°C (SRC, 2013). This may indicate that 1,3-butadiene will be amendable to removal from water by air stripping.

In a study conducted in air, Chou et al. (2005) determined that greater than 90% decomposition of a 1,3-butadiene concentration of 50 ppm could be achieved at 26°C and a gas retention time of 85 seconds with a molar ratio of ozone to 1,3-butadiene of 3.5. A similar removal was also achieved with UV/ozone with a molar ratio of UV/ozone to 1,3-butadiene of 2. Therefore, similar oxidation technologies may also be effective at removing 1,3-butadiene during drinking water treatment.

7. 1,1-Dichloroethane

7.1 Physico-Chemical Properties

Data on the physico-chemical properties of 1,1-dichloroethane are provided in Table 7.1.

Table 7.1 Physico-chemical properties of 1,1-dichloroethane

CAS Number 75-34-3

Chemical Formula C2H4Cl2

Structure

Molecular weight 98.9596 (ChemID, 2013)

Physical state Liquid (IUCLID, 2000)

Melting point -96.9°C (SRC, 2013)

Boiling point 57.4°C (SRC, 2013)

Water solubility 5040 mg/l at 25°C (SRC, 2013)

Log Kow 1.79 (SRC, 2013)

Koc No data

Vapour pressure 227 mm Hg at 25°C (SRC, 2013)

Henry’s Law constant 0.00562 atm.m³/mole at 24°C (SRC, 2013)

Dissociation constant No data

Density 1.174 g/cm³ at 20°C (IUCLID, 2000)

7.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with 1,1-dichloroethane is provided in Appendix B6, and only an overall summary is provided here (Table 7.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 7.2 Summary of key toxicological data for 1,1-dichloroethane

Route or Endpoint Summary study type

Acute toxicity Oral Low to moderate acute toxicity LD50 range: 14.1-14 100 mg/kg bw (rats)

Dermal No data were located

Inhalation Low acute toxicity LC50 range: 13 000 ppm (rats) to 17 300 ppm (mice)

Irritation and - Irritating to the skin and irritating to the eyes of experimental animals. Sensitisation No data are available on either respiratory irritation or sensitisation.

Genotoxicity In vitro and Tests for the genotoxicity of 1,1-dichloroethane give equivocal results in vitro, in vivo and there is very limited evidence available for genotoxicity in vivo.

Repeat Dose Oral Chronic Study: Key study used for assessment: Toxicity and Osborne-Mendel rats (50/sex/dose) and B6C3F1 mice (50/sex/dose) were Carcinogenicity administered 1,1-dichloroethane by gavage in corn oil at concentrations of 350 and 700 mg/kg bw/day and 750 and 1500 mg/kg bw/day in rats for males and females, respectively, and 900 and 1800 mg/kg bw/day in mice. Animals were dosed 5 days/week for 78 weeks, followed by an observation period of 33 weeks for rats and 13 weeks for mice. A time weighted average LOAEL of 475 mg/kg bw/day was identified in rats based on dose-dependent increases in mortality and clinical signs of toxicity.

Dermal No data were located

Inhalation Sub-chronic Study: Key study used for assessment: In a sub-chronic inhalation study 10 rats, 10 guinea pigs, 4 cats and 4 rabbits were administered 1,1-dichloroethane in inhalation chambers at a concentration of 2025 mg/m³, 6 hours/day, 5 days/week for 13 weeks. These animals were then administered 4050 mg/m3 for a further 13 weeks. A NOAEL of 2025 mg/m3 was identified for cats based on increase in serum urea and creatinine.

Reproductive and Oral No data were located Developmental Dermal No data were located Toxicity Inhalation Developmental Study: Mated female Sprague-Dawley rats (group sizes not reported) were administered 1,1-dichloroethane in inhalation chambers at concentrations of 0, 15 400 and 24 300 mg/m³ for 7 hours/day on days 6 to 15 of gestation. Maternal and foetal NOAEL (decreased maternal food consumption and weight gain, increase in delayed ossification of sternebrae in foetuses): 15 400 mg/m³

7.2.1 Evaluations by Authoritative Bodies

In 1993, and reaffirmed in 2004, the World Health Organization evaluated 1,1-dichloroethane, but, due to the very limited data available on its toxicity and carcinogenicity, no Guideline for Drinking-water Quality (GDWQ) was proposed.

In 1996, the US Environmental Protection Agency (EPA) evaluated 1,1-dichloroethane and classified it as Group C, i.e. a possible human carcinogen. This classification was based on a lack of human data and limited evidence of carcinogenicity in rats and mice.

7.2.2 Toxicity Summary

1,1-Dichloroethane is of low to moderate acute oral toxicity and is of low acute inhalation toxicity in experimental animals. No data were available on the acute dermal toxicity of 1,1-dichloroethane in experimental animals. The extremely limited data on the irritation and sensitisation of 1,1-dichloroethane to experimental animals indicate the presence of irritating effects to both eyes and skin. In vitro genotoxicity tests in bacteria were mostly negative, although there was evidence of genotoxicity in mammalian cells with metabolic activation. Very limited data were available on the in vivo genotoxicity of 1,1-dichloroethane. Based on the limited data available, some evidence of the carcinogenicity of 1,1-dichloroethane was observed. The one study of developmental and reproductive toxicity of 1,1-dichloroethane was equivocal.

7.3 Derivation of Tolerable Daily Intakes

7.3.1 Oral

In a chronic study in rats, a time-weighted LOAEL of 475 mg/kg bw/day was identified based on dose-dependent increases in mortality and clinical signs of toxicity. Using this LOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to account for the limited oral toxicity data), an oral Tolerable Daily Intake (TDI) of 0.475 mg/kg bw/day (475 µg/kg bw/day) is derived.

7.3.2 Dermal

No dermal toxicity studies were located upon which a dermal TDI can be derived. Although a single irritation study was located that indicated irritancy effects, these appear to be slight and reversible. Therefore, a dermal TDI can be derived by extrapolation from oral toxicity data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

1,1-Dichloroethane is reported to be absorbed via both the oral and the dermal route, however, no quantitative data on absorption via the dermal route were located. Therefore, equivalent bioavailability for the oral and dermal routes can be assumed. Therefore a dermal TDI of 0.475 mg/kg bw/day (475 µg/kg bw/day) can be derived.

However, it should be noted that this is likely to be a highly conservative value, as in reality, very few substances are absorbed as readily via the dermal route as they are via the oral route.

7.3.3 Inhalation

In a sub-chronic inhalation study, a NOAEC of 2025 mg/m³ was identified in cats exposed to 1,1-dichloroethane for 6 hours/day, 5 days/week based on histopathological lesions in kidney tubules and decreased body weight. Adjusting to a continuous exposure, this NOAEC would be equivalent to a concentration of 362 mg/m³ (rounded). Using this continuous NOAEC and following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to account for the limited inhalation database), a Tolerable Daily Concentration of 0.362 mg/m³ (3620 µg/m³; rounded) is derived.

7.4 Review of Current and Historical Usage

1,1-Dichloroethane is mostly used as an intermediate in the production of 1,1,1-trichloroethane, vinyl chloride and other chemicals. It is also used as a solvent in paint and varnish removers, as a degreaser and cleaning agent and in ore flotation (WHO, 1996). It is used as a coupling agent in antiknock petrol and It is also used in metal degreasing and organic synthesis (Verschueren, 1996, ATSDR, 1990). 1,1-Dichloroethane has previously been used as an anaesthetic in humans, but its use was discontinued due to the occurrence of cardiac arrhythmias at concentrations required for anaesthesia (>100 000 mg/m³) (year not reported) (WHO, 2003).

There is little data on the production volumes of 1,1-dichloroethane in Europe or the USA. It is reported that in 1977 in the USA 45 500 tonnes of 1,1-dichloroethane was produced. No data post 1977 is available (ATSDR, 1990).

7.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with 1,1-dichloroethane is provided in Appendix C6, and only an overall summary is provided here.

Data for 1,1-dichlorethane are available for drinking water, groundwater, fresh surface water and effluents. Concentrations of 1,1-dichlorethane in drinking water range from 0.105 to 24 µg/l, in ground water from 0.0081 to 1900 µg/l, in fresh surface water from 0 to 400 µg/land in effluents from 0.5 to 6300 µg/l. No data were located for marine surface water.

7.6 Review of Literature Data on Removal during Sewage Treatment

There are studies available reporting removal of 1,1-dichloroethane during sewage treatment indicating that it is readily biodegradable and of moderate to high toxicity to sewage treatment organisms. The available data are summarised in Table 7.3 and Table 7.4.

Table 7.3 Summary of studies reporting 1,1-dichloroethane removal during sewage treatment

Method/type of Concentration Inoculum Results Reference study (mg/l) Aerobic

Aerobic study, activated IUCLID 0.1 97% biodegraded duration not stated sludge (2000) sewage IUCLID NR 5 91 % degraded after 28 days sludge (2000)

sewage IUCLID NR 10 83 % degraded after 28 days sludge (2000) sewage 50% degraded after 7 days. Losses due Howard NR 5 sludge to evaporation were 19% (1990) sewage 27% degraded after 7 days. Losses due Howard NR 10 sludge to evaporation were 4% (1990)

NR: Not reported

Table 7.4 Summary of studies reporting toxicity of 1,1-dichloroethane to sewage treatment organisms

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

EC50 (inhibition of Nitrosomonas sp. NR 0.91 IUCLID (2000) nitrification) EC50 (inhibition of methane Methanogenic 48 6.2 IUCLID (2000) production) bacteria Vargas and Ahlert NOEC (inhibition) Anaerobic bacteria NR 3.5 (1987)

EC50: Median Effect Concentration i.e. the concentration that is estimated to cause a specified toxic effect in 50% of exposed organisms

NOEC: No Observed Effect Concentration. The highest concentration of a material in a toxicity test that has no statistically significant adverse effects on the exposed population with the controls

NR: Not reported

7.7 Review of Literature Data on Removal during Drinking Water Treatment

In a pilot plant trial, the GAC absorption capacity for an influent concentration of 9 µg 1,1-dichloroethane/l and an empty bed contact time of 2.5 minutes was 0.12-0.15 mg/g at saturation (Qi et al., 1992). Speth and Miltner (1990) have calculated carbon capacities of F400 using concentrations of 13.4 and 559 µg 1,1-dichloromethane/l of 0.4 and 5.62 mg/g,

respectively, at equilibrium. Urano et al. (1991) calculated loadings on F400 of 0.7 mg/g and 14.3 mg/g at equilibrium concentrations of 10 μg/l and 1000 μg/l, respectively. However, Urano et al., (1991) also noted that the presence of humic substances could reduce the loadings by 20-30%. Using F300, Dobbs and Cohen (1980) reported isotherms of 0.05, 0.16 and 0.53 mg/g using 1,1-dichloroethane concentrations of 1, 10 and 100 µg/l, respectively, in distilled water.

Fronk (1987) has reported that ozone doses of 2 and 6 mg/l were ineffective at removing 1,1,-dichloroethane. Chen et al. (2006) reported that a 1,1-dichloroethane concentration of 15 µg/l treated with an ozone dose of 1.7 mg/l at pH 8.2 underwent 20% removal after 40 minutes. However, an 11 µg/l solution underwent 90% degradation in the presence of a combination of ozone and hydrogen peroxide (ozone dose of 10 mg/l and a hydrogen peroxide dose of 3.4 mg/l).

1,1-Dichloroethane is also amenable to removal by aeration; a concentration of 6 µg 1,1-dichloroethane/l in groundwater underwent 83% removal following a 10 minute contact time and an air to water ratio of 4:1 (Love and Eilner, 1982).

8. Nitrobenzene

8.1 Physico-Chemical Properties

Data on the physico-chemical properties of nitrobenzene are provided in Table 8.1.

Table 8.1 Physico-chemical properties of nitrobenzene

CAS Number 98-95-3

Chemical Formula C6H5NO2 Structure

Molecular weight 123.11 (ChemID, 2013)

Physical state Colourless to pale yellow oily liquid (WHO, 2003)

Melting point 5.7°C (SRC, 2013a)

Boiling point 210.8°C (SRC, 2013a)

Water solubility 2090 mg/l at 25°C (SRC, 2013a)

Log Kow 1.85 (SRC, 2013a)

Koc Log Koc: 1.56 (WHO, 2003)

Vapour pressure 0.245 mm Hg (SRC, 2013a)

Henry’s Law constant 2.4 x10-5 atm.m³/mole at 25°C (SRC, 2013a)

Dissociation constant No data

Density 1.2037 g/cm³ at 20°C (WHO, 2003)

8.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with nitrobenzene is provided in Appendix B7, and only an overall summary is provided here (Table 8.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 8.2 Summary of key toxicological data for nitrobenzene

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50: 349 mg/kg bw (rats) LDL0: 1000 mg/kg bw (guinea pigs, rabbits, cats)

Dermal Low acute toxicity LD50 range: 480 mg/kg bw (mice) to 2100 mg/kg bw (rats)

Inhalation Low acute toxicity LC50: 556 ppm (reported to be 2847 mg/m³; rats)

Irritation and - Non-irritating to slightly irritating to the skin and eyes of experimental animals. Sensitisation Non-sensitising.

Genotoxicity In vitro and The available data indicate that nitrobenzene is not genotoxic in bacteria. The in vivo results of genotoxicity studies in mammalian in vitro systems have been mixed. Overall, these data indicate that nitrobenzene may be very weakly genotoxic in mammalian in vitro systems. The results of in vivo genotoxicity studies of nitrobenzene are mixed and the data indicate that nitrobenzene may be weakly genotoxic in vivo.

Repeat Dose Oral Sub-acute Study: Key study used for assessment: Toxicity and Male and female F344 rats (6 per group) were administered nitrobenzene by Carcinogenicity oral gavage at 0, 5, 25 and 125 mg/kg bw/day for 28 days. Additional groups of animals, exposed to 0 or 125 mg/kg bw/day, were allowed to recover for a period of 2 weeks prior to examination. A LOEL of 5 mg/kg bw/day can be identified from this study based on organ weights changes.

Dermal Sub-chronic Study: Key study used for assessment: Fischer 344 rats and B6C3F1 mice (10 per sex per group) were treated with nitrobenzene at 0, 50, 100, 200, 400 and 800 mg/kg bw/day, in acetone, for 90 days by skin painting. A LOEL of 50 mg/kg bw/day was identified in rats based on congestion of the spleen. A LOEL of 50 mg/kg bw/day was identified in mice based on alteration in hepatocytes.

Inhalation Chronic Study: Key study used for assessment: In a chronic inhalation study, female B6C3F1 mice were exposed to nitrobenzene at concentrations of 0, 5, 25 or 50 ppm (approximately 0, 25, 126 or 252 mg/m³, respectively), and male and female F344 rats and male CD rats were exposed to nitrobenzene at 0, 1, 5 or 25 ppm (approximately 0, 5, 25 or 126 mg/m³, respectively). Exposure was for 6 hours/day, 5 days/week (excluding holidays), for a total of 505 days over 2 years. A LOAEC of 1 ppm (approximately 5 mg/m³) was identified in rats based on methaemoglobin.

Route or Endpoint Summary study type

Reproductive and Oral Reproductive Study: Developmental Male and female SPF Sprague-Dawley rats (10 per sex per group) were Toxicity administered nitrobenzene, dissolved in 10% sesame oil, at 0, 20, 60 and 100 mg/kg bw by oral gavage. Animals were treated once per day prior to mating (14 days), during mating (up to 14 days), during gestation (22 days) and until day 3 of lactation. Parental NOAEL (testicular pathology in males): 20 mg/kg bw Reproductive NOAEL: 100 mg/kg bw (highest dose tested) Offspring LOAEL (reduced male pup body weight): 20 mg/kg bw

Dermal F344 rats and B6C3F1 mice (10 per sex per group) were dermally treated with nitrobenzene at 0, 50, 200 and 400 mg/kg bw/day for 13 weeks. Decreases in sperm motility, sperm density and the percentage of abnormal sperm were observed.

Inhalation Reproductive Study: In a two-generational reproductive study, male and female Sprague-Dawley rats (30 per sex per group) were exposed to nitrobenzene vapours at 0, 1, 10 and 40 ppm (approximately 0, 5, 50 and 202 mg/m³, respectively) for 6 hours per day, 5 days per week for 10 weeks, and then females were mated with the males. Following the mating period all males were sacrificed and pregnant females were again exposed to nitrobenzene from days 0-19 of gestation. Reproductive NOEL: 10 ppm (approximately 50 mg/m3)

Developmental Study: Pregnant female Sprague-Dawley rats (26/dose) were exposed to nitrobenzene vapours at 0, 1, 10 and 40 ppm (approximately 0, 5, 50 and 202 mg/m³, respectively) for 6 hours per day, on gestation days 6-15. Matenal NOAEL (bodyweight and spleen weight): 1 ppm (approximately 5 mg/m³) Developmental NOAEL: 40 ppm (approximately 5 mg/m³; highest dose tested) Foetal NOAEL: 40 ppm (approximately 5 mg/m³; highest dose tested)

8.2.1 Evaluations by Authoritative Bodies

The International Agency for Research on Cancer (IARC) has classified nitrobenzene as a group 2B carcinogen (possibly carcinogenic to humans), based on inadequate evidence in humans for the carcinogenicity of nitrobenzene and sufficient evidence in experimental animals for the carcinogenicity of nitrobenzene (IARC, 1996). In 2009, the US Environmental Protection Agency (US EPA) characterised nitrobenzene as ―likely to be carcinogenic to humans‖ (US EPA, 2009).

8.2.2 Summary

Nitrobenzene is of low acute oral, dermal and inhalation toxicity. It is reported to be non- irritating to slightly irritating to the skin and eyes of experimental animals, but is non- sensitising. Results of both in vitro and in vivo genotoxicity testing are mixed, indicating that nitrobenzene may be weakly genotoxic. Nitrobenzene has been shown to be carcinogenic in experimental animals, and has been classified as a group 2B carcinogen (possibly carcinogenic to humans) by IARC. The available data indicate that nitrobenzene affects fertility in experimental animals, probably due to its effects in male reproductive organs where these effects include testicular atrophy as well as reduced sperm concentrations and sperm motility. In the EU, nitrobenzene has been given the harmonised classification of H361F (suspected of damaging fertility). Another significant effect of nitrobenzene exposure is methaemoglobinaemia, where this effect has been observed in humans as well as experimental animals.

8.3 Derivation of Tolerable Daily Intakes

8.3.1 Oral

In 2009, the World Health Organization (WHO) derived a short-term health-based value for nitrobenzene of 0.03 mg/l (30 µg/l). This short-term value was based on a LOAEL of 5 mg/kg bw/day identified in a 28-day study on rats. An uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEL) was applied to the LOAEL to derive an oral Tolerable Daily Intake (TDI) of 0.005 mg/kg bw/day (5 µg/kg bw/day) (WHO, 2009; WHO, 2011).

8.3.2 Dermal

In skin painting studies, LOELs of 50 mg/kg bw/day were identified in rats and mice, based on congestion of the spleen in rats and methaemoglobin and lung congestion in mice. Using these LOELs and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for limitations in the dermal toxicity data) a dermal TDI of 0.05 mg/kg bw/day (50 µg/kg bw/day) is derived.

8.3.3 Inhalation

In a chronic inhalation study, female B6C3F1 mice were exposed to nitrobenzene at concentrations of 0, 5, 25 or 50 ppm (approximately 0, 25, 126 or 252 mg/m3, respectively), and male and female F344 rats and male CD rats were exposed to nitrobenzene at 0, 1, 5 or 25 ppm (approximately 0, 5, 25 or 126 mg/m3, respectively). Exposure was for 6 hours/day, 5 days/week (excluding holidays), for a total of 505 days over 2 years. A LOAEC of 1 ppm (approximately 5 mg/m³) was identified in rats based on methaemoglobin. Adjusting to a continuous exposure, this LOAEC would be equivalent to a concentration of 0.7 mg/m³ (rounded). Using this continuous LOAEC and following application of an uncertainty factor of

1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEC), a Tolerable Daily Concentration of 0.0007 mg/m³ (0.7 µg/m³) is derived.

8.4 Review of Current and Historical Usage

8.4.1 Uses

Within the European Union (EU) nitrobenzene is primarily used in the production of aniline with small additional uses in the production of pharmaceuticals and various other chemicals. There is no natural source of nitrobenzene although it may be formed by hydroxyl radical initiated photooxidation of benzene. Historically nitrobenzene was used as a solvent but this does not appear to be the case in the present (European Chemicals Bureau, 2007). Table 8.3 presents the estimated usage of nitrobenzene by application type (EU, 2007).

Table 8.3 Estimated usage per application type in the EU (EU, 2007)

Types of Use Tonnes/year Approximate % used

Processing to aniline 1 162 900 99 Processing to pharmaceuticals 9 300 0.8 Processing to other chemicals 2 800 0.2 Total 1 175 000 100

In Germany nitrobenzene was used for perfuming soaps and was called Mirbanoil, however the use of nitrobenzene in cosmetic products has been banned since the 1980‘s. Data on its use in soaps or its status as an ingredient in cosmetics in other countries are not available (EU, 2007).

There are other uses of nitrobenzene globally but these do not appear to be relevant within the EU (EU, 2007). These include: use as a solvent in petroleum refining; for manufacturing polyurethanes and dichloroanilines; and the synthesis of organic compounds such as acetaminophen (ingredient in over the counter analgesic) (ATSDR, 1990; WHO, 2003; US EPA, 2003).

It is reported that early in the 20th century nitrobenzene was used as an additive in food as a substitute for almond essence, and had extensive use in boot polish, inks (including inks for stamping clean baby nappies in hospitals) and disinfectants (WHO, 2003).

Outside of the EU it is reported that nitrobenzene is used in, floor and furniture polish, shoe polish and metal polish manufacture, pyroxylin compounds and some paints (EU, 2007; ATSDR, 1990). It is also used in solvent recovery plants, in rubber chemicals, in photographic chemicals, refining lubricant oils, as a solvent in trinitrotoluene (TNT) production, a solvent for cellulose ethers and in cellulose acetate manufacturing (Sax, 1985; Verschueren, 1996; HSDB, 2013).

8.4.2 Production

There are 8 production/processing sites within the EU and it is reported that the quantity of nitrobenzene produced within the EU is 1 180 000 tonnes/year (estimated from 2000 and 2002 data (EU, 2007). It is not known that any nitrobenzene is imported into or exported outside the EU.

In the United Kingdom (UK) there are two production sites for nitrobenzene with a total capacity of 167 000 tonnes/year. If it is assumed that 98% of the nitrobenzene in the UK is used to make aniline then 155,600 tonnes/year is used in the UK (WHO, 2003).

Table 8.4 Production capacities for European countries in 1985 (WHO, 2003)

Country Production capacity (tonnes/year)

Belgium 200 000 Germany 240 000 Italy 18 000 Portugal 70 000 Switzerland 5 000 United Kingdom 145 000

Global production of nitrobenzene in 1994 was estimated at 2 133 800 tonnes/year; about one-third was produced in the United States. In the US a gradual increase in production volume from 73 000 tonnes/year to 740 000 tonnes/year between 1960 and 1994 was reported (ATSDR, 1990; WHO, 2003). In Japan production was estimated to be 70 000 tonnes/year in 1980 and 135 000 tonnes/year in 1990. In India in 1989 an estimated to be 22 000 tonnes/year (WHO, 2003).

In 2003, a notable global increase in the production of nitrobenzene over the previous 30- 40 years was reported (WHO, 2003). It was also estimated that post-1987, production would continue to increase as the market for aniline increased. There were four nitrobenzene producers in the US in 1990 (ATSDR, 1990).

8.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with nitrobenzene is provided in Appendix C7, and only an overall summary is provided here.

Data for nitrobenzene are available for drinking water, groundwater, fresh and marine surface water and effluents. Concentrations of nitrobenzene in drinking water are detected but not quantifiable, in ground water range from 0 to 12 µg/l, in fresh surface water from 0.022 to <100 µg/l, in marine surface water range from 0 to 14.8 µg/l and in effluents from 0 to 91 000 µg/l.

8.6 Review of Literature Data on Removal during Sewage Treatment

There are studies available reporting removal of nitrobenzene during sewage treatment indicating that it is readily biodegradable and of low toxicity to a range of sewage treatment organisms. The available data are summarised in Table 8.5 and Table 8.6.

Table 8.5 Summary of studies reporting nitrobenzene removal during sewage treatment

Concentration Method/type of study Inoculum Results Reference (mg/l)

Aerobic

99.6% degradation after Davis et al. (1981) Warburg screening test sewage 50 6 days cited in SRC (2013b)

500 mg/l O2 uptake after Lutin et al. (1965) Warburg screening test activated sludge 500 6 days cited in SRC (2013b)

300 mg/l O2 uptake after Malaney (1960) cited Warburg screening test activated sludge 300 5-8 days in SRC (2013b) Malaney and 150 mg/l O2 uptake after Warburg screening test activated sludge 500 Mckinney (1966) 0.33 days cited in SRC (2013b)

Marion and Malaney 500 mg/l O2 uptake after Warburg screening test activated sludge 500 (1966) cited in SRC 5-7 days (2013b) Hallas and 100% loss of initial Rotary shaker sewage 10 Alexander (1983) absorbance after 14days cited in SRC (2013b) Hallas and 50% loss of initial Screening test sewage 10 Alexander (1983) absorbance after 14days cited in SRC (2013b) Heukelekian and BOD5 sewage 1.7-20 0% BOD after 5 days Rand (1955) cited in SRC (2013b)

Kincannon and Biological treatment 97% biodegradation activated sludge 0.01 – 0.012 Stover (1981) cited simulation after 8 hours in SRC (2013b) Kincannon et al Biological treatment 97.8% biodegradation activated sludge 100 (1983) cited in SRC simulation after 72 hours (2013b) Stover and Biological treatment 76 – 97.5% treatment activated sludge 100 Kincannon (1983) simulation efficiency 2 – 6 days cited in SRC (2013b)

Concentration Method/type of study Inoculum Results Reference (mg/l)

0 – 29% BOD after 14 Kawasaki (1980) Japanese MITI at 25˚C activated sludge 100 days cited in SRC (2013b) 0 – 29% BOD after 14 Kitano (1978) cited Japanese MITI at 30˚C activated sludge 100 days in SRC (2013b) Korte and Klein

Screening test activated sludge NR 1.7% CO2 after 5 days (1982) cited in SRC (2013b)

98% removal of CO2 Pitter (1976) cited in Screening test activated sludge 200 after 5 days SRC (2013b)

100% degradation after Tabak et al. (1981) Serial shake flask test sewage 5 7 days cited in SRC (2013b) 87 -100% degradation Tabak et al. (1981) Serial shake flask test sewage 10 after 7 days cited in SRC (2013b) OECD 301D Ready Biodegradability: Closed domestic 3 >90% after 30 days IUCLID (2000) bottle test with an initial Klebsiella ornithinolytica concentration 600 the Biodegradation study 60 Wang et al. (2012) NB1 biodegradation rate is 9.29 mg/l/hour

86% after 10 days, loss OECD 302B Modified attributed to activated sludge NR IUCLID (2000) Zahn-Wellens test evaporation, rather than biodegradation 100% after 21 days, loss Directive 79/831/EEC, filtered effluent from a attributed to physico- Annex V, C.3 Biotic municipal sewage NR chemical properties IUCLID (2000) degradation - modified treatment plant rather than OECD screening test biodegradation Manometric respirometry activated sludge 60 52% after 35 days IUCLID (2000) Sapromat respirometry test activated sludge 180 50% after 12 days IUCLID (2000) Sapromat respirometry test activated sludge 70 57% after 8 days IUCLID (2000) no decrease in treatment efficiency of sludge. Only at a Artificial municipal Gomolka and artificial activated sludge 5 – 6 g/m3 nitrobenzene wastewater Gomolka (1979) concentration of 25 g/m3 does the sludge index rise to 236 cm3/g.

Concentration Method/type of study Inoculum Results Reference (mg/l)

no inhibition by nitrobenzene. feeding Sequence aerobic batch granular sludge ≤600 time and shock loading Zhao et al. (2011) reactor had little effect on removal TOC measured using Zheng et al. (2009) isolated bacterium from Complete degradation of Micrococcus luteus <70 and Zhao et al. activated sludge with 2, 3 100 (2011) and 5% sodium chloride

TOC measured using Zheng et al. (2009) isolated bacterium from Complete degradation of Micrococcus luteus <96 and Zhao et al. activated sludge with 2, 3 150 (2011) and 5% sodium chloride TOC measured using Zheng et al. (2009) isolated bacterium from Complete degradation of Micrococcus luteus <120 and Zhao et al. activated sludge with 2, 3 200 (2011) and 5% sodium chloride TOC measured using Zheng et al. (2009) isolated bacterium from Complete degradation of Micrococcus luteus <196 and Zhao et al. activated sludge with 2, 3 250 (2011) and 5% sodium chloride TOC measured using Zheng et al. (2009) isolated bacterium from 43.32% degradation of Micrococcus luteus <120 and Zhao et al. activated sludge with 7% 200 (2011) sodium chloride Capable of degrading, Zheng et al. (2009) 2% sodium chloride without Streptomyces sp. NR but concentration not and Zhao et al. inhibition reported. (2011) Bacillus, Pseudomonas, Comamonas, Corynebacterium, Stapylococcus, Steptococcus, Klebsiella, Capable of degrading, Zheng et al. (2009) Acinetobacter, NR NR but concentration not and Zhao et al. Flavobacterium, reported. (2011) Alcaligenes, the fungi Rhodotorula milaginosa and the white rot fungi Phanerochaete chrysosporium

Concentration Method/type of study Inoculum Results Reference (mg/l) Anaerobic Chou et al. (1978) acclimated domestic 81% removed in 110 Anaerobic reactor NR cited in ATSDR sludge days (1990) Hallas and acclimated domestic 50% was degraded in 12 Alexander (1983) Anaerobic reactor NR sludge days cited in ATSDR (1990) 8% decrease in nitrobenzene after 8 Canton et al. (1985). Anaerobic test (not unadapted media NR days and half-life of less cited in ATSDR specified) than 2 weeks in adapted (1990) media.

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound

BOD5: Biological Oxygen Demand in 5 days. The amount of oxygen consumed by microorganisms when metabolising a compound in 5 days

MITI: Ministry of International Trade and Industry, Japan

NR: Not reported

Table 8.6 Summary of studies reporting toxicity of nitrobenzene to sewage treatment organisms

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

EC50 (inhibition of ammonia Blum & Speece (1991) Nitrosomonas sp. NR 0.92 consumption) cited in WHO (2003) Blum & Speece (1991) EC50 (gas production) Methanogen species NR 13 cited in WHO (2003)

Bringmann & Kühn protozoan Entosiphon NOEC (effect not stated) 72 1.9 (1980) cited in WHO sulcatum (2003) Ciliate (Tetrahymena Yoshioka et al. (1985) EC50 (growth rate) 24 98 pyriformis) Cited in WHO (2003) Ciliate (Tetrahymena Yoshioka et al. (1985) EC50 (growth rate) 48 106 pyriformis) Cited in WHO (2003) Ciliate (Tetrahymena Schultz (1981) cited in EC50 (growth rate) 60 83 - 246 pyriformis) US EPA (2013)

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

Bringmann and Kuhn LOEC (cell multiplication) Pseudomonas putida 16 7 (1980) Cited in WHO (2003) 100% degradation analysed Pseudomonas sp.a3 10 300 Wu et al. (2012) by liquid chromatography Yoshioka et al. (1986) EC50 (respiration) activated sludge 3 100 cited in WHO (2003) Volskay and Grady EC50 (respiration) activated sludge 0.5 320 (1988) cited in WHO (2003) Blum & Speece (1991) EC50 (oxygen consumption) activated sludge NR 370 cited in WHO (2003)

EC50: Median Effect Concentration i.e. the concentration that is estimated to cause a specified toxic effect in 50% of exposed organisms

LOEC: Lowest observed effect concentration. The lowest concentration of a material in a toxicity test that has a statistically significant adverse effect on the exposed population compared with the controls

NOEC: No observed effect concentration. The highest concentration of a material in a toxicity test that has no statistically significant adverse effects on the exposed population with the controls

NR: Not reported

TOC: Total Organic Carbon. The total amount of organic carbon in an aqueous solution/suspension

8.7 Review of Literature Data on Removal during Drinking Water Treatment

Slow sand filtration is expected to result in significant removal of nitrobenzene; in a pilot plant study, filtration (0.2 m/hour) of an initial concentration of 85 ng nitrobenzene/l was reported to produce removals of 70-80% (Hrubec et al., 1991).

At an equilibrium concentration of 5 mg/l, the isotherm absorption capacity for nitrobenzene on F100 activated carbon is reported to be approximately 60 mg/g. A concentration of 100 mg/l was effectively completely removed by addition of 10 g activated carbon/l (Rauthala et al., 2011).

A study conducted by Ma et al. (2005) indicates that ozone is not an effective treatment for nitrobenzene during drinking water treatment. Using an ozone dose of 3 mg/l for 8 minutes over a pH range of 2-12 and a nitrobenzene solution of 36 µg/l, >80% removal was only reported at pH 10 and above. In another study, an initial concentration of 740 µg nitrobenzene/l was >90% reduced by treatment with an ozone dose of 1.5 mg/l.minute for 15 minutes (Latifoglu et al., 2003).

The half-life of a 100 mg nitrobenzene/l solution treated with a high dose of ozone (approximately 350 mg/l) was reported to be approximately 30 minutes. The rate was unaffected by the application of UV irradiation or addition of iron(II) (Contreras et al., 2001).

Beltran et al. (1998) conducted a study of the removal of nitrobenzene with ozone, UV and hydrogen peroxide. UV irradiation (1.8 W/l) of a 17 mg/l solution gave about 40% conversion in 40 minutes. In combination with an ozone dose of 36 mg/l.minute, complete removal occurred in 10 minutes. Complete removal occurred after 40 minutes using ozone alone, and ozone with hydrogen peroxide gave a similar removal as ozone alone.

Chen et al. (2005) reported that an initial nitrobenzene solution of 64 mg/l underwent 30% removal after 2 hours under a 1 kW medium pressure mercury lamp. In another study, under a 1.2 eV xenon eximer lamp, an initial concentration of 500 mg nitrobenzene/l was approximately halved after 30 minutes, and in the presence of 970 mg hydrogen peroxide/l, approximately 90% removal was reported (Li et al., 2006).

UV irradiation (12 W) in the presence of 10 mg iron(II)/l and 65 mg hydrogen peroxide/l gave complete removal from a 100 mg/l solution after 30 minutes with a half-life of approximately 5 minutes (Al Momami, 2006).

A combination of vacuum UV (185 nm), titanium dioxide and ozone is reported to be highly effective at nitrobenzene removal; a 50 µg/l solution was completely removed in 60 seconds in deionised water and a solution of 170 µg/l was completely removed within 2 minutes. However, the rate is reduced in the presence of bicarbonate and humic acid; concentrations of 2 mmol bicarbonate/l and 3.2 mg humic acid/l reduced the rate constant by 82.9 and 71.6%, respectively (Yin and Zhang, 2009).

Kiso et al. (2001) reported removals of 42-70% using 10-50 mg nitrobenzene/l solutions with four different nanofiltration membranes. However, Van der Bruggen et al. (1999) reported a maximum removal of 22% using a 350 mg/l solution with four different nanofiltration membranes.

Over a 150 hour period of treatment of 10 mg nitrobenzene/l in 2000- 6000 mg sodium chloride/l using an FT-30 reverse osmosis membrane, removals of 20-60% (average 42%) were reported, whilst sodium chloride rejection was 95-99% (Urama and Marinas, 1997).

9. Oxirane methyl

9.1 Physico-Chemical Properties

Data on the physico-chemical properties of oxirane methyl are provided in Table 9.1.

Table 9.1 Physico-chemical properties of oxirane methyl

CAS Number 75-56-9

Chemical Formula C3H6O

Structure

Molecular weight 58.0794 (ChemID, 2013)

Physical state Colourless liquid (EU, 2002)

Melting point -111.9°C (SRC, 2013)

Boiling point 35°C (SRC, 2013)

Water solubility 590 000 mg/l at 25°C (SRC, 2013)

Log Kow 0.03 (SRC, 2013)

Koc Log Koc: 0.37 (Environment Canada, 2008)

Vapour pressure 538 mm Hg at 25°C (SRC, 2013)

Henry’s Law constant 6.96 x10-5 atm.m³/mole at 25°C (SRC, 2013)

Dissociation constant No data

Density 0.83 (relative density) (EU, 2002); 831 kg/m³ (EU, 2002)

9.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with oxirane methyl is provided in Appendix B8, and only an overall summary is provided here (Table 9.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 9.2 Summary of key toxicological data for oxirane methyl

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 382-1140 mg/kg bw (rats)

Dermal Low acute toxicity LD50 range: 950 mg/kg bw (rabbits) to 7168 mg/kg bw (guinea pigs)

Inhalation Low acute toxicity LC50: 4124 mg/m³ (mice)

Irritation and - Studies examining irritation and sensitisation to experimental animals and Sensitisation humans have given mixed results; however it appears that oxirane methyl has some potential to be irritating to the skin, eyes and respiratory tract and to be a sensitiser.

Genotoxicity In vitro and The weight of evidence indicates that oxirane methyl is genotoxic in vitro. in vivo When assessing the results of in vivo genotoxicity testing, the EU risk assessment concludes that oxirane methyl is a somatic cell mutagen, and that the general toxicological profile for oxirane methyl suggests that perhaps it shows genotoxicity only at sites of initial contact.

Repeat Dose Oral No suitable oral toxicity studies were located upon which an oral Tolerable Toxicity and Daily Intake (TDI) can be derived. Repeat dose studies indicate that tumour Carcinogenicity formation occurs at the site of administration.

Dermal No data were located

Inhalation Chronic Study: Key study used for assessment: In a combined chronic toxicity and carcinogenicity study, male and female Wistar rats (100 per sex per group) were exposed to oxirane methyl at 0, 30, 100 and 300 ppm (approximately 0, 71, 240 and 710 mg/m3, respectively), in a vehicle of nitrogen, for 6 hours per day, 5 days per week for 124 weeks (males) or 123 weeks (females) by whole body inhalation. A LOEC of 30 ppm (approximately 71 mg/m³) was identified from this study based on changes in the nasal cavity.

Reproductive and Oral Reduced sperm motility and damage to primary spermatocytes was observed Developmental in male rats administered a single oral dose of 520 mg/kg bw. However, it Toxicity should be noted that this dose is within the LD50 range.

Dermal No data were located

Inhalation In a combined repeat dose, reproductive and developmental study, male and female Crj:CD(SD)IGS rats (10/sex/dose) were exposed to oxirane methyl at 0, 125, 250, 500 or 1000 ppm (approximately 0, 300, 600, 1190 or 2340 mg/m³, respectively) by whole body inhalation for 6 hours/day, 7 days/week. Parental NOAEL: 125 ppm (approximately 300 mg/m³) Developmental NOAEL: 500 ppm (approximately 1190 mg/m³)

9.2.1 Evaluations by Authoritative Bodies

The International Agency for Research on Cancer (IARC) has classified oxirane methyl as a group 2B carcinogen (possibly carcinogenic to humans), based on inadequate evidence in humans for the carcinogenicity of oxirane methyl and sufficient evidence in experimental animals for the carcinogenicity of oxirane methyl (IARC, 1994).

9.2.2 Summary

Oxirane methyl is of low acute oral, dermal and inhalation toxicity. Studies examining irritation and sensitisation to experimental animals and humans have given mixed results; however it appears that oxirane methyl has some potential to be irritating to the skin, eyes and respiratory tract and to be a sensitiser. There is evidence that oxirane methyl is genotoxic in vitro and also in vivo, but this appears to be only at the site of administration. There is evidence that it is a somatic cell mutagen and an alkylating agent, where formation of DNA adducts has been observed in both experimental animals and occupationally exposed workers. In the EU, under the CLP regulation, oxirane methyl has been given the harmonised classification of H340 (may cause genetic defects). The most frequently observed effects in chronic studies are lesions of the stomach, in oral studies, and lesions of the nasal cavity, in inhalation studies. This suggests that oxirane methyl has a carcinogenic effect only in the tissues with maximum exposure after administration. It has been classified as a group 2B carcinogen (possibly carcinogenic to humans) by IARC. There is limited evidence of developmental variations following exposure to oxirane methyl by inhalation, however these effects are observed at doses that also lead to maternal toxicity. There is also limited evidence of neurotoxic effects on exposure to high concentrations of oxirane methyl.

9.3 Derivation of Tolerable Daily Intakes

9.3.1 Oral

No suitable oral toxicity studies were located upon which an oral Tolerable Daily Intake (TDI) can be derived. However, a number of inhalation studies are available and an oral TDI can be derived by extrapolation from these data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

Assuming an inhalation volume of 0.29 m³ per day and a bodyweight of 0.425 kg (425 g), this would equate to a systemic LOAEL of 8.7 mg/kg bw/day (rounded).

Following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to account for the use of a LOAEL) and multiplying by 100/50 (an arbitrary assumption within the IGHRC guidelines to account for the differences between oral and inhalation absorption), an oral Tolerable Daily Intake (TDI) of 0.017 mg/kg bw/day (17 µg/kg bw/day; rounded) can be derived.

It should be noted that this approach is likely to be conservative, as absorption via inhalation is likely to be at least as efficient, but may conceivably be more efficient, that oral exposure.

9.3.2 Dermal

No dermal toxicity studies were located upon which a dermal TDI can be derived, and it is not appropriate to extrapolate a dermal TDI from inhalation toxicity data.

9.3.3 Inhalation

In a 2-year study, rats were exposed to concentrations of oxirane methyl of at 0, 30, 100 or 300 ppm (approximately 0, 71, 240 and 710 mg/m3, respectively) for 6 hours/day, 5 days/week. A LOAEC of 30 ppm (approximately 71 mg/m³) was identified based on increased incidences of non-neoplastic changes. Adjusting to a continuous exposure, this LOAEC would be equivalent to a concentration of 12.7 mg/m³ (rounded) Following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEC), a Tolerable Daily Concentration of 0.071 mg/m³ (71 µg/m³) is derived.

9.4 Review of Current and Historical Usage

9.4.1 Uses

Oxirane methyl is used in three main areas:

 use as a monomer in polymer production;

 use as and intermediate in the synthesis of other chemicals; and

 use in direct applications (EU, 2002, Health Canada, 2008).

The main use of oxirane methyl in polymer production is to make polyether polyols which are mostly used in the manufacture of polyurethane foams (EU, 2002, Health Canada, 2008).

The main use of oxirane methyl as an intermediate is in the production of propylene glycol. Propylene glycol has many commercial uses from the manufacture of polyesters and polyester resins, plasticisers, emulsifiers in food, to solvents in paints, inks dyes, coatings, resins, cleaners, waxes, pharmaceuticals and cosmetics, and is a key ingredient in engine

coolants and aircraft de-icers. Oxirane methyl can also be used in the production of other monomers that are then used in polymer production such as acrylic resins (EU, 2002, Health Canada, 2008).

Oxirane methyl is also used directly as a stabiliser for dichloromethane and other chlorinated hydrocarbons. Oxirane methyl is used in the fumigation of dried fruit, cocoa, spices, processed nutmeats, starch and gums in the USA (EU, 2002; FAO/WHO, 2011; OECD, 2001; Health Canada, 2008). It is approved for use as an etherifying agent in the production of modified food starch by the United States Food and Drug Administration (EU, 2002). Oxirane methyl acts as a stabiliser of dichloromethane used for degreasing. Oxirane methyl is used at a concentration of 0.5% in this product as reported by suppliers within the European Union (EU).

Table 9.3 Usage of oxirane methyl by product type (EU, 2002)

Product EU EU tonnes/year USA Canada

Polyether polyols 72% 1 076 400 64% 65% Propylene glycol 23% 343 850 21% 29% Other uses 5% 74 750 15% 6% Total - 1 495 000 - -

9.4.2 Production

Global production of oxirane methyl was estimated to be 3 585 000 tonnes/year in 1990 (OECD, 2001). In the EU, production ranged between 580 000-2 750 000 tonnes/year, with seven companies manufacturing oxirane methyl. Imports into the EU were estimated at 10 000-50 000 tonnes/year (EU, 2002; OECD, 2001). In 1990 the EU capacity for oxirane methyl was 1 410 000 tonnes this was expected to rise by 300 000 tonnes by 1994. A total EU production figure of 1 495 000 tonnes/year was estimated in 2002 but it also noted that the real figure in 2000 could have been 400 000 tonnes greater than estimated (EU, 2002; OECD, 2001). Table 9.4 contains a summary of historical production data for oxirane methyl.

Table 9.4 Summary of historical global oxirane methyl production data (OECD, 2001 and EU, 2002)

Tonnes/year Year Global Western Europe USA Japan

1974 - - 80 000 130 000 1979 - 850 000 1 020 000 - 1982/83 - 810 000 720 000 190 000 1990 3 585 000 - - - 2002 - 1 495 000 - -

The UK office for National statistics reports import and export tonnages for the UK as presented in Table 9.5.

Table 9.5 Import and export tonnages of oxirane methyl in the UK (ONS, 2013)

Tonnes/year Year Intra EU Exports Intra EU Imports Extra EU Exports Extra EU Imports UK NET supply

2008 0.20 29444 0.01 43.2 29487 2009 0.10 19544 21.4 0.02 19522 2010 83.3 10164 135 25.3 9970 2011 7.88 12564 6.21 27.2 12577

In the USA in 1974, 800 000 tonnes of oxirane methyl was produced and 1 020 000 tonnes were produced in 1984.

9.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with oxirane methyl is provided in Appendix C8, and only an overall summary is provided here.

Data for oxirane methyl are available for drinking water, fresh surface water and effluents. Concentrations of oxirane methyl in drinking water are detected but not quantifiable, in fresh surface water not detected and in effluents have a maximum of 47 µg/l. No data are available for groundwater or marine surface water.

9.6 Review of Literature Data on Removal during Sewage Treatment

There are studies available reporting removal of oxirane methyl during sewage treatment indicating that it is readily biodegradable. There are no toxicity studies available to sewage treatment organisms. The available data are summarised in Table 9.6.

Table 9.6 Summary of studies reporting oxirane methyl removal during sewage treatment

Concentration Method/type of study Inoculum Results Reference (mg/l) Aerobic

Japanese MITI test, 28 Mixed activated 93-98% degraded (BOD) MITI (1988) cited in 100 days sludge (readily biodegradable) EU (2002) 12-14% after 15 days with Miller and sewage treatment minimal further degradation Closed bottle test 3 Watkinson (1985) effluent after 28 days (not readily cited in EU (2002) biodegradable) Manometric respiration test (Sapromat test), BASF AG (1977) Activated sludge 80 90 – 100% over 14 days measuring oxygen cited in EU (2002) consumption mixed inoculum of 14% after 5 days, 37% after Waggy and Payne Screening programme, municipal and 2.3-5.5 10 days, 65% after 15 days (1974) cited in EU non-adapted industrial sewage and 67% after 20 days. (2002) Standard dilution Inoculum taken from method (APHA biological sanitary 8% for non-adapted and 9% Bridié et al. (1979) NR Standard Method No waste treatment for adapted seed cited in EU (2002) 219) plant 15.8% after 5 days, 55.7% DOW (1978) cited BOD test Industrial sewage NR after 11 days and 74.1% in EU (2002) after 20 days Fill and Draw method, Domestic sewage Hatfield (1957) cited acclimated over 176 20% after 8 hours sludge in EU (2002) 1 month Grown on methane, No significant then incubated with Methylococcus Hou et al (1979) 24 disappearance of oxirane propylene oxide at capsulatus cited in EU (2002) methyl observed. 560 mg/l

Propylene oxide was De Bont et al. NR Nocardia A60 NR converted into 1,2- (1982) cited in EU propanediol (2002)

APHA: American Public Health Association

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound

MITI: Ministry of International Trade and Industry, Japan

NR: Not reported

9.7 Review of Literature Data on Removal during Drinking Water Treatment

No data were located on the removal of oxirane methyl during drinking water treatment, however, some predictions on the fate of this chemical in drinking water treatment can be made based on its physico-chemical properties.

A log Koc of 0.37 (Environment Canada, 2008) has been reported for oxirane methyl, which would suggest that it has high mobility in the water column and therefore is unlikely to be amenable to removal by GAC.

Oxirane methyl is volatile; a vapour pressure of 538 mm Hg and a Henry‘s Law constant of 6.96 x10-5 atm.m³/mole have been reported at 25°C (SRC, 2013). This may indicate that oxirane methyl will be amenable to moderate removal from water by air stripping.

Oxirane methyl is reported to undergo hydrolysis to propylene glycol, with half-lives of half-life of 10.7–21.7 days (Environment Canada, 2008). This process is reported to be accelerated in the presence of chloride ions (Environment Canada, 2008), which may indicate that chlorination of drinking water may also accelerate the hydrolysis of oxirane methyl.

10. 1,2,3-Trichloropropane

10.1 Physico-Chemical Properties

Data on the physico-chemical properties of 1,2,3-trichloropropane are provided in Table 10.1.

Table 10.1 Physico-chemical properties of 1,2,3-trichloropropane

CAS Number 96-18-4

Chemical Formula C3H5Cl3

Structure

Molecular weight 147.432 (ChemID, 2013)

Physical state Clear, colourless liquid (WHO, 2003)

Melting point -14.7°C (SRC, 2013)

Boiling point 157°C (SRC, 2013)

Water solubility 1750 mg/l at 25°C (SRC, 2013)

Log Kow 2.27 (SRC, 2013)

Koc 77-95 (WHO, 2003)

Vapour pressure 3.69 mm Hg at 25°C (SRC, 2013)

Henry’s Law constant 0.000343 atm.m³/mole at 25°C (SRC, 2013)

Dissociation constant No data

Density 1.38 g/cm³ at 20°C (WHO, 2003)

10.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with 1,2,3-trichloropropane is provided in Appendix B9, and only an overall summary is provided here (Table 10.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 10.2 Summary of key toxicological data for 1,2,3-trichloropropane

Route or Endpoint Summary study type

Acute toxicity Oral Low to moderate acute toxicity LD50 range: 120-500 mg/kg bw (rats)

Dermal Low acute toxicity LD50 range: 250-2457 mg/kg bw (rabbits)

Inhalation Low acute toxicity LC50 range: 4213-25 690 mg/m³ (rats)

Irritation and - Mildly irritating to skin and moderately irritating to eyes, but is not a skin Sensitisation sensitiser

Genotoxicity In vitro and Overall, the data indicate that 1,2,3-trichloropropane requires metabolic in vivo activation to produce genotoxic effects in vitro. However, while in vivo studies indicate that 1,2,3-trichloropropane may bind to DNA following intraperitoneal administration, oral administration did not alter DNA synthesis or induce lethal mutations.

Repeat Dose Oral Chronic Study: Key study used for assessment: Toxicity and In a 2-year study, F344/N rats (60/sex/dose) were administered Carcinogenicity 1,2,3-trichloropropane via oral gavage at doses of 0, 3, 10 or 30 mg/kg bw/day 5 days/week (approximately 0, 2.1, 7.1 and 21.4 mg/kg bw/day, respectively, adjusting to a 7 day/week dosing regimen) in corn oil. A LOAEL of 3 mg/kg bw/day (approximately 2.1 mg/kg bw/day, adjusting to a 7 day/week dosing regimen) was identified, based on the occurrence of neoplasms at all doses.

Dermal No data were located

Inhalation Sub-chronic Study: Key study used for assessment: In a 13-week study, Sprague Dawley rats (51 days old) were exposed to 1,2,3-trichloropropoane at concentrations of 0, 0.5 or 1.5 ppm (reported to be 0, 3.1 and 9.2 mg/m³, respectively) for 6 hours/day, 5 days/week. No toxicologically relevant effects were noted at any dose, therefore a NOAEC of 1.5 ppm (reported to be 9.2 mg/m³) was identified

Reproductive and Oral A four-stage reproductive and fertility assessment has been conducted in Developmental Swiss CD-1 mice. In stage 1, mice were administered doses of 0, 12.5, 25, 50, Toxicity 100 or 200 mg/kg bw/day for 14 days. In stage 2, mice were administered 0, 30, 60 or 120 mg/kg bw/day for 1 week prior to co-habitation, 14-weeks cohabitation and 3 weeks after mating. Stage 3 was a 24-week cross-over mating trial. Stage 4 was a 30 week study to examine the reproductive performance in the offspring of task 2 to produce a 2nd generation study. In stage 1, one male at the top dose died before the end of the study. In stage 2, fertility was decreased at the mid and top doses. In stage 3, bodyweight was decreased in both sexes. The number of live pups per litter was decreased and the proportion of males/litter was slightly reduced. In stage 4, decreased relative testes weight in males and increased relative ovary weight

Route or Endpoint Summary study type in females were noted.

Dermal No data were located

Inhalation A 1-generation study has been reported in rats, however, due to study deficiencies, it is not possible to identified NOAELs.

10.2.1 Summary

1,2,3-Trichloropropane is of low to moderate acute oral toxicity, of low acute inhalation toxicity, and of low acute dermal toxicity to experimental animals. It is mildly irritating to skin and moderately irritating to eyes, but is not a skin sensitiser. The results from several in vitro genotoxicity assays indicate that 1,2,3-trichloropropane requires a metabolic activation system to induce genotoxic effects. In vivo data have shown no evidence of genotoxic effects via the oral route; the only positive effects in vivo have been reported following intraperitoneal administration. However, oral gavage carcinogenicity studies have reported neoplastic effects in both rats and mice in a range of tissues including the oral mucosa, forestomach of both species, the liver in mice, and the liver, pancreas and mammary glands of rats. Studies on the reproductive toxicity of 1,2,3-trichloropropane are limited, however, an oral study has indicated decreased fertility in mice at doses of 60 mg/kg bw/day or above.

10.3 Derivation of Tolerable Daily Intakes

10.3.1 Oral

In a 2-year study, rats were administered 1,2,3-trichloropropane via oral gavage at doses of 0, 3, 10 or 30 mg/kg bw/day 5 days/week (approximately 0, 2.1, 7.1 and 21.4 mg/kg bw/day, respectively, adjusting to a 7 day/week dosing regimen). A LOAEL of 3 mg/kg bw/day (approximately 2.1 mg/kg bw/day, adjusting to a 7 day/week dosing regimen) was identified based on the occurrence of neoplasms in the forestomach and nephropathy in males at all doses. Using this LOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEL), an oral Tolerable Daily Intake (TDI) of 0.002 mg/kg bw/day (2 µg/kg bw/day; rounded) is derived.

10.3.2 Dermal

No dermal toxicity studies were located upon which a dermal TDI can be derived. However, as there is no evidence to suggest that dermal exposure will significantly impair the integrity of skin (i.e. 1,2,3-trichloropropane is not corrosive), a dermal TDI can be derived by extrapolation from oral toxicity data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

Assuming equivalent bioavailability for the oral and dermal routes, a dermal TDI of 0.002 mg/kg bw/day (2 µg/kg bw/day; rounded) can be derived.

However, it should be noted that this is likely to be a highly conservative value, as in reality, very few substances are absorbed as readily via the dermal route as they are via the oral route.

10.3.3 Inhalation

In a 13-week study, rats (51 days old) were exposed to 1,2,3-trichloropropane at concentrations of 0, 0.5 or 1.5 ppm (reported to be 0, 3.1 and 9.2 mg/m³, respectively) for 6 hours/day, 5 days/week. No toxicologically relevant effects were noted at any dose, therefore a NOAEC of 1.5 ppm (reported to be 9.2 mg/m³) was identified. Adjusting to a continuous exposure, this LOAEC would be equivalent to a concentration of 1.7 mg/m³ (rounded) Following application of an uncertainty factor of 100 (to account for inter- and intra- species variation), a Tolerable Daily Concentration of 0.017 mg/m³ (1.7 µg/m³) is derived.

10.4 Review of Current and Historical Usage

There are no known natural sources of 1,2,3-tripchloropropane. 1,2,3-Trichloropropane is mostly used as an intermediate in the synthesis of other compounds such as pesticides for soil fumigation, is used as a crosslinking agent in the production of polymers such as polysulphides and hexafluoropropylene. 1,2,3-Trichloropropane has previously been produced in vast quantities as a by-product from the production of other chlorinated compounds such as epichlorohydrin, although the majority of 1,2,3-trichloropropane (>80%) is incinerated on-site (WHO, 2003; OECD, 2004).

Historically outside of the USA it is reported that 1,2,3-trichloropropane was used as a solvent for hydrophobic compounds and resins, as a paint and varnish remover and a degreasing agent but there is no indication that it is sold in consumer products currently (WHO, 2003; OECD, 2004; ATSDR, 1992).

The US EPA (2009) reported that 1,2,3-trichloropropane was used in the chemical industry as a solvent of oils, fats, waxes and resins and is found in consumer products such as paint thinner and varnish remover. In the USA it is also used in the production of polymers such as polysulphide rubbers and some pesticides (US EPA, 2009). Pesticides and nematicides such as Telone can contain residual 1,2,3-trichloroproane of up to 0.17% and has been identified as a potential source of 1,2,3-trichloropropane into the environment (WHO, 2003).

An estimated 9000 - 14 000 tonnes/year of 1,2,3-tripchloropropane is manufactured in the United States (US) and 50 000 tonnes/year is produced as a by-product of epichlorohydrin. There are around 20-30 epichlorohydrin manufacturing facilities in North America, Europe and Asia (WHO, 2003; OECD, 2004) It is reported that in 1977 the production volume for 1,2,3-trichloropropane was approximately 9 525-49 895 tonnes/year (ATSDR, 1992).

In 2011, according to the European Chemicals Agency (ECHA), 1,2,3-trichloropropane is only manufactured at five sites within the European Union under strictly controlled conditions, with no known manufacturing occurring within the UK (ECHA, 2011).

10.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with 1,2,3-trichloropropane is provided in Appendix C9, and only an overall summary is provided here.

Data for 1,2,3-trichloropropane are available for drinking water, groundwater, fresh and marine surface water and effluents. Concentrations of 1,2,3-trichloropropane in drinking water range from <0.24 to 2.654 µg/l, in ground water from 0.02 to >100 µg/l, in fresh surface water from 0.027 to 100 µg/l, in marine surface water from <0.1 to <0.5 µg/l and in effluents from 0 to <200 µg/l.

10.6 Review of Literature Data on Removal during Sewage Treatment

There are no studies available reporting removal of 1,2,3-trichloropropane during sewage treatment and but there are data for the effects of 1,2,3-trichloropropane on sewage treatment organisms indicating low to high toxicity, depending on species of organism. These have been detailed in Table 10.3.

Table 10.3 Summary of studies reporting toxicity of 1,2,3-trichloropropane to sewage treatment organisms

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

IC50 (ammonia Konnecker and Nitrosomononas sp. 24 30 consumption) Schmidt (2003) Mixed culture of aerobic Konnecker and IC50 (oxygen uptake) 15 290 heterotrophs Schmidt (2003) Konnecker and IC50 (gas production) Methanogens 48 0.63 Schmidt (2003)

Photobacterium Konnecker and IC50 (bioluminescence) 5 min 19 phosphoreum Schmidt (2003)

IC50: Inhibition Concentration for 50%. A point estimate of the toxicant concentration that would cause 50% reduction in a non-lethal biological measurement of the test organisms.

10.7 Review of Literature Data on Removal during Drinking Water Treatment

Tratnyek et al. (2008) have reported that 1,2,3-trichloropropane has low to moderate absorption capacity for granular activated carbon (GAC).

In a bench-scale experiment using an advanced oxidation process (AOP), a raw water sample containing 1,2,3-trichloropropane at a concentration of 0.95 µg/l was added to a reactor feed tank with hydrogen peroxide solution and ozone (ozone dose of 53 mg/l). The hydrogen peroxide-to-ozone mole ratio was 0.7. The concentration of 1,2,3-trichloropropane was reported to be reduced to <0.005 µg/l (Dombeck and Borg, 2005). 1,2,3-Trichloropropane is expected to undergo limited removal by air stripping.

11. Urethane

11.1 Physico-Chemical Properties

Data on the physico-chemical properties of urethane are provided in Table 11.1.

Table 11.1 Physico-chemical properties of urethane

CAS Number 51-79-6

Chemical Formula C3H7NO2

Structure

Molecular weight 89.0933 (ChemID, 2013)

Physical state Colourless crystals or white granular powder (HSDB, 2013)

Melting point 49°C (SRC, 2013)

Boiling point 185°C (SRC, 2013)

Water solubility 480 000 mg/l at 15°C (SRC, 2013)

Log Kow -0.15 (SRC, 2013)

Koc 20 (HSDB, 2013)

Vapour pressure 0.262 mm Hg at 25°C (SRC, 2013)

Henry’s Law constant 6.43 x10-8 atm.m³/mole at 25°C (SRC, 2013)

Dissociation constant No data

Density 0.9813 (relative density) (HSDB, 2013)

11.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with urethane is provided in Appendix B10, and only an overall summary is provided here (Table 11.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 11.2 Summary of key toxicological data for urethane

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 1809 mg/kg bw (rats) to 2500 mg/kg bw (mice)

Dermal No data were located

Inhalation No data were located

Irritation and - No data were located Sensitisation

Genotoxicity In vitro and Mixed results were observed in in vitro genotoxicity tests, indicating that in vivo urethane is not genotoxic in bacteria or yeast, however evidence of genotoxicity has been observed in mammalian assays. Clear evidence of genotoxicity can be observed in in vivo tests.

Repeat Dose Oral Chronic Study: Key study used for assessment: Toxicity and B6C3F1 mice were administered urethane in drinking water at doses of 0, Carcinogenicity 0.6, 3, 6, 60 or 600 mg/kg diet (approximately 0, 0.07, 0.36, 0.72, 7.2 and 72 mg/kg bw/day, respectively). A NOAEL of 3 mg/kg diet (approximately 0.07 mg/kg bw/day) was identified based on dose-related increases in liver angiosarcomas.

Dermal Study has been reported, however, due to study deficiencies, it is not possible to identified NOAELs.

Inhalation Sub-acute Study: Key study used for assessment: Female JCL:ICR mice (group size not reported) were exposed to urethane via inhalation at doses of 0.25 µg/ml and 1.29 µg/ml for 1,3,5 or 10 days and 0.25, 1, 2, 4 or 5 days, respectively. Male JCL:ICR mice were exposed to doses of 0.25 µg/ml for 10 days (50 mice), or 1.29 µg/ml for 4 days (47 mice) A LOAEC of 0.25 µg/ml was identified.

Reproductive and Oral Reproductive Study: Developmental Fischer 344 rats and B6C3F1 mice (10/sex/dose) were administered Toxicity urethane in drinking water at concentrations of 0, 110, 330, 1100, 3300 or 10 000 mg/l for 13 weeks. Rat NOAEL (epididymal spermatozoal motility and concentration in males): 330 mg/l Mouse LOAEL (epididymal spermatozoal concentration in males): 110 mg/l

Dermal No data were located

Inhalation No data were located

11.2.1 Evaluations by Authoritative Bodies

In 2010, the International Agency for Research on Cancer (IARC) evaluated the available data and classified aniline in Group 2A, (i.e. probably carcinogenic to humans) based on

―sufficient‖ and ―inadequate‖ evidence for carcinogenicity in experimental animals and humans, respectively (IARC, 2010).

11.2.2 Toxicity Summary

Urethane is of low acute oral toxicity in experimental animals. No data were located on the acute dermal and inhalation toxicity of urethane. No data were located on the irritation and sensitisation of urethane via oral, dermal or inhalation routes. Mixed results were observed in in vitro genotoxicity tests, indicating that urethane is not genotoxic in bacteria or yeast, however evidence of genotoxicity has been observed in mammalian assays. Clear evidence of genotoxicity can be observed in in vivo tests. Urethane has been classified as IARC Group 2A (i.e. probably carcinogenic to humans). There is clear evidence for carcinogenicity in experimental animals, with high incidences of lung and liver adenomas and carcinomas observed. Evidence of the developmental and reproductive toxicity of urethane has been reported, with increased skeletal malformations of foetuses observed following administration to pregnant females, and decreased epididymal spermatozoal motility and concentration have been observed in males following treatment with urethane.

11.3 Derivation of Tolerable Daily Intakes

11.3.1 Oral

In a 70-week study, B6C3F1 mice were administered urethane in drinking water at doses of 0, 0.6, 3, 6, 60 or 600 mg/kg diet (approximately 0, 0.07, 0.36, 0.72, 7.2 and 72 mg/kg bw/day, respectively). A NOAEL of 3 mg/kg diet (approximately 0.07 mg/kg bw/day) was identified based on dose-related increases in liver angiosarcomas. Using this NOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for potential carcinogenicity), an oral Tolerable Daily Intake (TDI) of 0.00007 mg/kg bw/day (0.07 µg/kg bw/day) is derived.

Urethane is classified as Group 2A by the International Agency for Research on Cancer (IARC) (i.e. probably carcinogenic to humans) and there is clear evidence of genotoxicity in vivo. Therefore, although a TDI has been proposed, it may be appropriate to ensure the concentration of urethane in water is as low as reasonably practicable.

11.3.2 Dermal

Although several dermal toxicity studies were located, they are not considered appropriate for use in the derivation of a dermal TDI.

No data were located to suggest urethane is irritating to the skin. Therefore, a dermal TDI can be derived by extrapolation from oral toxicity data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

Assuming equivalent bioavailability for the oral and dermal routes, a dermal TDI of 0.00007 mg/kg bw/day (0.07 µg/kg bw/day) can be derived.

However, it should be noted that this is likely to be a highly conservative value, as in reality, very few substances are absorbed as readily via the dermal route as they are via the oral route.

11.3.3 Inhalation

A single inhalation study was located for urethane in which a LOAEC of 0.25 µg/ml (250 mg/m³) was identified in mice. Using this LOAEC and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEC, the limited inhalation database and for potential carcinogenicity), a Tolerable Daily Concentration of 0.25 mg/m³ is derived.

However, urethane is classified as Group 2A by the International Agency for Research on Cancer (IARC) (i.e. probably carcinogenic to humans) and there is clear evidence of genotoxicity in vivo. Therefore, although a Tolerable Daily Concentration has been proposed, it may be appropriate to ensure the concentration of urethane in water is as low as reasonably practicable.

11.4 Review of Current and Historical Usage

Urethane is a naturally occurring compound that is formed during the fermentation of foods such as fruit, beer, wine, yoghurt, bread and soy sauce (FAO/WHO, 2005; FSA, 2001). The UK Food Standards Agency‘s Food Advisory Committee discussed the presence of naturally occurring urethane in whiskey, brandies, black tea, wine, beer, and bread and stated that levels of urethane in these foods should be reduced to the lowest technologically achievable concentrations. Historically industrial, medical and veterinary uses of urethane have been reported but there is little data on the uses of urethane at present (FAO/WHO, 2005).

In the US it is reported that urethane is widely used as an ingredient in paint (US EPA, 2013). It is also used as a solvent for organic materials, an intermediate in organic synthesis, in the preparation and modification of amino resins, and as a solubiliser and co-solvent for pesticides, fumigants and cosmetics (HSDB, 2013). No data on the relative amounts of urethane used in the US could be located.

11.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with urethane is provided in Appendix C10, and only an overall summary is provided here.

Data for urethane are available for drinking water, groundwater and effluents. Concentrations of urethane in drinking water and groundwater are detected but not quantifiable and in effluents range from 18 to 37 µg/l. No data are available for fresh and marine surface water.

11.6 Review of Literature Data on Removal during Sewage Treatment

There are no studies available reporting the removal of urethane during sewage treatment or its toxicity to sewage treatment organisms.

11.7 Review of Literature Data on Removal during Drinking Water Treatment

No data were located on the removal of urethane during drinking water treatment, however, some predictions on the fate of this chemical in drinking water treatment can be made based on its physico-chemical properties.

A Koc of 20 (HSDB, 2013) has been reported for urethane, which would suggest that it has high mobility in the water column and therefore is unlikely to be amenable to removal by GAC.

Urethane has a vapour pressure of 0.262 mm Hg at 25°C and a Henry‘s Law constant of 6.43 x10-8 atm.m³/mole at 25°C (SRC, 2013). Therefore, it is unlikely to undergo significant removal by air stripping.

12. Ethylene oxide

12.1 Physico-Chemical Properties

Data on the physico-chemical properties of ethylene oxide are provided in Table 12.1.

Table 12.1 Physico-chemical properties of ethylene oxide

CAS Number 75-21-8

Chemical Formula C2H4O

Structure

Molecular weight 44.0526 (ChemID, 2013)

Physical state Colourless gas (WHO, 2003)

Melting point -111.7°C (SRC, 2013)

Boiling point 10.7°C (WHO, 2003)

Water solubility Infinity soluble (WHO, 2003); 1 000 000 mg/l at 25°C (SRC, 2013)

Log Kow -0.22 (WHO, 2003); -0.3 (WHO, 2003)

Koc Log Koc: 1.204 (WHO, 2003)

Vapour pressure 66 kPa at 0°C (WHO, 2003); 100 kPa at 10°C (WHO, 2003); 146 kPa at 20°C (WHO, 2003); 208 kPa at 30°C (WHO, 2003)

Henry’s Law constant 14 Pa.m³/mol (WHO, 2003); 12.16 Pa.m³/mol (WHO, 2003); 19.86 Pa.m³/mol (WHO, 2003)

Dissociation constant No data

Density Specific gravity: 0.882 at 10°C (HSDB, 2013)

12.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with ethylene oxide is provided in Appendix B11, and only an overall summary is provided here (Table 12.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 12.2 Summary of key toxicological data for ethylene oxide

Route or Endpoint Summary study type

Acute toxicity Oral Low to moderate acute toxicity LD50 range: 72 mg/kg bw (rats) to 365 mg/kg bw (mice)

Dermal No data were located

Inhalation Low to moderate acute toxicity LC50 range 643-9030 mg/m3 (rats)

Irritation and - Irritating effects to the eyes, skin and respiratory system. Sensitisation Not a skin sensitiser.

Genotoxicity In vitro and Clear evidence of genotoxicity has been reported in both in vitro and in vivo in vivo tests, indicating that ethylene oxide is genotoxic. Evidence of genotoxicity has also been reported in a number of epidemiological studies in albeit limited numbers of workers exposed to ethylene oxide.

Repeat Dose Oral Chronic Study: Toxicity and Female Sprague-Dawley rats (50/dose), were administered ethylene oxide by Carcinogenicity oral gavage in salad oil at concentrations of 0, 7.5 or 30 mg/kg bw, twice weekly for 3 years. A LOAEL of 7.5 mg/kg bw can be identified for this study, based on dose- dependent increases in squamous cell carcinomas, hyperkeratosis and hyperplasia. This study was not considered sufficiently robust for use in the risk assessment.

Dermal A single study has been reported, however, it is not possible to identify a NOAEL. No evidence of skin tumours was reported in this study.

Inhalation Chronic Study: B6C3F1 mice (50/sex/dose) were administered ethylene oxide via inhalation at concentrations of 0, 50 or 100 ppm (approximately 0, 90.31 and 180.61 mg/m³, respectively) for 6 hours/day, 5 days/week, for 2 years. Dose- dependent increases in the frequency of alveolar and bronchiolar carcinomas (both sexes), alveolar and bronchiolar adenomas (females), combined incidences of benign and malignant lung tumours (both sexes) and Harderian gland papillary cystadenomas (both sexes) were observed.

Route or Endpoint Summary study type Chronic Study: Fischer 344 rats (120/sex/dose) were administered ethylene oxide via inhalation at concentrations of 0, 10, 33 or 100 ppm (approximately 0, 18, 60 and 181 mg/m³, respectively) for 6 hours/day, 5 days/week, for 2 years. An increase in the number of animals with malignant mononuclear cells in the peripheral blood was observed. Histologic evaluation of the spleens confirmed mononuclear cell leukaemia. Peritoneal mesotheliomas were observed in treated males, and a positive trend was reported in male and female rats in the incidence of primary brain tumours, including gliomas, malignant reticuloses, and granular cell tumours.

Reproductive and Oral No data were located Developmental Dermal No data were located Toxicity Inhalation Reproductive Study: Key study used for assessment: Fischer 344 rats (30/sex/dose) were exposed to ethylene oxide via inhalation at concentrations of 0, 10, 33 or 100 ppm (0, 18, 60 and 181 mg/m³, respectively). For 12 weeks the rats were exposed to ethylene oxide for 6 hours/day, 5 days/week. After 12 weeks, rats were exposed to ethylene oxide for 6 hours/day, 7 days/week during the cohabitation period. After 2 weeks of mating exposure to ethylene oxide was continued for the female rats for 6 hours/day, 7 days/week from day 0 until day 19 of gestation. Five days after parturition, the dams were separated from their pups and exposed to ethylene oxide for 6 hours/day, 7 days/week until day 21 postpartum. Reproductive NOAEC: 33 ppm (60 mg/m3) based on fewer implantation sites per female, a small ratio of foetuses born to the number of implants, a decreased number of pups per litter and longer lengths of gestation.

12.2.1 Evaluations by Authoritative Bodies

In 2012, the International Agency for Research on Cancer (IARC) evaluated ethylene oxide and upgraded the overall classification based on mechanistic and other relevant data to Group 1 (i.e. carcinogenic to humans), on the basis that there is sufficient evidence for the carcinogenicity of ethylene oxide in experimental animals, relying heavily on the compelling data in support of the genotoxic mechanism (IARC, 2012).

12.2.2 Toxicity Summary

Ethylene oxide is of low acute oral toxicity and low to moderate acute inhalation toxicity in experimental animals. The available data on the irritation and sensitisation of ethylene oxide in humans and experimental animals indicate the presence of irritating effects to the eyes, skin and respiratory system. Clear evidence of genotoxicity has been reported in both in vitro and in vivo tests, indicating that ethylene oxide is genotoxic, and evidence of genotoxicity has also been reported in a number of epidemiological studies in limited numbers of workers

exposed to ethylene oxide. In repeated dose studies ethylene oxide was observed to induce formation of mammary gland tumours via the oral route, and alveolar/bronchiolar carcinomas and mononuclear cell leukaemia via inhalation. In carcinogenicity studies in humans some evidence of carcinogenicity was observed following occupational exposure to ethylene oxide, particularly the incidence of leukaemia. There is limited evidence of the reproductive toxicity of ethylene oxide in experimental animals, and some evidence of spontaneous abortions in humans occupationally exposed to ethylene oxide.

12.3 Derivation of Tolerable Daily Intakes

12.3.1 Oral

No suitably robust oral toxicity studies were located upon which an oral Tolerable Daily Intake (TDI) can be derived. However, a number of inhalation studies are available and an oral TDI can be derived by extrapolation from these data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

In a reproductive study, rats were exposed to ethylene oxide via inhalation at concentrations of 0, 10, 33 or 100 ppm (0, 18, 60 and 181 mg/m³, respectively). A NOAEC of 33 ppm (60 mg/m³) was identified. Adjusting to a continuous exposure, this NOAEC would be equivalent to a concentration of 10.7 mg/m³ (rounded).

Assuming an inhalation volume of 0.29 m³ per day and a bodyweight of 0.425 kg (425 g), this would equate to a systemic LOAEL of 7.3 mg/kg bw/day (rounded).

Following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to account for the limited database) and multiplying by 100/50 (an arbitrary assumption within the IGHRC guidelines to account for the differences between oral and inhalation absorption), an oral Tolerable Daily Intake (TDI) of 0.015 mg/kg bw/day (15 µg/kg bw/day; rounded) can be derived.

It should be noted that this approach is likely to be conservative, as absorption via inhalation is likely to be at least as efficient, but may conceivably be more efficient, that oral exposure.

However, ethylene oxide is classified as a Group 1 carcinogen by the International Agency for Research on Cancer (IARC) Therefore, its concentration in water should be as low as reasonably practicable.

12.3.2 Dermal

A single dermal toxicity study was located for ethylene oxide, which did not find any evidence of carcinogenic effects following dermal exposure. However, this conclusion cannot be considered robust as no histopathological examinations were conducted. Therefore, this study is not considered appropriate for use in a risk assessment.

Consideration has been made to extrapolating a dermal TDI from oral toxicity data, however, the oral study is not considered sufficiently robust to allow adequate extrapolation. Considered was also made with regards to extrapolating from inhalation data, however, it is not appropriate to extrapolate a dermal TDI from inhalation toxicity data, as there are no guidance available on conducting such extrapolations.

However, it should be noted that ethylene oxide is classified as a Group 1 carcinogen by the International Agency for Research on Cancer (IARC) Therefore, its concentration in water should be as low as reasonably practicable.

12.3.3 Inhalation

In a reproductive study in in rats, a reproductive NOAEC 33 ppm (60 mg/m³) was identified. Using this LOAEC and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation, and 10 to account for the limited database), a Tolerable Daily Concentration of 0.06 mg/m³ (rounded) is derived.

12.4 Review of Current and Historical Usage

12.4.1 Production

In 1978, the estimated global production of ethylene oxide was 4 540 000 tonnes. Half of this figure was estimated to be from the USA. In Western Europe, 865 000 tonnes were produced in 1972 and 1 370 000 tonnes were produced in 1981. In Japan, 470 000 tonnes were produced in 1982.

In the USA, ethylene oxide is a major industrial chemical and one of the 25 highest production volume chemicals in the USA (ATSDR, 1990). It has also been identified as a global high production chemical by the Organisation for Economic Co-operation and Development (OECD) (OECD, 2013). Annual production in the USA has been reported to be 1 906 800 tonnes (in 1973), 2 610 500 tonnes (in 1979), 2 100 000 tonnes (in 1977) and 2 172 530 tonnes (in 1987) (NIOSH, 1977, ATSDR, 1990).

In Canada, domestic production of ethylene oxide was estimated to be 625 000 tonnes in 1996 and 682 000 tonnes in 1999. Canadian import values of 43 900 and 10 970 tonnes between 1992 and 1996 were reported (Environment Canada/Health Canada, 2001).

Data from the UK Office of National Statistics is presented in Table 12.3, which show the imports and exports of ethylene oxide into the UK and the NET UK supply.

Table 12.3 Annual data for imports and exports into the UK

Tonnes/year Year Intra EU Intra EU Extra EU Extra EU UK NET Exports Imports Exports Imports supply

2008 2276 19.8 205 10.4 -2451 2009 495 4264 207 6.79 3568 2010 83.3 10080 135 25.3 9887 2011 7.88 12564 6.21 27.2 12577

12.4.2 Uses

The main use of ethylene oxide is reported to be as an intermediate for other chemical products, especially ethylene glycol (Environment Canada/Health Canada, 2001, HSDB, 2013). In Canada, it is reported that 89% of the ethylene oxide produced was used in the production of ethylene glycol in 1993 and in 1996 this increased to 95% used in ethylene glycol production and 4% used in the manufacture of surfactants (Environment Canada/Health Canada, 2001, WHO, 2003). The Concise International Chemical Assessment Document on ethylene oxide published by the World Health Organisation assumed that these usage percentages would be similar in most countries but the gross quantities would vary. In the USA, 99% of the ethylene oxide is used as an intermediate for chemical production the remainder is used in the gaseous form as a sterilising agent, fumigant, or insecticide (HSDB, 2013), 60% is used for ethylene glycol production. Ethylene oxide is also used in the manufacture of choline chloride, glycol ethers, polyglycols, rocket propellant, petroleum demulsifiers, acrylonitrile, and ethanolamines (Environment Canada/Health Canada, 2001, HSDB, 2013).

Ethylene oxide has routinely been used as a gas in medical and health care related facilities for sterilising heat-sensitive instruments and equipment that come into contact with biological tissue (NISOH 1977, ARTG, 2013, HSDB, 2013). Ethylene oxide gas has also been used historically to sterilise inexpensive single-use items such as syringes on an industrial scale. Ethylene oxide is also reported to be used to sterilise electronic cardiac ―pacemakers‖, blood oxygenators and dialysers in the USA (NIOSH, 1977, HSDB, 2013). In 1977, it is reported that 80% of all such items were sterilised using ethylene oxide in the USA (NIOSH, 1977).

Library and museum artefacts are treated with ethylene oxide to control fungi and insects that might cause damage. Ethylene oxide has also been used to sterilise food and food packaging however in 2006 it‘s classification as a food additive was withdrawn by the Joint Food and Agriculture Organization/World Health Organization Expert Committee on Food Additives (JECFA) (JECFA, 2013). In Canada, ethylene oxide is used as an active ingredient in the control of insects and bacteria in spices, natural seasonings, textiles and cosmetics. Ethylene oxide is also used as a formulant or component in 25 pesticides in Canada, used in fungicides, insecticides and herbicides. Concentration ranges from trace amounts to 0.423%

(Environment Canada/Health Canada, 2001). Ethylene oxide is currently not approved for use as an active ingredient in pesticides within the European Union (EU, 2013). In Australia, ethylene oxide is the register active ingredient in 21 pesticide products, mostly as fumigants or surfactants (Australia Pesticides and Veterinary Medicines Authority, 2013).

There are natural sources of ethylene oxide; within some plants ethylene (a natural growth regulator) degrades to ethylene oxide and it is also a product of ethylene catabolism in some microorganisms. Ethylene oxide can be generated from waterlogged soil, manure, and sewage sludge (Environment Canada/Health Canada, 2001, WHO, 2003). Ethylene oxide is also a by-product of fossil fuel combustion and tobacco smoke (WHO, 2003).

12.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with ethylene oxide is provided in Appendix C11, and only an overall summary is provided here.

Data for ethylene oxide are available for groundwater, freshwater and effluents. No data were located for drinking water or marine water. Concentrations of ethylene oxide in groundwater are reported as 28 µg/l, in fresh surface water as 21 µg/l, and in effluents as 2000 µg/l.

12.6 Review of Literature Data on Removal during Sewage Treatment

There are studies available reporting removal of ethylene oxide during sewage and data available for the toxicity of ethylene oxide to sewage treatment organisms indicate that it is readily biodegradable by a range of organisms. The available data are summarised in Table 12.4 and Table 12.5.

Table 12.4 Summary of studies reporting ethylene oxide removal during sewage treatment

Method/type of Concentration Inoculum Results Reference study (mg/l) Aerobic

69% biodegraded after 20 days Activated Respirometry NR (19% after 5 days and 61% after IUCLID (2000) sludge 11 days) Activated Respirometry NR 30-50% after 5 days IUCLID (2000) sludge Culture jar designed Activated for volatile 107% degradation: BOD sludge 100 NEDO (2004) substances measured (30 mg/l) measuring BOD

Method/type of Concentration Inoculum Results Reference study (mg/l)

Culture jar designed Activated for volatile 96% degradation: TOC sludge 100 NEDO (2004) substances measured (30 mg/l) measuring TOC Culture jar designed Activated for volatile 100% degradation : GC sludge 100 NEDO (2004) substances measured (30 mg/l) measuring BOD 3% degradation after 5 days and BOD NR NR NEDO (2004) 52% after 20 days Anaerobic

NR NR NR Biodegradable NEDO (2004)

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound

GC: Gas Chromatography

NR: Not reported

TOC: Total Organic Carbon. The total amount of organic carbon in an aqueous solution/suspension

Table 12.5 Summary of studies reporting toxicity of ethylene oxide to sewage treatment organisms

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

Inoculum acclimated to ethene Van Agteren et al converted ethylene oxide into acetyl- Mycobacterium NR NR (1998) cited in coenzyme A by an uncharacterized E20 and M. E44 HSDB (2013) enzyme complex

NR: Not Reported.

12.7 Review of Literature Data on Removal during Drinking Water Treatment

No data on the removal of ethylene oxide via drinking water treatment were located. However, some predictions on the fate of this chemical in drinking water treatment can be made based on its physico-chemical properties and its fate in the environment.

Ethylene oxide is not expected to sorb to activated carbon, based on a log Koc of 1.204 (WHO, 2003). However, volatilisation is expected to be an important fate process; half-lives in the environment are reported to be 1 hour with no wind, and 0.5 hours at a wind speed of

5 m/second (Health Canada, 1999; WHO, 2003). Therefore, it is likely that significant removal or ethylene oxide will occur by air stripping.

Ethylene oxide will also undergo rapid hydrolysis, although this process is relatively slow compared to volatilisation. Half-lives of 12-14 days have been reported in freshwater at pH 5- 7, and 9-11 days have been reported in saltwater (Health Canada, 1999). In freshwater, ethylene oxide is hydrolysed to ethylene glycol, while in saltwater, it is hydrolysed to ethylene glycol and ethylene chlorohydrin (Health Canada, 1999).

13. Formaldehyde

13.1 Physico-Chemical Properties

Data on the physico-chemical properties of formaldehyde are provided in Table 13.1.

Table 13.1 Physico-chemical properties of formaldehyde

CAS Number 50-00-0

Chemical Formula CH2O

Structure

Molecular weight 30.0258 (ChemID, 2013)

Physical state Colourless gas (OECD, 2002)

Melting point -92°C (OECD, 2002)

Boiling point -19.2°C (OECD, 2002)

Water solubility 400 000-550 000 mg/l (Health Canada, 1999)

Log Kow 0.35 (OECD, 2002)

Koc Log Koc: 0.70-1.57 (Health Canada, 1999)

Vapour pressure 5176 hPa (OECD, 2002); 5185 hPa (OECD, 2002)

2.2 ×10-2 to 3.4 ×10-2 Pa.m³/mol at 25°C (Health Henry’s Law constant Canada, 1999); 3.37 x10-7 atm.m³/mol at 25°C (SRC, 2013)

Dissociation constant pKa: 13.27 at 25°C (HSDB, 2013)

Density 0.8153 g/cm³ at -20°C (OECD, 2002)

13.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with formaldehyde is provided in Appendix B12, and only an overall summary is provided here (Table 13.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 13.2 Summary of key toxicological data for formaldehyde

Route or Endpoint Summary study type

Acute toxicity Oral Low to moderate acute toxicity LD50 range: 42 mg/kg bw (mice) to 800 mg/kg bw (rats)

Dermal Moderate acute toxicity LD50: 270 mg/kg bw (rabbits)

Inhalation Moderate to high acute toxicity LC50 range: 497 mg/m³ (mouse, 4-hour) to 984 mg/m³ (rat, 30-minute)

Irritation and - Skin, eye and respiratory irritant in both experimental animals and Sensitisation humans, and it is also a skin sensitiser

Genotoxicity In vitro and Overall, the data indicate that formaldehyde is genotoxic in vitro. The data in vivo from in vivo studies indicate that formaldehyde is genotoxic, but only in the tissues of initial contact.

Repeat Dose Toxicity Oral Chronic Study: Key study used for assessment: and Carcinogenicity In a 2-year study, Wistar rats (70/sex/dose) were administered formaldehyde orally via drinking water at doses of 0, 20, 260 or 1900 mg/l (reported to be 0, 1.2, 15 and 82 in males, and 0, 1.8, 21 and 109 mg/kg bw/day in females, respectively). A NOAEL of 260 mg/l (reported to be 15 mg/kg bw/day) was identified based on histopathological changes in the gastrointestinal system due to irritancy.

Dermal Two initiation-promotion studies did not result in increased incidence of skin tumours. However, a third study reported that formaldehyde may have weak promoting potential.

Inhalation Chronic Study: Key study used for assessment: Male Fischer 344 rats (32/group) were exposed to formaldehyde in aqueous methanol at 0, 0.3, 2.0 or 15 ppm (reported to be 0, 0.36, 2.4 and 17 mg/m3, respectively) for 6 hours/day, 5 days/week for up to 28 months A LOAEC of 0.3 ppm (reported to be 0.36 mg/m³) was identified based on dose-related squamous cell metaplasia and epithelial cell hyperplasia at all doses.

Reproductive and Oral Developmental Study: Developmental Pregnant CD-1 mice were administered formaldehyde by oral gavage at 0, Toxicity 74, 148 or 185 mg/kg bw/day on gestation days 6-15. Developmental NOAEL: 185 mg/kg bw/day (highest dose tested)

Developmental Study: Pregnant female Beagle dogs were administered formaldehyde in the diet at 0, 125 or 375 mg/kg diet (reported to be 0, 3.1 and 9.4 mg/kg bw/day, respectively).

Route or Endpoint Summary study type Developmental NOAEL: 9.4 mg/kg bw/day (highest dose tested)

Dermal Developmental Study: Key study used for assessment: Pregnant female hamsters were topically administered 0 or 0.5 ml formaldehyde solution (37%) on gestation days 8, 9, 10 or 11, for 2 hours, after which the skin was washed to remove remaining formaldehyde. A developmental NOAEL of 0.5 ml (37% solution) was identified, based on a lack of treatment-related effects.

Inhalation Developmental Study: In a teratogenicity study, pregnant female Sprague-Dawley rats (25/group) were exposed to formaldehyde at 0, 2, 5 or 10 ppm (reported to be 0, 2.5, 6.2 and 12.3 mg/m3, respectively) for 6 hours/day on gestation days 6-15. Maternal NOAEC (decreased food consumption and bodyweight gain): 5 ppm (reported to be 6.2 mg/m3) Foetal NOAECs: 10 ppm (reported to be 12.3 mg/m3) (highest dose tested) Developmental NOAEC: 10 ppm (reported to be 12.3 mg/m3) (highest dose tested)

13.2.1 Evaluations by Authoritative Bodies

The International Agency for Research on Cancer (IARC) has classified formaldehyde as a Group 1 carcinogen (carcinogenic to humans). This is based on sufficient evidence in humans for the carcinogenicity of formaldehyde and sufficient evidence in experimental animals for the carcinogenicity of formaldehyde. In humans, formaldehyde is reported to cause cancer of the nasopharynx and leukaemia, and a positive association has been observed between exposure to formaldehyde and sinonasal cancer (IARC, 2012).

13.2.2 Toxicity Summary

Formaldehyde is of low to moderate acute oral toxicity, of moderate acute dermal toxicity and of moderate to high inhalation toxicity. Formaldehyde is a skin, eye and respiratory irritant in both experimental animals and humans, and it is also a skin sensitiser. Formaldehyde is genotoxic in vitro, inducing DNA damage and DNA-protein crosslinks, but it is genotoxic in vivo only at the tissues of initial contact. Similarly, in repeat dose/carcinogenicity studies, non- neoplastic and neoplastic lesions are predominantly observed at the sites of initial contact. There is also limited evidence for weak promoting potential via the oral and dermal routes. Studies in experimental animals do not indicate a specific toxicity of formaldehyde to reproduction or foetal development. It has been reported that concentrations of formaldehyde which produce toxicity at the sites of entry, do not result in a significant systemic dose and therefore do not produce systemic toxicity. In humans, formaldehyde exposure has been

associated with nasopharyngeal cancer, sinonasal cancer and leukaemia (particularly myeloid leukaemia).

13.3 Derivation of Tolerable Daily Intakes

13.3.1 Oral

In 2005 (and retained in the 4th edition of the guidelines published in 2011), the World Health Organization (WHO) concluded that, due to the high reactivity of formaldehyde, effects in the tissue of first contact following ingestion are more likely to be related to the concentration of the formaldehyde consumed rather than total intake. WHO derived a tolerable concentration of 2.6 mg/l, based on a NOEL of 260 mg/l. This NOEL was identified based on histopathological effects of the oral and gastric mucosa identified in a 2-year study in rats. An uncertainty factor of 100 (to account for inter- and intra-species variation) was applied to the NOEL to derive the tolerable concentration. (WHO, 2005; WHO, 2011).

Formaldehyde is an International Agency for Research on Cancer (IARC) Group 1 carcinogen and therefore its concentration in water should be as low as reasonably practicable, although the weight-of-evidence indicates that formaldehyde is not carcinogenic via the oral route.

13.3.2 Dermal

In a dermal developmental study, pregnant female hamsters were topically administered 0 or 0.5 ml formaldehyde solution (37%) on gestation days 8, 9, 10 or 11, for 2 hours, after which the skin was washed to remove remaining formaldehyde. A developmental NOAEL of 0.5 ml (37% solution) was identified, based on a lack of treatment-related effects.

Assuming a density for formaldehyde of 0.82 g/cm³ and a bodyweight for female hamsters of 110 g, this NOAEL would equate to a dose of 3727 mg/kg bw (rounded) for a 37% solution, or 1380 mg/kg bw/day (rounded) for a neat solution. Using this NOAEL for a neat solution and following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to account for the limited dermal toxicity database), a dermal Tolerable Daily Intake (TDI) of 1.38 mg/kg bw/day is derived.

Formaldehyde is an International Agency for Research on Cancer (IARC) Group 1 carcinogen and therefore its concentration in water should be as low as reasonably practicable, although no data were located to indicate that formaldehyde is not carcinogenic via the dermal route.

13.3.3 Inhalation

In a 28-month study, male rats were exposed to formaldehyde in aqueous methanol at 0, 0.3, 2.0 or 15 ppm (reported to be 0, 0.36, 2.4 and 17 mg/m3, respectively) for 6 hours/day, 5 days/week. A LOAEC of 0.3 ppm (reported to be 0.36 mg/m³) was identified based on dose- related squamous cell metaplasia and epithelial cell hyperplasia at all doses. Adjusting to a continuous exposure, this LOAEC would be equivalent to a concentration of 0.06 mg/m³

(rounded). Following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEC), a Tolerable Daily Concentration of 0.00006 mg/m³ (0.06 µg/m³) is derived.

Formaldehyde is an International Agency for Research on Cancer (IARC) Group 1 carcinogen and therefore its concentration in water should be as low as reasonably practicable.

13.4 Review of Current and Historical Usage

13.4.1 Production

In 1999, the global production of formaldehyde was estimated to be 5 to 6 million tonnes (OECD, 2002).

Table 13.3 Global production of formaldehyde

Country/continent Production (tonnes/year)

Asia 1 000 000 – 1 500 000 North America 1 000 000 – 1 500 000 Western Europe 2 000 000 – 2 500 000 Global 5 000 000 – 6 000 000

European production figures for 1982, 1986 and 1990 are presented in Table 13.4.

Table 13.4 European production data from 1982 to 1990 (IPCS, 2002)

Production tonnes/year Country 1982 1986 1990

United Kingdom 107 000 103 000 80 000 Former Czechoslovakia 254 000 274 000 Denmark 3 000 300 Finland 5 000 48 000 France 79 000 80 000 100 000 Germany 630 000 714 000 680 000 Hungary 13 000 11 000 Italy 125 000 135 000 114 000 Poland 219 000 154 000 Portugal 70 000 Spain 91 000 136 000

Sweden 223 000 244 000 Former Yugoslavia 108 000 99 000 88 000

In the USA, the annual growth rate of formaldehyde has increased since the late 1950‘s with an average annual growth rate of 2.7% between 1988 to 1997, between 1995 and 200 the growth rate was 1.8% which has halved to 0.9% between 2001 and 2006 and was predicted to continue to grow by 1% through 2010 (ATSDR, 2010). Table 13.5 presents the production data for the USA.

Table 13.5 Annual production volumes for the USA (ATSDR, 1999 and 2010)

Year Tonnes/year

1960 848 217 1965 1 408 857

1970 2 008 052 1975 2 067 019 1977 2 741 964 1978 2 947 894 1982 2 184 953 1986 2 516 982 1990 3 401 940 1992 3 755 742 1995 3 945 797

2000 4 649 772

In 1997, USA import volumes of formaldehyde were approximately 63 500 tonnes and in 1994 import volumes were 39 500 tonnes and export volumes were 11 300 tonnes. In 2005, 9500 tonnes were imported to the USA and 12 200 tonnes were exported. In 2006, approximately 10 000 tonnes were imported and 14 000 tonnes were exported (ATSDR, 1999 and 2010).

13.4.2 Uses

Formaldehyde is produced in large quantities industrially. It has a large range of uses including the production of urea-formaldehyde, phenolic, melamine, pentaerythritol, and polyacetal resins (WHO, 2005). Formaldehyde is also used in the synthesis of methylene dianiline (MDA), diphenyl methane diisocyanate (MDI), hexamethylenetetraamine (HTMA), trimethylol propane and neopentaglycol (OECD, 2002). Formaldehyde is also used as a chemical intermediate in the production of cosmetics, textiles and embalming fluid (WHO, 2005).

Formaldehyde is used as a disinfectant and fumigant in hospitals, ships, dwellings, and animal handling facilities. A 5% solution of formaldehyde with water as it is effective in killing most bacteria, viruses and fungi (HSE, 2008, ATSDR, 1999). Approximately 1.5% of the global use of formaldehyde (i.e. 75 000 to 90 000 tonnes/year) is as a biocide (OECD, 2002).

Medicinal uses of formaldehyde are fairly small. Formaldehyde has been used during vasectomies, it has been used as a foot antiperspirant, as a steriliser of echinicoccus cysts prior to their surgical removal. In veterinary medicine it has been used therapeutically as an antiseptic and as a fumigant. It has been used to treat tympany, diarrhoea, mastitis, pneumonia and internal bleeding in animals (ATSDR, 1999).

Formaldehyde is not approved for use as an active ingredient in pesticides within the European Union under (EC) No 1107/2009 (repealing Directive 91/414/EEC) (EU, 2013). However in the USA, it is reported that formaldehyde has been used as a fumigant for preventing mildew and spelt in wheat, a pre-planting soil steriliser in mushroom houses, a germicide and fungicide for plants and vegetables, and as an insecticide for destroying flies and other insects. It was also commonly been used in the manufacture of slow release fertilisers with 80% of the slow release fertiliser market being based on urea-formaldehyde containing products in 1999 (ATSDR, 1999). In 2002 it was reported that Canada applied 131 tonnes of formaldehyde to land as a pesticide and approximately 80% of the slow release fertiliser market in Canada was based on urea-formaldehyde products (IPCS, 2002).

A 37% solution of formaldehyde is commonly used as a tissue fixative for histology and pathology (HSE, 2008). In Canada, formaldehyde is added to foodstuffs such as maple syrup, at a concentration of <2 mg/kg (IPCS, 2002).

Table 13.6 lists the amount of the globally produced formaldehyde used for each purpose.

Table 13.6 Approximate percentages of global formaldehyde use (OECD, 2002; ATSDR, 2010)

Formaldehyde use Approximate global use % Approximate USA use in 2007 %

Urea-formaldehyde resins 40 22 Phenol formaldehyde resins 10 17 Polyacetal resins 10 13 Melamine-formaldehyde resins 5 3 MDA, MDI, HTMA, trimethylol propane, 25 10 neopentylglycol Pentaerythritol 5 5 Acetalytic agents 5 9 Disinfectants 0.5 - Biocides 1.5 -

In Swiss, Danish and Swedish Products Registers there are a large number of products that contain formaldehyde. These are both domestic and commercial products with 4000 products containing formaldehyde in Switzerland (1000 for consumer use), 1400 products in Sweden

(200 for consumer use), and 2289 products in Demark (OECD, 2002). Table 13.7 lists the approximate concentration of formaldehyde in various consumer products.

Table 13.7 Swiss products register approximate concentrations in commercial and domestic products (OECD, 2002)

Product Approximate formaldehyde content %

Paints and lacquers 0 – 10 Adhesives 0.1 – 10 Cleaning agents 0.1 – 50 Biocides 0.1 – 100 Disinfectants 0.1 – 100

Formaldehyde is a natural chemical that can be produced by the photodegradation of methane in the atmosphere (estimated to be 4 x 1011 kg/year), by the combustion of organic matter and is a product of metabolism by most organisms (HSE, 2008, IPCS, 2002 WHO, 1989). Formaldehyde can arise by the oxidation of natural organic matter during ozonation and chlorination (WHO, 2005). It is reported that it is also released from biomass combustion such as forest and brush fires (IPCS, 2002).

In 2000, it was estimated that 21 tons of formaldehyde were released to air from one site in Germany, no estimates for release to water were made. At other production site in Germany, <5 tonnes of formaldehyde were emitted to air with none going to water. In Canada, about 1424 tonnes of formaldehyde were released into the environment from industrial sites during 1997, with 20 tonnes/year released to surface waters by four sites. In the USA, 6000 tonnes formaldehyde/year were released to air and 175 tonnes/year were released to water in 1999 from industrial sites (OECD, 2002).

It has been reported that formaldehyde is a product of incomplete combustion of fuel especially in internal combustion engines. In 1997, on-road motor vehicles were reported to be the largest direct source of formaldehyde release into the environment in Canada (11 284 tonnes) (IPCS, 2002). It was also estimated that aircraft emitted 1730 tonnes and the marine sector released 1175 tonnes in the same year. These figures are expected to decrease as automotive and fuel technologies change (IPCS 2002). Other sources of fuel burning that release formaldehyde to air include: wood-burning stoves; agricultural burns; waste incinerators; cigarette smoking; and cooking food (IPCS, 2002).

In Canada, environmental release from industry was estimated to equate to: 1339.3 tonnes to air; 60.5 tonnes to deep-well injections; 19.4 tonnes to surface water; and 0 tonnes to soil. It is also reported that releases of formaldehyde to groundwater from bodies that contain embalming fluids buried in cemeteries is expected be very small (IPCS, 2002).

13.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with formaldehyde is provided in Appendix C12, and only an overall summary is provided here.

Data for formaldehyde are available for drinking water, groundwater, fresh and marine surface water, and effluents. Concentrations of formaldehyde in drinking water range from 20 to <100 µg/l; in groundwater from <50 to 690 000 µg/l; in fresh surface water from ≤9 to 133.86 µg/l; in marine surface water it has not been detected, and in effluents levels range from detected but not quantified to 325 µg/l.

13.6 Review of Literature Data on Removal during Sewage Treatment

Studies are available reporting removal of formaldehyde during sewage and data available for the toxicity of formaldehyde to sewage treatment organisms indicating that it is readily biodegradable by a range of organisms and of low to moderate acute and chronic toxicity to sewage treatment organisms. The available data are summarised in Table 13.8 and Table 13.9.

Table 13.8 Summary of studies reporting formaldehyde removal during sewage treatment

Concentration Method/type of study Inoculum Results Reference (mg/l) Aerobic

Acclimatised for 30 activated Verschuren (1996) 333 94% reduction after 5 days. days at 20°C sludge cited in HSDB (2013) Acclimatised for 30 activated Verschuren (1996) 500 47% reduction after 5 days. days at 20°C sludge cited in HSDB (2013) Acclimated for less activated Verschuren (1996) 720 No biodegradation after 5 days. than 1 day sludge cited in HSDB (2013) Belly and Goodhue 14 day CO2 evolution activated NR 10.1 – 18.2 % CO2 recovered (1976) cited in SRC screening test sludge (2013) activated 90% BOD in 5 days, 103% BOD in Dickerson et al. (1955) BOD screening test NR sludge 10 days and 99% BOD in 20 days cited in SRC (2013) Warburg BOD activated Dickerson et al. (1955) 300 96% BOD in 0.25 days screening test sludge cited in SRC (2013) Warburg BOD activated Dickerson et al. (1955) 400 94% BOD in 0.25 days screening test sludge cited in SRC (2013) Warburg BOD activated Dickerson et al. (1955) 600 110% BOD in 0.25 days screening test sludge cited in SRC (2013)

Concentration Method/type of study Inoculum Results Reference (mg/l)

Warburg BOD activated Dickerson et al. (1955) 800 106% BOD in 0.25 days screening test sludge cited in SRC (2013) Warburg BOD activated Dickerson et al. (1955) 1000 90% BOD in 0.25 days screening test sludge cited in SRC (2013) Warburg BOD activated Dickerson et al. (1955) 1500 57% BOD in 0.25 days screening test sludge cited in SRC (2013) Gellman and Warburg BOD sewage 130 0% BOD in 1 day Heukelekian (1950) screening test, 20˚C cited in SRC (2013)

Gellman and Warburg BOD sewage 445-750 95% BOD in 1 day Heukelekian (1950) screening test, 20˚C cited in SRC (2013) Gerhold and Malaney Warburg BOD activated 500 0% BOD in 1 day (1966) cited in SRC screening test, 20˚C sludge (2013) Heukelekian and Rand Warburg BOD sewage 130 94% BOD in 7 days (1955) cited in SRC screening test, 20˚C (2013) Heukelekian and Rand Warburg BOD sewage 260 103% BOD in 5 days (BOD5) (1955) cited in SRC screening test, 20˚C (2013) Heukelekian and Rand Warburg BOD sewage 2.5-10 31% BOD in 5 days (BOD5) (1955) cited in SRC screening test, 20˚C (2013) Heukelekian and Rand Warburg BOD sewage 1.7-20 42% BOD in 5 days (BOD5) (1955) cited in SRC screening test, 20˚C (2013) Pauli and Franke SAPROMAT screening sewage 60 50% degradation after 1 day (1971) cited in SRC test at pH 7.2-7.6 (2013) Pauli and Franke SAPROMAT screening sewage 60 100% degradation after 1.5 days (1971) cited in SRC test at pH 7.2-7.6 (2013) Pauli and Franke SAPROMAT screening sewage 240 35% degradation after 1 day (1971) cited in SRC test at pH 7.2-7.6 (2013) Pauli and Franke SAPROMAT screening sewage 240 85% degradation after 1.5 days (1971) cited in SRC test at pH 7.2-7.6 (2013)

Concentration Method/type of study Inoculum Results Reference (mg/l)

Pauli and Franke SAPROMAT screening sewage 240 100% degradation after 2.5 days (1971) cited in SRC test at pH 7.2-7.6 (2013) Placak and Ruchhoft BOD screening test, activated 0% BOD, but toxic at higher 720 (1947) cited in SRC 20˚C sludge concentrations (2013) Stafford and Northup BOD screening test, sewage 1000 37% BOD after 5 days (1955) cited in SRC 20˚C (2013)

Biological Treatment simulation, BOD, domestic Hatfield (1957) cited in acclimated to methanol 500 57% BOD reduction after 3 hours sewage SRC (2013) or acetic acid for 30 days Biological Treatment simulation, BOD, domestic Hatfield (1957) cited in acclimated to methanol 333 82% BOD reduction after 8 hours sewage SRC (2013) or acetic acid for 30 days Biological Treatment activated Dickerson et al. (1955) simulation, pH 4.6 with 3000 0% removal sludge cited in SRC (2013) 24 hours aeration Biological Treatment activated Dickerson et al. (1955) simulation, pH 6.0 with 3000 80% removal sludge cited in SRC (2013) 24 hours aeration Biological Treatment activated Dickerson et al. (1955) simulation, pH 7.2 with 3000 99% removal sludge cited in SRC (2013) 24 hours aeration degrades readily in aerobic systems with acclimation, however, above this level the fall in pH, Dickerson et al. (1955) NR NR ≤100 caused by the accumulation of cited in HSDB (2013) oxidation products, inhibits further degradation oxidation of concentrations up to Dickerson et al. (1955) NR NR ≤1750 1750 mg/l cited in HSDB (2013) Howard (1989) cited in NR NR NR biodegradable following acclimation HSDB (2013) Anaerobic

biodegradable under anaerobic Howard (1989) cited in NR NR NR conditions following acclimation HSDB (2013)

NR: Not reported

Table 13.9 Summary of studies reporting toxicity of formaldehyde to sewage treatment organisms

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

Toxicity threshold (cell Protozoa (Uronema 20 6.5 IUCLID (2000) multiplication) parduczi) Toxicity threshold (cell Protozoa (Entosiphon 72 22 IUCLID (2000) multiplication) sulcatum) Toxicity threshold (cell Protozoa (Chilomonas 48 4.5 IUCLID (2000) multiplication) paramaecium)

Toxicity threshold (cell Pseudomonas putida 16 14 IUCLID (2000) multiplication) Toxicity threshold (cell Mycrocystis aeruginosa 8 days 0.39 IUCLID (2000) multiplication) Minimum inhibitory Alcaligenes sp. 72 50 IUCLID (2000) concentration Toxicity threshold (glucose Pseudomonas 16 14 IUCLID (2000) assimilation) fluorescens EC10 (effect not specified) Industrial activated sludge 7 day >1995 IUCLID (2000) EC20 (effect not specified) Activated sludge 0.5 >700 IUCLID (2000) Gerike and Gode, Inhibitory threshold Aerobic sewage sludge NR 30 (1990) Hickey et al. IC20 (effect not specified) Methanogenic bacteria 5 10 (1987) Hickey et al. IC90 (effect not specified) Methanogenic bacteria 5 100 (1987) IC100 (effect not specified) Methanogenic bacteria - >200 Vidal et al. (1998)

EC10: Effect Concentration – A point estimate of the toxicant concentration that would cause an observable effect (such as death, immobilisation or a serious incapacitation) in 10% of the test organisms.

EC20: Effect Concentration – A point estimate of the toxicant concentration that would cause an observable effect (such as death, immobilisation or a serious incapacitation) in 20% of the test organisms.

IC20: Inhibition Concentration – A point estimate of the toxicant concentration that would cause 20% reduction in a non-lethal biological measurement of the test organisms, such as reproduction or growth.

IC90: Inhibition Concentration – A point estimate of the toxicant concentration that would cause 90% reduction in a non-lethal biological measurement of the test organisms, such as reproduction or growth

NR: Not Reported

Toxicity Threshold: A predicted concentration of a substance that may lead to toxic effects.

13.7 Review of Literature Data on Removal during Drinking Water Treatment

Formaldehyde is reported to be rapidly hydrolysed to glycol (WHO, 2002). Therefore, it is not expected to persist in water. The information provided below is based on a literature search specifically on the removal of formaldehyde, rather than this breakdown product, and a distinction between the two has not been made in these studies. However, it is probable that in these studies, at least some of the measured removal would be the result of hydrolysis.

Krasner et al., (1993) have assessed the effectiveness of biological filtration, using a dual media of sand/granular activated carbon (GAC) (coal or wood-based) or sand/anthracite coal on the removal of a range of aldehydes, including formaldehyde. Water was pre-ozonated and had undergone coagulation and sedimentation and the filters were operated at 14 metres/hour to give an empty bed contact time (EBCT) of 2.1 minutes. Influent concentrations were 7-12 mg/l. The results of this study are presented in Table 13.10. Krasner et al. (1993) stated that GAC developed biological activity sooner than anthracite.

Table 13.10 Removal of formaldehyde by sand/coal-based GAC, sand/wood-based GAC and sand/anthracite coal media

Sand/Coal-based Sand/Wood-based Sand/anthracite GAC GAC coal

Days to achieve 50% removal <15 8 <32

Days to achieve 80% removal 18 15 36

In another study using sand/anthracite biological filters and an EBCT of 7 minutes, approximately 85% of an initial concentration of 10 µg formaldehyde/l was reported to be removed. Backwashing with chlorinated water (1 mg/l) resulted in poorer removal than backwashing with non-chlorinated water (Miltner et al., 1995).

Wang and Summers (1966) reported that using a concentration of 7 µg formaldehyde/l and sand filters with a range of contact times that approximately 60% removal of formaldehyde occurred after an EBCT of 2 minutes, but no further removal occurred with EBCTs up to 7 minutes.

Greater than 80% removal of formaldehyde (7 to 25 μg/l in ozonised humic water) was obtained using biofilters packed with different media. EBCTs in the range 5 to 50 minutes had little effect on removal efficiency (Melin and Odegaard, 2000).

Based on a Henry‘s Law constant of 3.37 x10-7 atm.m³/mol at 25°C (SRC, 2013), volatilisation from water surfaces is not expected to be significant, and therefore, it is unlikely to undergo substantial removal by air stripping.

Chlorination is not expected to reduce formaldehyde concentrations during water treatment; no significant chlorine consumption was reported by Fielding et al. (1989) using a concentration of 10 mg formaldehyde/l in aqueous solution buffered at pH 7 and using 1 mg chlorine/l.

Ozonation is not an effective removal mechanism for formaldehyde, and reaction with organic material may result in its formation. Jammes et al. (1995) conducted a study at a full-scale treatment plant that reported increases in formaldehyde concentration by 2-4 times the level in raw water following ozonation. Filtration with GAC reduced the concentration of formaldehyde to levels similar to the raw water.

However, another study has indicated that formaldehyde may be removed by ozonation, particularly in combination with UV. Formaldehyde (48 mg/l) was treated with ozone (7.8 mg/l.minute), UV (18.6 mW/cm² at 254 nm), and ozone plus UV for up to 3 hours. Removal, based on total organic carbon (TOC), is provided in Table 13.11 (Takahashi, 1990).

Table 13.11 Removal of formaldehyde by ozone, UV and ozone/UV

Removal (%) Treatment 0.5 hours 1 hour 1.5 hours 2 hours 2.5 hours 3 hours

Ozone 17 35 54 73 80 86

UV 2 4 7 10 14 16

Ozone/UV 80 97 100 100 100 100

14. o-Toluidine

14.1 Physico-Chemical Properties

Data on the physico-chemical properties of o-toluidine are provided in Table 14.1.

Table 14.1 Physico-chemical properties of o-toluidine

CAS Number 95-53-4

Chemical Formula C7H9N

Structure

Molecular weight 107.155 (ChemID, 2013)

Physical state Light yellow liquid that turns reddish brown on exposure to air and light (OECD, 2004)

Melting point -24.4 °C (α-form), -16.3 °C (β-form) (OECD, 2004)

Boiling point 200.25°C at 1013 hPa (OECD, 2004)

Water solubility 15 000 mg/l at 25°C (OECD, 2004); 16 220 mg/l at 20°C (OECD, 2004)

Log Kow 1.4 at 24.5°C (OECD, 2004)

Koc 52 (OECD, 2004); 40-250 (HSDB, 2013)

Vapour pressure 34.5 Pa at 25°C (OECD, 2004)

Henry’s Law constant 1.98 x10-6 atm.m³/mol at 25°C (SRC, 2013)

Dissociation constant pKa: 4.44-4.45 (OECD, 2004)

Density 0.9984 g/cm3 at 20°C (OECD, 2004)

14.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with o-toluidine is provided in Appendix B13, and only an overall summary is provided here (Table 14.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 14.2 Summary of key toxicological data for o-toluidine

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 520 mg/kg bw (mice) to 1710 mg/kg bw (rats)

Dermal Low acute toxicity LD50: 3250 mg/kg bw (rabbits)

Inhalation Low acute toxicity LC50: 862 ppm (reported to be 3827 mg/m3) (rats)

Irritation and - Non-irritating to moderately irritating to the skin and to be highly irritating to Sensitisation the eyes of experimental animals

Genotoxicity In vitro and Overall o-toluidine shows some genotoxic potential and clastogenic activity in in vivo vitro, however, the in vivo genotoxic potential of o-toluidine remains uncertain.

Repeat Dose Oral Chronic Study: Key study used for assessment: Toxicity and In a 2-year study, mice were administered o-toluidine hydrochloride at 0, Carcinogenicity 1000 or 3000 mg/kg diet (reported to be approximately 0, 150 and 450 mg/kg bw/day, respectively) A LOAEL of 1000 mg/kg diet (reported to be approximately 150 mg/kg bw/day) was identified reductions in bodyweight.

Dermal One dermal study was located, however, it was not considered appropriate for use in this assessment

Inhalation No data were located

Reproductive and Oral In a subchronic dietary study in male rats Fisher 344 rats administered o- Developmental toluidine hydrochloride for various lengths of time, slight reductions in Toxicity absolute testis weight and absolute epididymis weight and an increase in relative testis weights have been observed. However, these effects were considered secondary to changes in bodyweight.

Dermal Developmental Study: Key study used for assessment: In a dermal developmental study, rats were administered o-toluidine at doses of 0, 8 or 80 mg/kg bw to their tail skin. A NOAEL of 8 mg/kg bw was identified based on changes in body and organ weight in offspring.

Inhalation No data were located

14.2.1 Evaluations by Authoritative Bodies

The International Agency for Research on Cancer (IARC) has classified o-toluidine as a Group 1 carcinogen (carcinogenic to humans). This is based on sufficient evidence in humans for the carcinogenicity of o-toluidine (cancer of the urinary bladder) and sufficient evidence in

experimental animals for the carcinogenicity of o-toluidine. IARC also state that there is moderate mechanistic evidence indicating that the carcinogenicity of o-toluidine involves metabolic activation, formation of DNA adducts, and induction of DNA-damaging effects (IARC, 2012).

This classification of o-toluidine as carcinogenic to humans is controversial due to the fact that in epidemiological studies, many of the workers were exposed to other aromatic amines, such as aniline and 4-aminobiphenyl, in addition to o-toluidine (IARC, 2010).

14.2.2 Toxicity Summary

o-Toluidine is of low acute oral, dermal and inhalation toxicity. It is reported to be non-irritating to moderately irritating to the skin and to be highly irritating to the eyes of experimental animals. Results of both in vitro and in vivo genotoxicity studies are mixed. However, overall, o-toluidine is considered to have some genotoxic and clastogenic potential. Effects observed in repeat dose toxicity studies are similar to those seen with other monocyclic aromatic amines, such as aniline. These effects include methaemoglobin formation, haemosiderosis, and effects on the spleen and urinary bladder. o-Toluidine has also been associated with urinary bladder cancer in occupationally exposed workers, and o-toluidine has been classified as a Group 1 carcinogen (i.e. carcinogenic to humans). Very limited information is available on the reproductive and developmental toxicity of o-toluidine. However, there is some evidence of effects on parental reproductive organs (weight changes and degeneration) and various organ weight changes in offspring. In light of its genotoxic and carcinogenic potential, it has been stated that o-toluidine should be regarded as potentially toxic to reproduction.

14.3 Derivation of Tolerable Daily Intakes

14.3.1 Oral

In a 2-year study, mice were administered o-toluidine hydrochloride at 0, 1000 or 3000 mg/kg diet (reported to be approximately 0, 150 and 450 mg/kg bw/day, respectively). A LOAEL of 1000 mg/kg diet (reported to be approximately 150 mg/kg bw/day) was identified from this study. Using this LOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the use of a LOAEL) an oral Tolerable Daily Intake (TDI) of 0.15 mg/kg bw/day (150 µg/kg bw/day) is derived.

However, o-toluidine is an International Agency for Research on Cancer (IARC) Group 1 carcinogen and therefore its concentration in water should be as low as reasonably practicable. IARC state that there is moderate mechanistic evidence indicating that the carcinogenicity of o-toluidine involves metabolic activation, formation of DNA adducts, and induction of DNA-damaging effects.

14.3.2 Dermal

In a dermal developmental study, rats were administered o-toluidine at doses of 0, 8 or 80 mg/kg bw to their tail skin. A NOAEL of 8 mg/kg bw was identified based on changes in body and organ weight in offspring. Using this NOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra-species variation and 10 to account for the limitations of the study and general limitations of the dermal toxicity database), a dermal TDI of 0.008 mg/kg bw/day (8 µg/kg bw/day) is derived.

14.3.3 Inhalation

No data inhalation toxicity data were located. However, an inhalation Tolerable Daily Concentration can be derived by extrapolation from oral toxicity data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

In a 2-year study in mice, a LOAEL of 1000 mg/kg diet (reported to be approximately 150 mg/kg bw/day) was identified.

Assuming an inhalation volume of 0.043 m³ per day and a bodyweight of 0.038 kg (38 g), this would equate to a LOEAC of 133 mg/m³ (rounded).

Following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 for the use of a LOAEL) and multiplying by 50/100 (an arbitrary assumption within the IGHRC guidelines to account for the differences between oral and inhalation absorption), an inhalation Tolerable Daily Concentration of 0.067 mg/m³ (67 µg/m³; rounded) can be derived.

It should be noted that extrapolation from the oral route to the inhalation route may underestimate toxicity via the inhalation route, as absorption via inhalation is likely to be at least as efficient, but may conceivably be more efficient, that oral exposure. However, the adjustment of 50/100 is considered sufficiently precautionary to account for such underestimates.

14.4 Review of Current and Historical Usage

14.4.1 Production

In 2001, world-wide production of o-toluidine was estimated to be 59 000 tonnes by 11 producers (OECD, 2004).

Table 14.3 Estimated production volumes in 2001 (OECD, 2004)

Region Estimated production volume (tonnes/year)

Western Europe 35 000 USA 10 000 China 12 000 India 2000

Total production of o-toluidine in Great Britain was approximately 6000 tonnes per year, as of 1998, with 90% of the chemical being exported. In 1975 in the USA, 900 tonnes of o-toluidine were produced and 1000 tonnes were imported. Approximately 610 tonnes and 545 tonnes for o-toluidine were imported into Japan in 1992 and 1993, respectively (WHO, 1998).

In the USA, commercial production began in 1956. Annual production in the USA of o- toluidine was estimated to be between 500 tonnes and 5000 tonnes in the late 1970s. This increased to between 6600 tonnes to 12 800 tonnes in the early 1990s (National Toxicology Programme, 2011).

14.4.2 Usage

o-Toluidine is used as a chemical intermediate in chemical synthesis of herbicides, rubber chemicals, dye and pigment intermediates, resin hardeners, fungicide intermediates, corrosion inhibitor in paints and pharmaceutical intermediates. In 1997 in Western Europe, 25 800 tonnes were used for these purposes (OECD, 2004, WHO, 1998).

The largest use of o-toluidine is for the preparation of methyl ethyl aniline (MEA, 6-ethyl-o-toluidine) which is used as an intermediate in the manufacture of herbicides such as acetochlor, metochlor and propisochlor. It is also used in the synthesis of rubber chemicals such as di-tolyl-phenyl-p-phenylenediamine (DTPD), which is a rubber antioxidant (OECD, 2004).

It is also converted into dye and pigment intermediates like acetoacet-o-toluidine and 3-hydroxy-2-napthoyl-o-toluidine (OECD, 2004). It is used in acid fast dyestuffs, azo pigments and dyes, sulphur dyes, indigo compounds and as a photographic dye. o-Toluidine is also used for the photometric determination of glucose in the blood (OECD, 2004, National Toxicology Programme, 2011).

No direct consumer use for o-toluidine is listed in the Danish, Finnish, Norwegian or Swedish product registers. The Swiss product register lists analytical kits, and analytical substance and metallic mordants as containing o-toluidine it has been assumed these products are for industrial use (OECD, 2004).

Within the European Union (EU) the use of azo dyes which release o-toluidine upon degradation are not permitted for textiles or other consumer articles. Exposure to consumers and releases to the environment from consumer products is believed to be negligible (OECD, 2004).

o-Toluidine is usually manufactured and processed within close systems although small amounts have been detected in wastewater effluents. O-Toluidine occurs naturally in some vegetables, tobacco leaves and black tea. It is an intermediate in the biodegradation of o-nitrotoluene which can be found at old munitions sites. It can also be formed during pyrolysis and is a component of coal oil (OECD, 2004).

14.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with o-toluidine is provided in Appendix C13, and only an overall summary is provided here.

Data for o-toluidine are available for groundwater, fresh surface water, and effluents. No data were located for drinking water or marine water. Concentrations of o-toluidine in groundwater range from 0.06 to 9.2 µg/l (combined o and p isomers), in fresh surface water from 0.03 to 20 µg/l and in effluents from 42.5 to 18843 µg/l.

14.6 Review of Literature Data on Removal during Sewage Treatment

There are studies available reporting removal of o-toluidine during sewage and data available for the toxicity of o-Toluidine to sewage treatment organisms indicate that it is readily biodegradable by a range of organisms under aerobic conditions, but not biodegraded under anaerobic conditions and it is of low to moderate acute toxicity to sewage bacteria. The available data are summarised in Table 14.4 and Table 14.5.

Table 14.4 Summary of studies reporting o-toluidine removal during sewage treatment

Concentration Method/type of study Inoculum Results Reference (mg/l) Aerobic

OECD guideline 301A 88-90% after 28 days (74- Activated readily biodegradability test, NR 89% after 7 days and 73 - IUCLID (2000) sludge DOC analysis. >90% after 14 days) OECD 301C, readily Activated biodegradability: Modified 100 65.4% after 28 days IUCLID (2000) sludge MITI test, non-adapted.

Concentration Method/type of study Inoculum Results Reference (mg/l)

OECD 301E, readily >90% after 28 days (57 to biodegradability: Modified Activated NR >90% after 7 days and >90% IUCLID (2000) OECD screening test, DOC sludge after 14 days) analysis. OECD guideline 301E, Waste water readily biodegradability: 97% after 21 days (7 days treatment 20 IUCLID (2000) Modified OECD screening 86% and 14 days 95%) effluent test OECD guideline 302 Industrial 96% after 11 days (3 hours inherent biodegradability : activated NR 20-30%, 5 days 20%, 10 IUCLID (2000) Modified Zahn Wellens test. sludge days 77%) Activated sludge Activated 100 92% after 1 day IUCLID (2000) degradability test, COD sludge Activated sludge Activated 100 83% after 1 day IUCLID (2000) degradability test, TOC sludge

Industrial OECD guideline 302 activated inherent biodegradability : NR 96% after 15 days IUCLID (2000) sludge – non- Modified Zahn Wellens test. adapted Warburg Screening test, GC Baird et al. (1977) Activated analysis, incubated for 0.25 20 100% degradation cited in SRC sludge days at 25˚C (2013) Warburg Screening test, Malaney (1960) Activated MANO analysis, incubated 500 57% BODT degradation cited in SRC sludge for 5-8 days at 20˚C (2013) Brown and OECD screening test with Laboureur (1983) DOC analysis, 28 day Sewage 20 (DOC) ≥90% cited in SRC incubation (2013) Brown and OECD Modified test: French Laboureur (1983) AFNOR with DOC analysis, Sewage 20 (DOC) ≥90% cited in SRC 28 day incubation (2013) BOD screening test at 20 ˚C Heukelekian and with 5 days incubation Sewage 1.25-2.5 56 – 57 %BODT Rand (1955) cited (BOD5) in SRC (2013) BOD: Japanese MITI Activated Kitano (1978) cited screening test with 14 days 100 30 – 100% degradation sludge in SRC (2013) incubation at 25 ˚C and pH7 Activated COD test 100 92% after 1 day IUCLID (2000) sludge

Concentration Method/type of study Inoculum Results Reference (mg/l)

COD removal screening test with 5 days incubation and Activated Pitter (1976) cited 20 days acclimatisation, 20 200 97.7% removal sludge in SRC (2013) ˚C and pH 7.2, vigorous system Biological treatment Matsui et al, Activated 92% COD removal after 24 simulation test with COD 306 (1975) cited in sludge hours analysis SRC (2013) Anaerobic

Kuhn and Suflita Biodegradation study, 10 Aquifer slurry NR Not biodegraded (1989) and cited in month incubation period. HSDB (2013)

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound

COD: Chemical Oxygen Demand. The amount of oxygen consumed during oxidation of a compound with hot acid dichromate; it provides a measure of the oxidisable matter present

DOC: Dissolved Organic Carbon. The amount of carbon present in organic compounds in aqueous solution

GC: Gas chromatography

MITI: Ministry of International Trade and Industry, Japan

NR: Not reported

OECD: The Organisation for Economic Co-operation and Development. Standardised methods of testing for assessing the potential effects of chemicals on human health and the environment.

TOC: Total Organic Carbon. The total amount of organic carbon present in an aqueous solution.

Table 14.5 Summary of studies reporting toxicity of o-toluidine to sewage treatment organisms

Duration Concentration Method/type of study Inoculum Reference (hours) (mg/l)

EDTA fermentation Anaerobic bacteria from a domestic IUCLID 24 1250 tube method water treatment plant (2000) IUCLID Toxicity threshold Pseudomonas putida 16 16 (2000) IUCLID Toxicity threshold Entosiphon sulcatum 72 76 (2000)

EDTA: Ethylenediaminetetraacetic acid

Toxicity Threshold: A predicted concentration of a substance that may lead to toxic effects.

14.7 Review of Literature Data on Removal during Drinking Water Treatment

o-Toluidine may be readily removed from drinking water by Fenton reactions. In a study using 2+ a fluidised-bed process (1.35 litre fluidised-bed reactor), 1 mM Fe and 17 mM H2O2 at pH 3±0.5, removal of o-toluidine was reported to be 99.8% (Anotai et al., 2012).

No other data on the removal of o-toluidine during drinking water treatment were located, however, some inference on its likely fate can be made based on its physico-chemical properties and fate in the environment.

o-Toluidine is not expected to adsorb to carbon, based on Kocs of 52 (OECD, 2004) and 40-250 (HSDB, 2013). Hydrolysis is also not expected to be an important fate process (OECD, 2004). However, it is expected to volatilise very slowly from water surfaces, based on a Henry‘s Law constant of 1.98 x10-6 atm.m³/mol at 25°C (SRC, 2013), and therefore, it is not expected to be amenable to air stripping. Volatilisation half-lives of 19 and 140 days have been estimated for a model river and model lake, respectively (HSDB, 2013).

15. Quinoline

15.1 Physico-Chemical Properties

Data on the physico-chemical properties of quinoline are provided in Table 15.1.

Table 15.1 Physico-chemical properties of quinoline

CAS Number 91-22-5

Chemical Formula C9H7N

Structure

Molecular weight 129.1613 (ChemID, 2013)

Physical state Colourless to brown liquid (HSDB, 2013)

-19.5°C (IUCLID, 2000); -15.6°C (IUCLID, Melting point 2000); -14.2°C (IUCLID, 2000)

2371°C at 1013 hPa (IUCLID, 2000); Boiling point 238.1°C at 1013 hPa (IUCLID, 2000)

6110 mg/l at 20°C (IUCLID, 2000); Water solubility 60 000 mg/l at 25°C (IUCLID, 2000)

Log Kow 1.88-2.06 (IUCLID, 2000)

Koc 2.84-10.9 (HSDB, 2013)

0.012 hPa at 25°C (IUCLID, 2000); Vapour pressure 0.06 mmHg at 25°C (SRC, 2013)

Henry’s Law constant 1.67 x10-6 atm.m³/mol at 25°C (SRC, 2013)

Dissociation constant pKa: 4.9 at 20°C (SRC, 2013)

Density 1.0938 g/cm³ at 20°C (IUCLID, 2000)

15.2 Toxicological Data Review

For the ease of reading, a detailed review of mammalian toxicity studies with quinoline is provided in Appendix B14, and only an overall summary is provided here (Table 15.2). Where a specific study is used later in this report in the derivation of a Tolerable Daily Intake (TDI), that study has been identified as a ―Key study used for assessment‖.

Table 15.2 Summary of key toxicological data for quinoline

Route or Endpoint Summary study type

Acute toxicity Oral Low acute toxicity LD50 range: 270-460 mg/kg bw (rats)

Dermal Low acute toxicity LD50 range: 540 mg/kg bw (rabbits) to 1370 mg/kg bw rats)

Inhalation Low acute toxicity No mortalities in quinoline saturated air

Irritation and - Irritating to the skin and eyes, but is not considered to be a skin sensitiser Sensitisation

Genotoxicity In vitro and The results of in vitro and in vivo assays are generally positive for in vivo mutagenicity.

Repeat Dose Toxicity Oral Sub-chronic Study: Key study used for assessment: and Carcinogenicity In a 40-week study, male Sprague-Dawley rats were administered quinoline in their diet at doses of 0, 0.05, 0.1 or 0.25% (0, 500, 1000 or 2500 mg/kg diet; approximately 0, 25, 50 and 125 mg/kg bw/day, respectively). A LOAEL of 0.05% (500 mg/kg diet; approximately 25 mg/kg bw/day) was identified based on haemagioendothelioma and increased aspartate aminotransferase (AST) and alkaline phosphatase.

Dermal In a tumour promotion initiation study, quinoline was applied as a 0.5% solution in acetone (0.1 ml) to the shaved backs of female Sencar mice (dose reported to be 5 mg/day), and was considered to be a tumour initiator.

Inhalation No data were located

Reproductive and Oral No data were located Developmental Toxicity Dermal No data were located

Inhalation No data were located

15.2.1 Toxicity Summary

Quinoline is of low acute oral, dermal and inhalation toxicity to experimental animals. It is irritating to the skin and eyes, but is not considered to be a skin sensitiser. The weight of evidence indicates that quinoline is genotoxic in vitro and in vivo. Oral repeat dose studies indicate that the liver is the target organ of toxicity, and tumours are reported to largely develop in this organ, with tumours in the lungs believed to be metastatic tumours from the liver. Dermal studies also indicate that quinoline may be a tumour promotor. No oral, dermal or inhalation reproductive or developmental studies were located on quinoline.

15.3 Derivation of Tolerable Daily Intakes

15.3.1 Oral

In a 40-week study, male Sprague-Dawley rats were administered quinoline in their diet at doses of 0, 0.05, 0.1 or 0.25% (0, 500, 1000 or 2500 mg/kg diet; approximately 0, 25, 50 and 125 mg/kg bw/day, respectively). A LOAEL of 0.05% (500 mg/kg diet; approximately 25 mg/kg bw/day) was identified based on haemagioendothelioma and increased aspartate aminotransferase (AST) and alkaline phosphatase at all doses. Using this LOAEL and following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to account for the use of a LOAEL), an oral Tolerable Daily Intake (TDI) of 0.025 mg/kg bw/day (25 µg/kg bw/day) is derived.

15.3.2 Dermal

Although several dermal toxicity studies were located, they are not considered appropriate for use in the derivation of a dermal TDI.

No data were located to suggest that dermal exposure will significantly impair the integrity of skin (i.e. quinoline is not corrosive), a dermal TDI can be derived by extrapolation from oral toxicity data in accordance with the Interdepartmental Group on Health Risks from Chemicals (IGHRC) guidelines (IGHRC, 2006).

Assuming equivalent bioavailability for the oral and dermal routes, a dermal TDI of 0.025 mg/kg bw/day (25 µg/kg bw/day) can be derived.

However, it should be noted that this is likely to be a highly conservative value, as in reality, very few substances are absorbed as readily via the dermal route as they are via the oral route.

15.3.3 Inhalation

In a 40-week study in rats, a LOAEL of 0.05% (500 mg/kg diet; approximately 25 mg/kg bw/day) was identified based on haemagioendothelioma and increased aspartate aminotransferase (AST) and alkaline phosphatase at all doses.

Assuming an inhalation volume of 0.29 m³ per day and a bodyweight of 0.425 kg (425 g), this would equate to a NOEAC of 37 mg/m³ (rounded).

Following application of an uncertainty factor of 1000 (100 to account for inter- and intra- species variation and 10 to account for the limited database) and multiplying by 50/100 (an

arbitrary assumption within the IGHRC guidelines to account for the differences between oral and inhalation absorption), an inhalation Tolerable Daily Concentration of 0.019 mg/m³ (19 µg/m³; rounded) can be derived.

15.4 Review of Current and Historical Usage

15.4.1 Production

In 1990 in the US, the annual production of quinolone was estimated to be 17.2 tonnes. In 1994 the estimated amount of quinolone released to the environment in the US was 41.2 tonnes (California Environmental Protection Agency 1997).

15.4.2 Usage

Quinoline can be found in petroleum, coal processing, coal tar, wood preservation (creosote) and shale oil. Quinoline can be produced by combustion of a number of substances including tobacco (California Environmental Protection Agency 1997, US EPA, 2001). It is used as an intermediate in the production of 8-hydroxyquinoline (fungi growth inhibitor and chemical intermediate for pharmaceuticals such as anti-infectives), hydroxyquinoline sulphate and copper-8-hydroxyquinolate (flavouring). It is used as a solvent for resins and terpenes and in the production of paints (US EPA, 2001, HSDB, 2013). Quinoline has anti-malarial activity but does not appear to be used for this purpose (California Environmental Protection Agency 1997).

Quinoline is used as a solvent and decarboxylation reagent and is a raw material in the production of dyes, antiseptics, fungicides, niacin (vitamin B3), and pharmaceuticals (California Environmental Protection Agency 1997).

The main application of quinolone is the production of 8-quinolinol. It can be used in the production of methane dyes and nicotinic acid. Quinoline can be used as an effective solvent and extractor of polyaromatic compounds (HSDB, 2013).

15.5 Review of Occurrence Data

For the ease of reading, a detailed review of occurrence studies with quinoline is provided in Appendix C14, and only an overall summary is provided here.

Data for quinoline are available for groundwater, fresh and marine surface water and effluents. No data were located for drinking water. Concentrations of quinoline in groundwater

range from <0.006 to 11 200 µg/l, in fresh surface water of 20 µg/l, in marine surface water of 0.0017 µg/l and in effluents from 0.02 to 10 000 µg/l.

15.6 Review of Literature Data on Removal during Sewage Treatment

There are two studies available reporting removal of quinoline during sewage and data available for the toxicity of quinoline to sewage treatment organisms and indicate that it not readily or inherently biodegradable. Quinoline is of moderate to low acute toxicity to sewage treatment organisms. The available data are summarised in Table 15.3 and Table 15.4.

Table 15.3 Summary of studies reporting quinoline removal during sewage treatment

Concentration Method/type of study Inoculum Results Reference (mg/l) Aerobic

BOD screening test at Heukelekian and Rand (1955) 20 ˚C with 5 days sewage NR 69 %BODT cited in SRC (2013) incubation (BOD5) BOD: Japanese MITI screening test with 14 activated Sasaki (1978) cited in SRC 100 0 – 29% degradation days incubation at 25 sludge (2013) ˚C and pH7 Closed bottle test with Pseudomonas 2.5 mmol/l 100% after 18 hours IUCLID (2000) UV and GC analysis putida Closed bottle test with Pseudomonas 2.5 mmol/l 100% after 24 hours IUCLID (2000) UV and GC analysis aeruginosa Closed bottle test with sewage 100% within 40-60 NR IUCLID (2000) UV and GC analysis effluent hours 69% of Theoretical BOD5 NR NR IUCLID (2000) BOD 65% of Theoretical COD NR NR IUCLID (2000) COD

BOD: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms when metabolising a compound.

BOD5: Biological Oxygen Demand. The amount of oxygen consumed by microorganisms in 5 days when metabolising a compound.

COD: Chemical Oxygen Demand. The amount of oxygen consumed during oxidation of a compound with hot acid dichromate; it provides a measure of the oxidisable matter present

MITI: Ministry of International Trade and Industry, Japan

NR: Not reported

Table 15.4 Summary of studies reporting toxicity of quinoline to sewage treatment organisms

Method/type of Duration Concentration Inoculum Reference study (hours) (mg/l)

EC0 (effect not Escherichia coli NR 1000 IUCLID (2000) stated) EC10 (effect not Escherichia coli 48 600 IUCLID (2000) stated) EC100 (effect not Protozoa (Tetrahymena 24 6.19 mmol/l IUCLID (2000) stated) pyriformis)

EC50 (effect not Protozoa (Tetrahymena 60 34.34 – 130.29 IUCLID (2000) stated) pyriformis) EC50 (growth Neuwoehner et al, Pseudomonas putida NR >98.8 inhibition) (2009)

EC0: Effect Concentration – A point estimate of the toxicant concentration that would cause an observable effect (such as death, immobilisation or a serious incapacitation) in 0% of the test organisms.

EC50: Effect Concentration – A point estimate of the toxicant concentration that would cause an observable effect (such as death, immobilisation or a serious incapacitation) in 50% of the test organisms.

EC100: Effect Concentration – A point estimate of the toxicant concentration that would cause an observable effect (such as death, immobilisation or a serious incapacitation) in 100% of the test organisms.

NR: Not Reported

15.7 Review of Literature Data on Removal during Drinking Water Treatment

Quinoline is expected to undergo some removal by absorption onto activated carbon. In a study recently conducted by Liao et al., (2012) using bamboo charcoal, a maximum absorption capacity of 91.74 mg/g was calculated using an initial concentration of 200 mg/l.

Quinoline is amendable to removal by photodegradation. Using titanium oxide (TiO2) nanoparticles with an average size of 16 nm and UV irradiation at a wavelength of 365 nm, Jing et al., (2012) reported photodegradation efficiencies of up to 91.5%.

Zhang et al., (2003) have reported that degradation of quinoline by oxidation with UV/H2O2 occurs according to pseudo-first order kinetics, with the rate of photochemical degradation

increasing with increasing H2O2 concentration. In a similar study, Wang et al., (2004) reported

quinoline underwent a second-order reaction with ozone/UV and calculated rate constants at 15°C of 51 and 7.24 x10-9 M-1s-1 for reaction with ozone and hydroxyl radicals, respectively.

Quinoline is expected to volatilise very slowly from water surfaces, based on a Henry‘s Law constant of 1.67 x10-6 atm.m³/mol at 25°C (SRC, 2013). Therefore, it is not expected to be amenable to air stripping.

16. Modelling of the Routes of VOCs to Drinking Water

16.1 Introduction

Due to the limited information on the manufacturing and usage of VOCs, and their subsequent release into the environment, it has been necessary in many cases to estimate the volume of VOCs used and manufactured in the UK based on European or international data. As a result of these estimates, there are significant uncertainties in the quality of these data. As such, the quality of the data does not justify the development of complex models. Therefore, a simple, conservative model has been developed to estimate the potential concentrations of VOCs in drinking water. This model assumes that there are two routes by which VOCs can enter drinking water supply.

The first route is through the release of VOCs at sites of manufacture and use into wastewater. The wastewater containing VOCs then undergoes treatment at the wastewater treatment works and the fraction of VOCs that is not removed by treatment processes is released into surface water. Abstraction for drinking water treatment then takes place from this surface water and abstracted water is treated by either air stripping, ozonation or GAC, and any fraction of the VOCs not removed by treatment will subsequently enter the drinking water supply.

The second route is via groundwater that is subsequently abstracted and treated. This model assumes two main routes to groundwater; diffuse release from the site of manufacture/use onto land and subsequent migration to groundwater, and a point-source pollution incident onto land and migration into groundwater 50 and 100 metres from the source of the pollution incident.

It is important to note that due to the simplicity of this model, a number of potential routes in the environment have not been considered. For example, although distribution to the air is considered at multiple stages in the model, it is not considered at the manufacturing stage, and no consideration of distribution of VOCs from the air to surface water has been made (such as may be expected through precipitation and dry deposition). There is also no consideration of transport of VOCs between surface and groundwater. If more reliable data on the volumes of VOCs manufactured and used (and the processes by which they are manufactured and used) become available in the future, it may be appropriate to refine the model to include these additional considerations.

Figure 16.1 Diagrammatic representation of models predicting the occurrence of VOCs in drinking water

16.2 Input Volumes

16.2.1 1,2-Dichloropropane

No specific data for the volumes of 1,2-dichloropropane used in the UK were located; however, in 2001, the annual global production of 1,2-dichloropropane was estimated to be 350 000 tonnes, with European production accounting for 19-25% of this production (OECD, 2003). Therefore, using this tonnage and assuming a European production of 22%, the production volume in Europe is estimated to be 77 000 tonnes. The European Union calculates that the UK accounts for ca 10.5% of the total European manufacturing sector (EC, 2013), therefore, a production volume for the UK of 8085 tonnes can be estimated.

1,2-Dichloropropane is manufactured and used in closed systems and hence releases to the environment are expected to be low.

The CEFIC guidance on ―Specific Environmental Release Categories (SPERCs) Chemical Safety Assessments, Supply Chain Communication and Downstream User

Compliance1‖suggests the following release factors for ―formulation and (re)packing of substances and mixtures – Industrial (Solvent-borne)‖:

 Water solubility <1 mg/l: 0.000005%

 Water solubility 1-10 mg/l: 0.00002%

 Water solubility 10-100 mg/l: 0.0002%

 Water solubility 100-1000 mg/l: 0.002%

 Water solubility >1000 mg/l: 0.005%

Similarly, for solvents, under the description ―manufacturing of substances: industrial‖, CEFIC has suggested the following release factors2:

 Water solubility <1 mg/l: 0.00001%

 Water solubility 1-10 mg/l: 0.00003%

 Water solubility 10-100 mg/l: 0.0003%

 Water solubility 100-1000 mg/l: 0.003%

 Water solubility >1000 mg/l: 0.01%

Given the uncertainty with regards to the manufacturing and usage of 1,2-dichloropropane, it is appropriate to use the CEFIC release factor for solvents, ―manufacturing of substances: industrial‖ to provide a reasonable ‗worst-case‘ scenario. As a result, any comparisons of these outputs with toxicological endpoints should provide a precautionary risk assessment.

1,2-Dichloropropane has a reported water solubility of 2700 mg/l at 20°C, therefore, the release factor of 0.01% is applied to the model.

1 CEFIC (2012). Cefic Guidance Specific Environmental Release Categories (SPERCs). Chemical Safety Assessments, Supply Chain Communication and Downstream User Compliance. Revision 2. Available from http://www.cefic.org/Documents/IndustrySupport/REACH-Implementation/Guidance- and-Tools/SPERCs-Specific-Envirnonmental-Release-Classes.pdf

2 CEFIC (2010). SPERC Overview Table. Final for Publication, April 2010. Available from http://cefic- staging.amaze.com/Documents/IndustrySupport/-SPERC-Overview-Table-Final-for-publication- April%202010.xls

Input volume for the surface water model: 0.81 tonnes/year

16.2.2 Dichloromethane

In 2011, the UK net supply of dichloromethane was 13 568 tonnes (ONS, 2013). Therefore, this value has been used as the input tonnage for the model. The World Health Organization estimates that approximately 80% of the world-wide production of dichloromethane is released into the atmosphere, therefore, it is reasonable to anticipate that releases to water will be low.

Dichloromethane has been used for a wide range of purposes, including paint and varnish strippers, solvent degreasers, cleaning fluids and dewaxing solutions. However, a number of these uses are no longer permitted with the EU. As such, the CEFIC release factors for solvents, ―manufacturing of substances: industrial‖ may be considered appropriate for this chemical, although they may be considered highly conservative:

 Water solubility <1 mg/l: 0.00001%

 Water solubility 1-10 mg/l: 0.00003%

 Water solubility 10-100 mg/l: 0.0003%

 Water solubility 100-1000 mg/l: 0.003%

 Water solubility >1000 mg/l: 0.01%

Dichloromethane has a water solubility of 20 000 mg/l at 20°C, therefore, the release factor of 0.01% is applied to the model.

Due to the solubility of dichloromethane, it is likely that washout from the atmosphere into surface water may occur, however, this is not considered within the model.

Input volume for the surface water model: 1.36 tonnes/year

16.2.3 Aniline

The UK total production capacity of nitrobenzene in 1990 was 167 000 tonnes/year, of which 98% is used to produce aniline (WHO, 2003a). Therefore, a production volume of 155 600 tonnes/year can be estimated for aniline.

Aniline is used in closed systems, and primarily as a chemical intermediate, as such, releases to the environment are expected to be limited. However, aniline is also used in the manufacture of dyestuffs, rubber accelerators, antioxidants, and in solvents, vulcanising

agents and resins. As such, the CEFIC release factors for solvents, ―manufacturing of substances: industrial‖ may be considered appropriate for this chemical:

 Water solubility <1 mg/l: 0.00001%

 Water solubility 1-10 mg/l: 0.00003%

 Water solubility 10-100 mg/l: 0.0003%

 Water solubility 100-1000 mg/l: 0.003%

 Water solubility >1000 mg/l: 0.01%

Aniline has a water solubility 36 000 mg/l at 25°C, therefore, the release factor of 0.01% is applied to the model.

Input volume for the surface water model: 15.56 tonnes/year

16.2.4 Benzyl chloride

No data on the UK usage and production of benzyl chloride were located. However, in 2006, the annual usage of benzyl chloride in Canada was 100-1000 tonnes (Environment Canada and Health Canada 2009). The Canadian market represents 2.1% of the global manufacturing market, while the UK represents approximately 4.1% of the global manufacturing market. Therefore, assuming a volume for Canada of 500 tonnes, a UK volume of 980 tonnes can be estimated. According to the OECD, benzyl chloride is manufactured and used only in closed systems.

Benzyl chloride is widely used in the production of other substances such as plastics, dyes, lubricants, petrol, photographic developer, flavour products and pharmaceuticals. While none of the CEFIC release factors adequately describe these uses, the release factors for solvents, under the description ―manufacturing of substances: industrial‖, could be ascribed for some of these uses. These release factors are generally higher than the release factors for other categories that could also be loosely ascribed to these uses, and as such would provide a conservative assessment. Under ―manufacturing of substances: industrial‖ CEFIC has suggested the following release factors:

 Water solubility <1 mg/l: 0.00001%

 Water solubility 1-10 mg/l: 0.00003%

 Water solubility 10-100 mg/l: 0.0003%

 Water solubility 100-1000 mg/l: 0.003%

 Water solubility >1000 mg/l: 0.01%

Benzyl chloride has a reported water solubility of 525 mg/l at 25°C, therefore, for the purposes of this model, it has been assumed that 0.003% of the total tonnage will be released.

Input volume for the surface water model: 0.03 tonnes/year

16.2.5 1,3-Butadiene

According to the European Union Risk Assessment Report (RAR), the total European Union release to wastewater of 1,3-butadiene in 2002 was 1074 tonnes/year (EU, 2002a). The European Union calculates that the UK accounts for 10.5% of the total European manufacturing sector (EC, 2013), therefore, assuming an equivalent release to wastewater in the UK, a release of 113 tonnes/year can be estimated.

Input volume for the surface water model: 113 tonnes/year

16.2.6 1,1-Dichloroethane

There were no available data on the production volumes of 1,1-dichloroethane in Europe or the USA. It was reported that in 1977 in the USA, 45 500 tonnes of 1,1-dichloroethane was produced (ATSDR, 1990). The USA represents approximately 28.1% of global manufacturing while the UK represents approximately 4.1% of global manufacturing. Therefore, the UK production volume is estimated to be approximately 6640 tonnes.

1,1-Dichloroethane is primarily used as an intermediate in chemical production; however, it is also used as a solvent in paint and varnish removers, as a degreaser and cleaning agent and in ore flotation.

As such, the CEFIC release factors for solvents, ―manufacturing of substances: industrial‖ may be considered appropriate for this chemical, although they may be considered highly conservative:

 Water solubility <1 mg/l: 0.00001%

 Water solubility 1-10 mg/l: 0.00003%

 Water solubility 10-100 mg/l: 0.0003%

 Water solubility 100-1000 mg/l: 0.003%

 Water solubility >1000 mg/l: 0.01%

1,1-Dichloroethane has a water solubility of 5040 mg/l at 25°C, therefore, the release factor of 0.01% is applied to the model.

However, it should be emphasised that dichloroethane is a Red List compound, and therefore, current production levels are likely to be significantly lower than those reported in the 1970s and this input volume will be an overestimate.

Input volume for the surface water model: 0.66 tonnes/year

16.2.7 Nitrobenzene

The UK total production capacity of nitrobenzene in 1990 was 167 000 tonnes/year, of which 98% is used to produce aniline (WHO, 2003a). Nitrobenzene is primarily used in the production of aniline.

While none of the CEFIC release factors adequately describe these uses, the release factors for solvents, under the description ―manufacturing of substances: industrial‖ offer the closest description for this type of use. These release factors are generally higher than the release factors for other categories, and as such would provide a conservative assessment. Under ―manufacturing of substances: industrial‖ CEFIC has suggested the following release factors:

 Water solubility <1 mg/l: 0.00001%

 Water solubility 1-10 mg/l: 0.00003%

 Water solubility 10-100 mg/l: 0.0003%

 Water solubility 100-1000 mg/l: 0.003%

 Water solubility >1000 mg/l: 0.01%

Nitrobenzene has a reported water solubility of 2090 mg/l at 25°C, therefore, for the purposes of this model, it has been assumed that 0.01% of the total tonnage will be released.

Input volume for the surface water model: 16.7 tonnes/year

16.2.8 Oxirane methyl

In 2011, the UK net supply of oxirane methyl was reported to be 12 577 tonnes (ONS, 2013). According to the European Union Risk Assessment Report (RAR), the release of oxirane methyl to water within the EU is 0.26 kg/tonne, which is considered by the EU to be an

overestimate (EU, 2002b). Therefore, by applying this release factor to the UK net supply, an input volume for the surface water model of 3.27 tonnes/year can be estimated.

Input volume for the surface water model: 3.27 tonnes/year

16.2.9 1,2,3-Trichloropropane

According to the European Chemicals Agency (ECHA), 1,2,3-trichloropropane is only manufactured at five sites within the European Union under strictly controlled conditions, with no known production occurring within the UK (ECHA, 2011). Total releases to air and wastewater combined across the EU were estimated to be less than 1 tonne/year (precise amount not stated) (ECHA, 2011).

The European Union calculates that the UK accounts for 10.5% of the total European manufacturing sector (EC, 2013), therefore, assuming an equivalent release in the UK, a release of 0.1 tonnes/year can be estimated. However, it should be noted that as the release in Europe was reported to be the total for air and wastewater, this value is likely to be an overestimate.

Input volume for the surface water model 0.1 tonnes/year

16.2.10 Urethane

In 2009, the US National Toxicology Program (NTP) stated that there was only one urethane manufacturing site world-wide, which is based in the USA (NTP, 2011). Although polyurethane coating are used in drinking water pipes in the UK, these coatings are formed by reactions between a diisocyanate and a polyol (Forrest, 1999). No data were located that indicate that urethane molecules are used to produce polyurethanes. As such, it is considered unlikely that there are any significant routes of urethane into UK water systems.

However, it should be noted that urethane is a naturally occurring compound that is formed during the fermentation of foods such as fruit, beer, wine, yoghurt, bread and soy sauce (FAO/WHO, 2005; FSA, 2001). UK Food Standards Agency‘s Food Advisory Committee has stated that levels of urethane in foods should be reduced to the lowest technologically achievable concentrations. Therefore, while it is considered unlikely that any significant exposure to urethane will occur via drinking water, some exposure may occur via food.

Therefore, for the consideration of the small amounts likely to be released into wastewater, a nominal release volume of 0.1 tonnes/year has been assumed for this model.

Input volume for the surface water model: 0.1 tonnes/year

16.2.11 Ethylene oxide

In 2011, the UK net supply of ethylene oxide was reported to be 12 577 tonnes (ONS, 2013). In 1996, Canada produced 625 kilotonnes of ethylene oxide, resulting in a total release of 23 tonnes. Therefore, 0.04% of total production was released (Environment Canada/Health Canada, 2001). Assuming a similar level of release will occur in the UK, an input volume of 5.03 tonnes/year can be estimated.

However, it should be noted that the resulting release from this assumption is likely to be a significant over-estimation, as it was noted that release to air was the significant pathway for Canada (Environment Canada/Health Canada, 2001).

Input volume for the surface water model: 5.03 tonnes/year

16.2.12 Formaldehyde

In 1990, the UK production volume for formaldehyde was 80 000 tonnes (IPCS, 2002).

According to the US Agency for Toxic Substances and Disease Registry (ATSDR), 9642 tonnes of formaldehyde were released in the USA in 1996, with 2% (145 153 kg) of the total environmental release of formaldehyde to surface water. Production volume in the US in 1990 was 3 048 000 tons (3 078 480 tonnes), therefore, assuming a similar production in 1996, release to water would account for 0.005% of total production (ATSDR, 1999).

By applying this release factor to the UK production data, an input volume for the surface water model of 4 tonnes/year can be estimated.

Input volume for the surface water model: 4.0 tonnes/year

16.2.13 o-Toluidine

In 1998, the UK production volume for toluidine was 6000 tonnes, with approximately 90% of that production subsequently exported to the USA (WHO, 1998). o-Toluidine is used as a chemical intermediate in chemical synthesis.

 Water solubility <1 mg/l: 0.00001%

 Water solubility 1-10 mg/l: 0.00003%

 Water solubility 10-100 mg/l: 0.0003%

 Water solubility 100-1000 mg/l: 0.003%

 Water solubility >1000 mg/l: 0.01%

o-Toluidine has a reported water solubility of 15 000 mg/l at 25°C, therefore, for the purposes of this model, it has been assumed that 0.01% of the total tonnage will be released.

Input volume for the surface water model: 0.6 tonnes/year

16.2.14 Quinoline

Data on the production and usage of quinoline are extremely limited. In 1994 the estimated amount of quinoline released to the environment in the USA was 41.2 tonnes (California Environmental Protection Agency 1997). The USA represents 28.1% of global manufacturing, while the UK represents 4.1% of total global manufacturing. If it is assumed that the scale of release into the environment is analogous to the scale of manufacture, a release to the environment of 6.01 tonnes can be estimated for the UK.

It should be noted that the data from the USA have not distinguished between releases to air, water or soil. Therefore, this input volume may significantly over-estimate the release to water.

Input volume for the surface water model: 6.01 tonnes/year

16.2.15 Summary

The input volumes used in the surface water and groundwater models are provided in Table 16.1.

Table 16.1 Input volumes used in the surface and groundwater models

Input volume for the Input volume for the Input volume % Release to diffuse groundwater VOC surface water model (tonnes/year) surface water water model (tonnes/year) (tonnes/year)a

0.01 1,2-Dichloropropane 8085 0.81 8085 (CEFIC value)

0.01 Dichloromethane 13 568 1.36 13 568 (CEFIC value)

0.01 Aniline 155 600 15.56 155 600 (CEFIC value)

0.003 Benzyl chloride 980 0.03 980 (CEFIC value)

1,3-Butadiene 198 660 b 113 198 660

0.01 1,1-Dichloroethane 6640 0.066 6640 (CEFIC value)

0.01 Nitrobenzene 167 000 16.70 167 000 (CEFIC value)

Oxirane methyl 12577 b 3.27 12 577

1,2,3-Trichloropropane 0.1 100 0.1 0.1

Urethane 0.1 100 0.1 0.1

Ethylene oxide 12 577 0.04 5.03 12577

Formaldehyde 80 000 0.005 4.00 80 000

0.01 o-Toluidine 6000 0.60 6000 (CEFIC value)

Quinoline 6.01 c 6.01 6.01

a. The diffuse model assumes that the total tonnage of VOCs manufactured in the UK per annum is released into the environment to provide a ‗worst-case‘ assessment.

b. The European Union has estimated release to water for this chemical. Therefore an estimate has been made, based on the relative size of the UK manufacturing sector to the EU manufacturing sector.

c. The input volume is estimated based on reported releases to the environment in the USA.

16.3 Surface Water Model

For the ease of reading within this report, the combination of release of VOCs to wastewater, removal during wastewater treatment, discharge to surface water and abstraction and treatment at a drinking water treatment works is referred to as the ‗surface water model‘.

16.3.1 Wastewater Treatment

The removal of VOCs from the water column during wastewater treatment was modelled using data from various sources (summarised in Appendix D). These data were collated into a spread sheet which details the minimum and maximum removal that has been reported in the literature and the fate of the chemical within the treatment works (i.e. amount of the input concentration that ends up in the sludge, amount that is biodegraded and amount that stays in the water column). These removal rates were then applied to the input concentration to infer estimates of the concentration that would be released to surface water using the following formula:

*( ) ( ) ( )+

Where:

 Ceffluient is the concentration in effluent (µg/l);

 Cinfluent is the concentration in influent (µg/l);

 Sl is the % adsorbed to sludge;

 B is the % biodegraded; and

 V is the % volatilised.

A summary of the removal efficiencies by sorption to sludge, biodegradation and volatilisation during sewage treatment processes for each of the VOCs is provided in Table 16.2.

Table 16.2 Removal of VOCs during sewage treatment

Input value Removal Output value

Concentration Daily Average Average Effluent VOC entering concentration Average % % to % concentration WwTW entering biodegraded sludge volatilised (µg/l) (tonnes/year) WwTW (µg/l)1

1,2-Dichloropropane 0.81 0.234 0 0 99 0.0023

Dichloromethane 1.36 0.393 0 92 3 0.0197

Aniline 15.56 4.511 0 95 0 0.2256

Benzyl chloride 0.03 0.009 0 70 0 0.0026

1,3-Butadiene 113 32.761 0 0 96 1.3104

1,1-Dichloroethane 0.066 0.193 0 78 15 0.0135

Nitrobenzene 16.70 4.842 0 98 0 0.0968

Oxirane methyl 3.27 0.948 0 90 4 0.0569

Input value Removal Output value

Concentration Daily Average Average Effluent VOC entering concentration Average % % to % concentration WwTW entering biodegraded sludge volatilised (µg/l) (tonnes/year) WwTW (µg/l)1

1,2,3-Trichloropropane 0.1 0.029 0 0 83 0.0049

Urethane 0.1 0.029 0 90 0 0.0029

Ethylene oxide 5.03 1.459 0 90 5 0.0729

Formaldehyde 4.00 1.160 0 94 0 0.0696 o-Toluidine 0.60 0.174 0 97 0 0.0052

Quinoline 6.01 1.742 0 98 0 0.0348

1. Assuming one person produces 150 litres of waste per day (population of UK 63 million).

16.3.2 Distribution in the Environment

To estimate the distribution in the environment, a Mackay Level III Fugacity Model (US EPA/SRC EpiSuite programme version 4.1) was used to estimate the distribution between the water phase, the air, sediment and soil (US EPA/SRC, 2011).

As stated in Section 16.1, due to the limitations in the data, it has been assumed that all release to surface waters is occurring via release into sewage treatment, therefore, it has been assumed in the fugacity calculation that VOC input into the environment is entirely by release to water. The available data on the physico-chemical properties of these chemicals (as detailed in the respective section of each VOC) have been used in the fugacity model, rather than using the estimates of physico-chemical properties that would otherwise be generated by EpiSuite.

It is important to note that only the advection (i.e. distribution between air, water, soil and sediment) part of the model has been used for this project to estimate losses from water. The biodegradation part of the model has not been included within these predictions. For the purposes of this project, distribution to air, soil or sediment has been treated as a removal process, i.e. only the concentration of VOC remaining in the water column is taken forward as an output concentration. This has been calculated using the following formula:

*( ) ( ) ( )+

Where:

 CSurface is the concentration in surface water (µg/l);

 Ceffluent is the concentration in the WwTW effluent (µg/l), as detailed in Section 16.3.1;

 A is the % distributed to air;

 S is the % distributed to soil; and

 Sed is the % distributed to sediment.

No consideration of the effects of dilution in the environment has been made by the Fugacity Model. The likelihood is that, if released into surface water, there will be significant dilution of VOCs in UK rivers prior to abstraction for drinking water treatment. Additionally, no consideration has been made of abiotic degradation processes that may occur in water, for example, formaldehyde is known to be rapidly hydrolysed in water to glycol, which is not considered within this model. Therefore, in these respects, it is likely that the outputs of this aspect of the model will provide a precautionary estimate of the concentrations used for the drinking water treatment part of the model.

Table 16.3 Fugacity distribution of VOCs

Input value Fugacity distribution (%) Output value

WwTW Surface water Chemical name effluent Air Water Soil Sediment concentration concentration (µg/l) (µg/l)1

1,2-Dichloropropane 0.0023 16.9 82.6 0.0453 0.397 0.0019

Dichloromethane 0.0197 18.2 81.5 0.0208 0.24 0.0160

Aniline 0.2256 0.00649 99.5 0.00704 0.467 0.2245

Benzylchloride 0.0026 7.86 90.4 0.164 1.62 0.0023

1,3-Butadiene 1.3104 1.17 98.5 0.000129 0.342 1.2906

1,1-Dichloroethane 0.0135 18.2 81.6 0.0156 0.279 0.0110

Nitrobenzene 0.0968 1.97 96.8 0.227 1.01 0.0937

Oxirane methyl 0.0569 4.46 95.2 0.101 0.195 0.0542

1,2,3-Trichloropropane 0.0049 11.4 87.6 0.306 0.657 0.0043

Urethane 0.0029 0.00095 99.7 0.0239 0.233 0.0029

Ethylene oxide 0.0729 8.19 91.5 0.0839 0.18 0.0668

Formaldehyde 0.0696 0.00797 99.8 0.0303 0.187 0.0694

o-Toluidine 0.0052 0.0507 99.3 0.00653 0.637 0.0052

Quinoline 0.0348 0.037 95 0.0558 4.92 0.0331

1. This is the output value from Table 16.2.

16.3.3 Removal During Drinking Water Treatment

A summary of the removal efficiency of each VOC by air stripping, ozone or GAC is provided in Table 16.4. This was based on the data summarised in Appendix D. It should be noted that while air stripping is an effective treatment process for a number of VOCs (such as

1,2-dichloropropane and dichloromethane), it is not necessarily an environmentally sustainable option.

These highest and lowest magnitudes of removal were applied to the outputs of the fugacity distribution to provide indicative concentrations in finished drinking water with the most effective and least effective removal technique using the following formulae:

( )

Where:

 CDW is the concentration in drinking water (µg/l);

 Csurface is the concentration in surface water (µg/l), as detailed in Section 16.3.2; and

 is the largest [most effective removal process] or smallest [least effective removal process] value of the three removal options:

o % removed by air stripping;

o % removed by ozone; or

o % removed by GAC.

It should be noted that for the purpose of this model, each drinking water treatment technology has been considered separately, rather than in combination, i.e. no scenarios have been created using two or more of the three drinking water treatment techniques.

Table 16.4 Indicative magnitude of removal

` Input value Removal Efficiency Output value

Concentration Concentration in DW using in DW using VOC Surface water Air 4 5 most effective least effective concentration 3 Ozone GAC stripping removal removal (µg/l)2 technique technique (µg/l) (µg/l)

1,2-Dichloropropane 0.0019 >90% - >80% 0.0002 0.0019 Dichloromethane 0.0160 >90% - >90% 0.0016 0.0160 Aniline 0.2245 - >90% >90% 0.0224 0.2245 Benzyl chloride 0.0023 25% - >90%6 0.0002 0.0023 1,3-Butadiene 1.2906 >90% >90%7 >90% 0.1291 0.1291 1,1-Dichloroethane 0.0110 >90% - >90% 0.0011 0.0110

Nitrobenzene 0.0937 - <5% >90% 0.0094 0.0937

` Input value Removal Efficiency Output value

Oxirane methyl 0.0542 <10% ? - 0.0488 0.0542 1,2,3-Trichloropropane 0.0043 25% - >90% 0.0004 0.0043 Urethane 0.0029 - - - 0.0029 0.0029 Ethylene oxide 0.0668 <10% - - 0.0601 0.0668 Formaldehyde 0.0694 - - - 0.0694 0.0694 o-Toluidine 0.0052 - >90%8 >90% 0.0005 0.0052 Quinoline 0.0331 - 10% >90% 0.0033 0.0331

1. '-' indicates little or no removal anticipated.

2. This is the output value from Table 16.3.

3. Assumes a model air stripper capable of 99% removal of trichloroethylene at 25oC:

a. Water loading rate 62.25 m3/m2/h b. Gas/liquid volumetric ratio 20:1 c. 3.3 m bed of 1‖ plastic saddles

4. Assumes 1 mg/l ozone residual for 5 minutes.

5. Assumes virgin GAC with appropriate empty bed contact time. No information can be inferred from these values in relation to service life of the GAC.

6. Assumes similar adsorption properties to benzene.

7. High reaction rate assumed based on structure.

8. Assumes reaction rate similar to aniline.

16.3.4 Model Outputs

The results of the surface water model are provided in Table 16.5 and a comparison of the model with occurrence data from publically available literature is presented in Table 16.6. In a number of cases, there is a general paucity of data on the occurrence of these VOCs in water. This may imply that they do not commonly occur in water; however, it is also possible that the occurrence of some of these chemicals has not been regularly investigated.

However, where data are available for comparison, the estimates from the model are generally similar to, or lower than, the occurrence data reported in the literature. However, in cases where the model appears to be under-estimating concentrations in drinking water this

may be because the concentrations reported in the drinking water may not be ‗typical‘ levels. Comparison of model outputs to reported concentrations showed that:

 The concentrations of 1,2-dichloropropane predicted are lower than those detected in a survey from the Netherlands in 1995. However, it should be noted that the positive results only accounted for 7.7% of the total number of samples. Therefore, these measured concentrations are not representative of typical concentrations in drinking water.

 The estimated concentration of dichloromethane is consistent with typical concentrations reported in the literature.

 The estimated concentrations of aniline are consistent with the measured levels in a Spanish drinking water study.

 No data were located on the occurrence of benzyl chloride to compare with the model prediction.

 No data on the occurrence of 1,3-butadiene in drinking water were located, however, based on a comparison of the model with studies on the occurrence of 1,3-butadiene in surface water, it is likely that the predictions from the model is providing a good estimate of the risks to drinking water supplies from 1,3-butadiene.

 The concentrations of 1,1-dichloroethane estimated from the model are lower than the concentrations reported in the literature. Therefore, the model may be underestimating the risk from this chemical. However, the concentrations reported in the literature are from the 1970s and 1980s, and therefore will likely provide a poor representation of levels likely to currently be present in drinking water supplies.

 The available literature reports several instances of nitrobenzene qualitatively detected in drinking water, but no quantitative data were located. The model predicts a concentration in surface water of 0.094 µg/l. This is above the lowest reported concentration, and consistent with the reported concentrations of <1 µg/l. However, in the absence of more robust literature data, it is not possible to determine the reliability of this prediction.

 No suitable occurrence data were located to assess the predictions for oxirane methyl.

 The concentration range for 1,2,3-trichloropropane estimated from the model is lower than the detected concentrations reported in this survey. However, these measured concentrations do not appear to be representative of typical concentrations in drinking water, as it was detectable on only a few occasions.

 No suitable occurrence data were located to assess the predictions for urethane.

 No data on the concentrations of ethylene oxide in drinking water were located. The model predicts a surface water concentration lower than concentration reported in one sample at one surface water site. However, ethylene oxide was not detected in the other 861 sites examined in that survey, therefore the site with a positive result may not be representative.

 One study was located that estimated a formaldehyde concentration in drinking water of 20 µg/l, which is higher than the estimated concentrations from the model.

 No data on the concentrations of o-toluidine in drinking water were located and the available surface water data are insufficient for the purposes of comparison with the model.

 No data on the concentrations of quinoline in drinking water were located and the available surface water data are insufficient for the purposes of comparison with the model.

Table 16.5 Summary of results of the surface water model

Annual Concentration in drinking water Daily Input concentration Concentration Concentration in Concentration in DW concentration Concentration in DW using Chemical name volume entering in wastewater surface water using least effective 1 entering most effective removal (tonnes/yr) WwTW effluent (µg/l) (µg/l) removal technique WwTW (µg/l) technique (µg/l) (mg/l/year) (µg/l)

1,2-Dichloropropane 0.81 0.086 0.234 0.0023 0.0019 0.0002 0.0019

Dichloromethane 1.36 0.144 0.393 0.0197 0.0160 0.0016 0.0160

Aniline 15.56 1.647 4.511 0.2256 0.2245 0.0224 0.2245

Benzyl chloride 0.03 0.003 0.009 0.0026 0.0023 0.0002 0.0023

1,3-Butadiene 113 11.958 32.761 1.3104 1.2906 0.1291 0.1291

1,1-Dichloroethane 0.66 0.070 0.193 0.0135 0.0110 0.0011 0.0110

Nitrobenzene 16.70 1.767 4.842 0.0968 0.0937 0.0094 0.0937

Oxirane methyl 3.27 0.346 0.948 0.0569 0.0542 0.0488 0.0542

1,2,3-Trichloropropane 0.10 0.011 0.029 0.0049 0.0043 0.0004 0.0043

Urethane 0.10 0.011 0.029 0.0029 0.0029 0.0029 0.0029

Ethylene oxide 5.03 0.532 1.459 0.0729 0.0668 0.0601 0.0668

Formaldehyde 4.00 0.423 1.160 0.0696 0.0694 0.0694 0.0694 o-Toluidine 0.60 0.063 0.174 0.0052 0.0052 0.0005 0.0052

Quinoline 6.01 0.636 1.742 0.0348 0.0331 0.0033 0.0331

Table 16.6 Comparison of the surface water model output with literature data

Concentration in Concentration Drinking water DW using most in DW using concentrations Chemical name effective least effective Comments reported in the removal removal literature (µg/l) technique (µg/l) technique (µg/l)

None or the concentrations reported in the literature are from the UK, and most results are from the 1980s and 1990s.

0.01-21 The example in the literature which may be expected to be most similar to UK drinking (drinking water water is a result from the Netherlands in 1995 which positively identified 1,2-Dichloropropane 0.0002 0.0019 supplies from the 1,2-dichloropropane in 13 of 169 drinking water supplies derived from surface water at a Netherlands in concentration range of 0.06-2.1 µg/l. These concentrations are higher than those predicted 1995) by the model. However, it should be noted that the positive results only accounted for 7.7% of the total number of samples. Therefore, these measured concentrations may not be representative of typical concentrations in drinking water. 0.008-3600 None or the concentrations reported in the literature are from the UK, and most results are (worldwide drinking from the 1970s and 1980s. Most of the data report a wide range of concentrations and are Dichloromethane 0.0016 0.0160 water supplies thus difficult to compare, however, the World Health Organization report mean levels of during the 1970s dichloromethane in drinking water less than 1 µg/l. The concentrations reported are and 1980s) consistent with the lower-end of the concentrations reported in the literature.

0-0.18 Data on the occurrence of aniline in drinking water are extremely limited, however, (drinking water Aniline 0.0224 0.2245 concentrations were recently reported in Spain (2010-2011). The concentrations reported from Spain in in the model are consistent with the results from this study. 2010-11)

No data on quantifiable amounts of benzyl chloride in the environment were located. Benzyl chloride 0.0002 0.0023 No data located Therefore, an assessment of the reliability of this prediction cannot be made.

No data on the concentrations of 1,3-butadiene in drinking water were located. A concentration of 2 µg/l has been reported in surface water of a heavily industrialised river 1,3-Butadiene 0.1291 0.1291 No data located basin in San Francisco Bay. This is similar to the concentration predicted by the model of 1.3 µg/l.

0.105-24 (drinking water The concentrations estimated from the model are lower than the concentrations reported in 1,1-Dichloroethane 0.0011 0.0110 supplies primarily the literature. Therefore, the model may be underestimating the risk from this chemical in the USA during the 1980s)

The available literature reports several instances of nitrobenzene qualitatively detected in Qualitatively drinking water, but no quantitative data were located. Concentrations in fresh surface water detected in from 0.022 to <100 µg/l, with the majority of studies reporting concentrations of less than Nitrobenzene 0.0094 0.0937 drinking water, but 1 µg/l. The model predicts a concentration in surface water of 0.094 µg/l. This is above the no quantitative lowest reported concentration, and consistent with the reported concentrations of <1 µg/l. data were located However, in the absence of more robust literature data, it is not possible to determine the reliability of this prediction.

Oxirane methyl was not detected in drinking water or surface water samples reported in the literature. Three studies were reported on sewage effluent; in the first two oxirane Oxirane methyl 0.0488 0.0542 Not detected methyl was not detected (limits of detection of 5 and 20 µg/l, respectively), while in the third (USA, 1974), it was detected at 47 µg/l. The model predicts a concentration in sewage of 0.057 µg/l.

<0.24-2.654 None or the concentrations reported in the literature are from the UK, and most results in (drinking water the literature were below the limit of detection. In the biggest survey (USA), less than 0.5% 1,2,3-Trichloropropane 0.0004 0.0043 supplies in the of the 2092 domestic wells and less than 1% of the 997 private wells contained detectable USA from 1985 to levels of 1,2,3-trichloropropane. 2002)

The concentration range estimated from the model is lower than the detected concentrations reported in this survey. However, these measured concentrations do not appear to be representative of typical concentrations in drinking water.

Qualitatively detected in Only one study from the USA was located in the literature on the occurrence of urethane in Urethane 0.0029 0.0029 drinking water, but drinking water. Urethane was detected, but not quantified. No data on its occurrence in no quantitative surface water were located. data were located

No data on the concentrations of ethylene oxide in drinking water were located. One study was located in surface water that reported a mean concentration of 21 µg/l at one of 862 Ethylene oxide 0.0601 0.0668 No data located tested sites. The model predicts a surface water concentration of 0.067 µg/l. However, as ethylene oxide was not detected in the remaining 861 sites, this site may not be representative.

Most of the reported occurrence data for formaldehyde indicate that it is not detected in drinking water, however, the limits of detection for these studies are high (30 or 100 µg/l). Not detected to As a result, it is difficult to compare these data with the results of the model, as the model Formaldehyde 0.0694 0.0694 <100 µg/l predicts a concentration below these limits of detection. One study was located that estimated a concentration in drinking water of 20 µg/l, which is higher than the estimated concentrations from the model.

No data on the concentrations of o-toluidine in drinking water were located. Studies in surface water have reported concentrations of up to 20 µg/l. Most of the studies report o-Toluidine 0.0005 0.0052 No data located concentrations much lower than this (less than 1 µg/l). The model predicts a surface water concentration of 0.005 µg/l.

No data on the concentrations of quinoline in drinking water were located. Quinoline has been reported in surface water at a maximum concentration of 11 µg/l in the Netherlands Quinoline 0.0033 0.0331 No data located (1984). However, this does not appear to be a typical concentration, as many of the samples in sewage, groundwater and the marine environment were below 1 µg/l.

16.4 Groundwater Model

UK data on the use of the target VOCs are sparse. Production figures are not available for individual locations or specific to chemical processing or product manufacture. Without information on specific UK usage patterns or monitoring data to provide a preliminary assessment of existing contamination levels in UK groundwater a simple Source-Pathway- Receptor approach is used to prioritise those VOCs that due to their chemical properties or quantities being used present a greater risk to UK drinking water.

The Source-Pathway-Receptor (or Target) approach (Environment Agency, 2013) is an accepted methodology to assess risk and can be used effectively where a relative comparison is required. For the purposes of this assessment the Source is defined as the amount of a substance released (based on production or an accidental spillage) to the environment, the receptor the nominated point of impact of the contamination and the pathway the means by which it gets from the source to the receptor.

16.4.1 Approach

An approach was specifically designed for the assessment of risks to groundwater posed by the VOCs.

Source: Two methods have been used to produce a potential concentration in groundwater. In the first, EU production data was scaled to the UK, and in the second the maximum reported solubility of the VOC in water is used.

Pathway: The route for VOC pollution would initially be through soil and then through the unsaturated and saturated zone of an aquifer. To define the length of the pathway it has been assumed that there is a minimum safeguard zone between the point of use of a chemical and the receptor of 50 m (as required for an Environmental Permit). This corresponds to a groundwater Source Protection Zone 1 - SPZ1 (or Inner Protection Zone) which is equivalent to a 50 day groundwater travel time. The 50 day travel time has been used as the basis to define the degree of attenuation and level of dilution of a contaminant that would be achieved in the aquifer system. A second calculation has been made for a safeguard distance of 100 m or 100-days groundwater travel time which corresponds to SPZ2.

Receptor: the point of impact is assumed to be a drinking water supply borehole. The modelled predicted concentrations assume that there is no treatment and potentially provide a worst case assessment due to the high concentrations assumed to be spilt at the source and short travel times modelled.

16.4.2 Methodology

Two approaches have been used in the assessment of risk. In the first it is assumed that the target VOCs may reach groundwater as a diffuse source of pollution which would be typical of

the situation if they were in widespread usage across the UK and secondly that as a result of an accidental leak or spill the contamination arises as a point source of pollution.

This section outlines the main parameters and assumptions used within these two scenarios. For the purpose of the assessment a number of generic values have been used to represent UK soil and aquifer conditions.

Diffuse source

To provide a worst case assessment it has been assumed that the total tonnage of VOCs manufactured within the UK per annum is released and is diluted by the amount of available rainfall minus evapotranspiration (67 003 005 053 300 m³/year).

An ‗urbanisation‘ factor has been applied to reflect the fact that use of these chemicals is likely to be restricted to urban areas. This urban factor increases the concentration of the diffuse source to provide a better representation of where these chemicals are primarily manufactured and used (i.e. developed land) by reducing the influence, and hence potential dilution, of rainfall from non-urban areas. The urban factor was assumed to be 6.82% for the UK (0.0682 in the equation below) (UK NEA, 2011).

The diffuse source inputs are estimated using the general equation:

As a starting point, using the relative density of the VOC compounds, a first assessment of the capacity of the VOCs to enter the groundwater system can be made.

A VOC compound with a relative density of less than 1 (g/cm3) is unlikely to pollute groundwater over a significant area due to the fact that the VOC compound will effectively ―float‖ in the unsaturated zone. The capacity for degradation of organic compounds in this zone is higher than in the saturated zone of the aquifer as the aerobic conditions will favour highly efficient degrading organisms and increased air-water phase mixing. The half-lives of chemicals in an aerobic environment are therefore considerably shorter than in anaerobic conditions.

Dilution Factor and Attenuation Factor

The two most common mechanisms for the reduction in contaminant concentrations in groundwater are dilution and attenuation. Dilution occurs through dispersion of a contaminant in the groundwater and attenuation based on the physico-chemical characteristic of the VOC.

Dilution Factor

The dilution factor (DF) is a function of the distance between the source and the receptor and is also based on the characteristics of the aquifer. In this study, a DF of 100 has been used in the risk assessment.

Attenuation Factor

The Soil Association estimated in 2009 that the average UK soil organic content was around 2% in the National Soil Inventory survey (Bellamy et al., 2005) and Countryside Survey (NERC, 2007).

Generic soil parameters are presented in Table 16.7 and are representative of a ―typical soil‖. Their influence in the resulting concentration is low compared to the soil water partition

coefficient (Kd) which is derived from the VOC organic carbon partition coefficient multiplied

by the organic content of the soil (foc).

Table 16.7 Typical physical soil parameters

Parameter Value Unit

Water content (θw) 0.18 Fraction

Air content (θa) 0.006 Fraction Bulk density (ρ) 2.36 g/cm3

Organic carbon (foc) 0.02 Fraction

The equation for the attenuation factor (AF) uses the equation for the calculation of pore water quality based on soil concentration from the Environment Agency Remedial Target Methodology (EA, 2006). The AF reflect the addition of the processes of sorption, where solute become sorbed following reaction with organic carbon or clay minerals, and retardation, which reduces the rate of contaminant migration. The AF represents the ratio between the concentration in the soil (in mg/kg) and the concentration in the pore water equals to 1 mg/l. The resulting dimensionless AF is calculated as:

[ ]

Where:

The resulting concentration at the receptor is the source concentration divided by both the dilution and attenuation factors.

Point source

In the prediction of risk posed by a point source, it has been assumed that a constant source ―leak or spill‖ occurred at the manufacturing plant or during handling of the product in downstream manufacturing and that the concentration of individual VOC that could enter the groundwater system would equate to the maximum solubility for that compound in water.

The Remedial Targets Methodology (Environment Agency, 2006) provides a groundwater flow equation between a source and a receptor. This approach has been used to predict VOC concentrations at the receptor using both a 50-day travel time (50 m) and 100-day travel time (100 m). Table 16.8 provides a list of generic aquifer characteristics that have been assumed to complete the modelling exercise.

The critical assumptions that have been used which will influence predicted VOCs concentrations at the receptor include:

 That the system is in steady state.

 Use of the Ogata Banks flow equation which simulates dispersion in two dimensions (vertical and horizontal).

 That published degradation rates applies to all phases of the pollutants (liquid and gas in this case).

 Dispersivities (longitudinal, transversal and vertical) have been taken as respectively 10%, 1% and 0.1% of the pathway length.

Table 16.8 Remedial targets methodology parameters

Parameter Value Unit

Initial contaminant concentration in groundwater at plume core (C0) Maximum solubility mg/l

Half-life for degradation of contaminant in groundwater (t1/2) VOC Specific days Width of plume in aquifer at source (perpendicular to flow) (Sz) 1.00 m

Plume thickness at source (Sy) 1.00 m

Saturated aquifer thickness (da) 10 m

Bulk density of aquifer materials () 2.36 g/cm3

Effective porosity of aquifer (n) 0.05 fraction

Hydraulic gradient (i) 0.05 fraction

Parameter Value Unit

Hydraulic conductivity of aquifer (K) 1.00 m/d

Distance to compliance point (x) 50 or 100 m

Time since pollutant entered groundwater (t) 1x10100 days

Partition coefficient (Kd) VOC Specific l/kg Longitudinal dispersivity (ax) 5.00 m

Transverse dispersivity (az) 0.5 m

Vertical dispersivity (ay) 0.05 m

Groundwater flow velocity (calculated) (v) 1.00 m/d

It should be noted that as this methodology calculates the concentration 50 and 100 metres from the source of a spill based on the water solubility, rather than the tonnage, it provides a ‗worst-case‘ scenario. Therefore, although it has been considered unlikely that either 1,2,3-trichloropropane or urethane will be released into the environment in the UK, predictions are included for these chemicals that will suggest the potential for significant concentrations to reach groundwater in the event of a pollution incident involving these chemicals due to their high water solubility.

Results

The assessment of VOC diffuse and point source pollution can be used to assess the relative risks from each compound versus the rest and to identify the overall risk of finding detectable levels of VOCs in UK water supply boreholes.

Diffuse Source Methodology Results

Table 16.9 provides the results of an assessment into the risk of diffuse pollution causing groundwater pollution. It should be noted that this is very much a worst case approach as it is assumed that the available source term (total concentration diluted by rainfall) is released onto urbanised land regardless of actual use.

The table shows the predicted VOC receptor concentration after attenuation and dilution factors. Receptor concentration for 1,3-butadiene, oxirane methyl, urethane, ethylene oxide, formaldehyde, o-toluidine have been calculated using the same methodology but it should be highlighted that it is unlikely that these VOCs would enter the groundwater system.

Table 16.9 Diffuse source methodology results

Volume "Source” Enter GW Partition Concentration at Chemical name AF2 (tonnes) (mg/l)1 system? coefficient (Kd) receptor (µg/l)3

1,2-Dichloropropane 8085 1.768 x10-3 Yes 1.36 1.44 1.23 x10-2

Dichloromethane 13568 2.968 x10-3 Yes 0.156 0.23 1.28 x10-1

Aniline 155600 3.404 x10-2 Yes 8.2 8.28 4.11 x10-2

Benzyl chloride 980 2.144 x10 -4 Yes 13.836 13.91 1.54 x10-4

1,3-Butadiene 198660 4.345 x10-2 No 3.015 3.10 1.40 x10-1

1,1-Dichloroethane 6640 1.452 x10-3 Yes 0.576 0.65 2.22 x10-2

Nitrobenzene 167000 3.653 x10-2 Yes 0.726 0.80 4.55 x10-1

Oxirane methyl 12577 2.751 x10-3 No 0.046 0.12 2.25 x10-1

1,2,3-Trichloropropane 0.1 2.187 x10-8 Yes 1.72 1.80 1.22 x10-7

Urethane 0.1 2.187 x10-8 No 0.4 0.48 4.59 x10-7

Ethylene oxide 12577 2.751x10-3 No 0.32 0.40 6.94 x10-2

Formaldehyde 80000 1.750x10-2 No 0.422 0.50 3.51 x10-1 o-Toluidine 6000 1.312 x10-3 No 2.28 2.36 5.57 x10-3

Quinoline 6.01 1.315 x10-6 Yes 0.1374 0.21 6.15 x10-5

2. * +

Point Source Methodology Results

Table 16.10 shows the results of the point source pollution assessment with concentrations (in µg/l) at 50 m and at 100 m using the Ogata Banks equation alongside half-life values for each VOC in groundwater taken from reported literature.

Table 16.10 Point source methodology results

Maximum Half-life Literature Concentration (µg/l) Concentration (µg/l) VOC concentration (i.e. (days) Source at 50 m at 100 m solubility) (mg/l)

ATSDR, 1,2-Dichloropropane 2700 456.25 712.21 14.64 1989

ATSDR, Dichloromethane 20 000 540 119707.19 18885.24 2000

SCTEE, Aniline 36 000 2007.5 3640.79 44.83 2003

CDC, Benzyl chloride 525 0.60 8.62 x10-262 1.00 x10-297 1978

US EPA, 1,3-Butadiene 735 28 4.11 x10-13 2.98 x10-21 2010

Lawrence, 1,1-Dichloroethane 5040 114 201.13 1.54 2006

WHO, Nitrobenzene 2090 133 62.35 0.41 2003a

ECHA, Oxirane methyl 590 000 22 151752.13 3077.20 2002

Stepek, 1,2,3-Trichloropropane 1750 547.5 399.80 7.60 2009

HSDB, Urethane 480 000 1.0 x1099 4787651.14 1205138.27 2013

WHO, Ethylene oxide 1 000 000 11.2 0.17 4.89 x10-6 2003b

WHO, Formaldehyde 475 000 8 3.93 x10-5 4.28 x10-11 2003c

OECD, o-Toluidine 15 1.0 x1099 149.61 37.66 2004

HSDB, Quinoline 60 000 160 144678.06 11435.60 2013

Figure 16.2 presents result for each VOC from both methodologies. It presents high concentration around the edge (1 g/l – 1 gram per litre) of the chart and low concentration to the centre (1 ag/l – 1 attogram per litre) When concentrations are lower than 1 x10-15 mg/l or where production estimates are missing, results are not plotted.

Figure 16.2 Groundwater risk assessment results

16.4.3 Evaluation of Predictions

Results of the assessment using the diffuse source methodology shows most VOCs are predicted to be in concentrations of less than 1 µg/l. Occasional point sources, e.g. accidental spillage, leak from manufacturer or transport incident, may raise groundwater concentration to levels identified in the point source methodology SPZ1 or SPZ2.

16.4.4 Removal during Drinking Water Treatment

Using the data derived for the removal of VOCs by drinking water treatment described in Section 16.3.3, predictions for the concentrations of VOCs in treated water abstracted from groundwater can be derived using the following equations:

( )

Where:

 CDW is the concentration in drinking water (µg/l);

 Creceptor is the concentration at the receptor (µg/l), as detailed in Table 16.9; and

 is the largest [most effective removal process] or smallest [least effective removal process] value of the three removal options:

o % removed by air stripping;

o % removed by ozone; or

o % removed by GAC.

The estimates of the concentrations of VOCs in drinking water derived from groundwater following diffuse pollution are provided in Table 16.11. These diffuse pollution estimates are subsequently used, in combination with the estimates for concentrations in surface water to predict the potential risk to human health by comparison of exposure with Tolerable Daily Intakes (TDI).

The estimates of the concentrations of VOCs in drinking water derived from groundwater following point sources of pollution are provided in Table 16.12 and Table 16.13 (for 50 and 100 metres from the point source of pollution, respectively). These point source estimates are not considered in comparison with TDIs, as they are concentrations that would only be anticipated in the event of a chemical spill, rather than being representative of typical human exposure.

Table 16.11 Estimated concentrations in drinking water using the groundwater diffuse pollution model

Estimated concentration in Estimated concentration in Concentration at drinking water using the least drinking water using the VOC receptor (µg/l)1 effective removal technique most efficient removal (µg/l)2 technique (µg/l)3

1,2-Dichloropropane 1.23 x10-2 1.23 x10-2 1.23 x10-3

Dichloromethane 1.28 x10-1 1.28 x10-1 1.28 x10-2

Aniline 4.11 x10-2 4.11 x10-2 4.11 x10-3

Benzyl chloride 1.54 x10-4 1.54 x10-4 1.54 x10-5

1,3-Butadiene 1.40 x10-1 1.40 x10-2 1.40 x10-2

1,1-Dichloroethane 2.22 x10-2 2.22 x10-2 2.22 x10-3

Nitrobenzene 4.55 x10-1 4.55 x10-1 4.55 x10-2

Oxirane methyl 2.25 x10-1 2.25 x10-1 2.02 x10-1

1,2,3-Trichloropropane 1.22 x10-7 1.22 x10-7 1.22 x10-8

Urethane 4.59 x10-7 4.59 x10-7 4.59 x10-7

Ethylene oxide 6.94 x10-2 6.94 x10-2 6.25 x10-2

Formaldehyde 3.51 x10-1 3.51 x10-1 3.51 x10-1 o-Toluidine 5.57 x10-3 5.57 x10-3 5.57 x10-4

Quinoline 6.15 x10-5 6.15 x10-5 6.15 x10-6

1. See Table 16.9.

2. Based on the smallest [least effective removal process] value of the three removal options of removal by air stripping, removal by ozone, or removal by GAC.

3. Based on the largest [most effective removal process] value of the three removal options of removal by air stripping, removal by ozone, or removal by GAC.

Table 16.12 Estimated concentrations in drinking water using the groundwater point source pollution model (at 50 metres)

Estimated concentration in Estimated concentration in Concentration at drinking water using the drinking water using the most VOC 50 m (µg/l)1 least effective removal efficient removal technique technique (µg/l)2 (µg/l)3

1,2-Dichloropropane 712.21 712.2 71.2

Dichloromethane 119707.19 119707 11971

Aniline 3640.79 3640.8 364.1

Benzyl chloride 8.62 x10-262 8.6 x10-262 8.6 x10-263

1,3-Butadiene 4.11 x10-13 4.1 X10-14 4.1 x10-14

1,1-Dichloroethane 201.13 201.1 20.1

Nitrobenzene 62.35 62.3 6.2

Oxirane methyl 151752.13 151752.1 136576.9

1,2,3-Trichloropropane 399.80 399.8 40.0

Urethane 4787651.14 4787651.1 4787651.1

Ethylene oxide 0.17 0.17 0.15

Formaldehyde 3.93 x10-5 3.9 x10-5 3.9 x10-5 o-Toluidine 149.61 149.6 15.0

Quinoline 144678.06 144678.1 14467.8

1. See Table 16.10.

2. Based on the smallest [least effective removal process] value of the three removal options of removal by air stripping, removal by ozone, or removal by GAC.

3. Based on the largest [most effective removal process] value of the three removal options of removal by air stripping, removal by ozone, or removal by GAC.

Table 16.13 Estimated concentrations in drinking water using the groundwater point source pollution model (at 100 metres)

Estimated concentration Estimated concentration in Concentration at in drinking water using drinking water using the VOC 100 m (µg/l)1 the least effective removal most efficient removal technique (µg/l)2 technique (µg/l)3

1,2-Dichloropropane 14.64 14.6 1.5

Dichloromethane 18885.24 18885.24 1888.52

Aniline 44.83 44.8 4.5

Benzyl chloride 1.00 x10-297 1.0 x10-297 1.0 x10-298

1,3-Butadiene 2.98 x10-21 3.0 x10-22 3.0 x10-22

1,1-Dichloroethane 1.54 1.5 0.2

Nitrobenzene 0.41 0.4 0.04

Oxirane methyl 3077.20 3077.2 2769.5

1,2,3-Trichloropropane 7.60 7.6 0.8

Urethane 1205138.27 1205138.3 1205138.3

Ethylene oxide 4.89 x10-6 4.9 x10-6 4.4 x10-6

Formaldehyde 4.28 x10-11 4.3 x10-11 4.3 x10-11 o-Toluidine 37.66 37.7 3.8

Quinoline 11435.60 11435.6 1143.6

1. See Table 16.10.

2. Based on the smallest [least effective removal process] value of the three removal options of removal by air stripping, removal by ozone, or removal by GAC.

3. Based on the largest [most effective removal process] value of the three removal options of removal by air stripping, removal by ozone, or removal by GAC.

16.5 Surface Water-Groundwater Interactions

Due to the limitations in the input data, groundwater and surface water have been considered as separate models with no interaction between them (i.e. no flow from groundwater to surface water or vice versa. This may be a significant limitation if it is considered that a significant proportion of flow from a river is from a groundwater source.

Therefore, to provide some assessment of such circumstances, the concentrations of VOCs in surface water and in the groundwater (diffuse source) models can be combined and the removal of VOCs during drinking water treatment from this combined concentration assessed. This assessment is provided in Table 16.14.

In this assessment, it has been assumed that there is an equal ‗blending‘ of surface and groundwater concentrations of VOCs to provide the maximum concentration in blended water (i.e. the concentration in surface water has been added to the concentration in groundwater to give a total concentration). This provides an extreme ‗worst-case‘ scenario; as if a truly equal blending of different waters were to occur, the concentration would be expected to be an average of the two sources (surface and groundwater), rather than an additive concentration. However, this extreme approach will mean that subsequent comparisons with exposure data will be highly precautionary. Therefore, if the subsequent exposure assessment indicates that there is unlikely to be a concern to human health, it is reasonable to assume that under more realistic conditions (where concentrations in the water will be even lower), that exposure to these VOCs would be even less likely to be of concern to human health.

Table 16.14 Estimated concentrations in drinking water from a combined surface water-groundwater (diffuse source) model

Concentration in drinking water

Concentration in Estimated Estimated Concentration “Blended” water groundwater concentration in concentration in VOC in surface concentration (diffuse source) drinking water drinking water water (µg/l) (µg/l)1 (µg/l) using the least using the most effective removal effective removal technique (µg/l)2 technique (µg/l)

1,2-Dichloropropane 0.0019 0.0123 0.0142 0.0142 0.0014

Dichloromethane 0.0160 0.128 0.144 0.144 0.0144

Aniline 0.2245 0.0411 0.2656 0.2656 0.0266

Benzyl chloride 0.0023 1.54 x10-4 0.0025 0.0025 0.0002

1,3-Butadiene 1.2906 0.14 1.4306 0.1431 0.1431

1,1-Dichloroethane 0.0110 0.0222 0.0332 0.0332 0.0033

Nitrobenzene 0.0937 0.455 0.5490 0.5490 0.0549

Oxirane methyl 0.0542 0.225 0.2792 0.2792 0.2512

1,2,3-Trichloropropane 0.0043 1.22 x10-7 0.0043 0.0043 0.0004

Urethane 0.0029 4.59 x10-7 0.0029 0.0029 0.0029

Ethylene oxide 0.0668 0.0694 0.1362 0.1362 0.1226

Formaldehyde 0.0694 0.3512 0.4206 0.4206 0.4206 o-Toluidine 0.0052 0.00557 0.0108 0.0108 0.0011

Quinoline 0.0331 6.15 x10-5 0.0332 0.0332 0.0033

1. Sum concentration of the concentration in surface water and the concentration in groundwater (diffuse source).

2. These concentrations are subsequently used for comparison with Tolerable Daily Intakes for the 14 VOCs (see Section 17.4).

16.6 Limitations within the Estimations

Several limitations have been identified and discussed within the preceding sections. The most significant of these limitations is the lack of robust information on the current production and use of these VOCs within the UK. Where data have been available on the volumes of these chemicals within the UK, it has largely been for an overall volume. No information was located on the number of sites at which these chemicals may be produced and/or used, the different processes this volume may undergo or the potential for release of chemicals to the environment for each of these processes.

In a number of cases, no information on the volumes used or produced in the UK were located. As such, estimations have been made, based on data from other countries. However, this approach firstly assumes that these chemicals are being produced in the UK, and secondly that the production volumes correlate with the relative sizes of the manufacturing industries of these countries.

As a result of this lack of data, broad assumptions have also been made regarding the volumes of these chemicals that will be released into the sewer system. A default release factor of 0.01% has been applied in a number of instances, which is based on information derived by CEFIC. These release factors are considered to be conservative, and as such represent a reasonable ‗worst-case‘ scenario.

Where data are available for comparison, the estimates from the model are generally similar to, or lower than, the occurrence data reported in the literature. In some cases, the model appears to be under-estimating concentrations in drinking water, although it should be emphasised that in a number of these cases, the concentrations reported in water may not be representative of ‗typical‘ levels. Therefore, while some validation of the predictions is possible, additional data are needed to validate the models for all chemicals with any degree of confidence.

It should also be noted that although at multiple stages in the surface water model calculations are made to account for migration of VOCs from water to the atmosphere (e.g. volatilisation, air stripping), no account is made of the migration from the atmosphere to water (e.g. precipitation, dry deposition). The model also assumes that the wastewater treatment works discharge to inland surface waters, whereas, in some instances discharge may be to marine waters or inland waters downstream of any drinking water abstraction points. In such cases, there would be no pathway to drinking water, and therefore, the model may overestimate the risks.

Similarly, although surface water and groundwater are considered as separate compartments, with no interaction between them, this would only be a significant limitation if there was a

significant contribution to the flow of a river from a groundwater source. This limitation is addressed, to a limited degree, by combining the results of the surface water model with the diffuse source model. However, the interactions between groundwater and surface water are likely to be more complex than is implied by the simple combination of these two models.

It should also be noted that the drinking water treatment aspect of the model only considers one treatment process (air stripping, ozone or GAC). This is not a comprehensive list of treatment processes, and it is possible that many treatment works will use a combination of processes, which may further reduce the concentration in treated drinking water.

17. Exposure Modelling

Consideration has been given in this project to the choice of the most appropriate method to compare exposure of the fourteen VOCs, not only via drinking water, but also via bathing and showering. Ideally, the concentrations of VOCs in finished drinking water estimated from the model described in Section 16 would be used in a second model to estimate the conditions present within a bathroom during a bathing or showering event to calculate exposure via the dermal and inhalation routes. These data would then be compared with tolerable intakes for these routes.

However, as discussed in Section 16, a number of assumptions have been made in developing the drinking water model, most notably with regards to the input data. While this has resulted in reasonable predictions of the concentrations of VOCs in drinking water, it has not been possible to fully validate these data. A model that estimates exposure during bathing and showering would also require the use of a significant number of precautionary assumptions, which would compound existing uncertainty within the data to produce outputs that are unlikely to be robust. As a result, such predictions are likely to over-estimate exposure, while the use of large uncertainty factors in the toxicological data will result in relatively small tolerable intakes. This may result in overly-precautionary results that indicate a potential for health concern due to uncertainty in the data, rather than genuine health concerns. In addition, the development of such a detailed model requires Tolerable Daily Intakes (TDIs) for the oral, dermal and inhalation routes to allow any meaningful assessment to take place. However, there are a number of VOCs where it has not been possible to derive dermal TDIs and as a result an assessment using this approach would be incomplete with the currently available data.

Therefore, it is considered more appropriate to assess multi-route exposure on the basis of the relative contribution of each route compared to consumption of water in terms of litre- equivalents per day. This approach has previously been considered by the World Health Organization (WHO) in the assessment of chloroform consumption, and by Health Canada in consideration of several of their Guidelines for Canadian Drinking Water Quality (GCDWQ).

For the purposes of this assessment, it has been assumed that a 60 kg adult will drink 2 litres of water per day and take either one shower or one bath of 30 minutes duration each day. This is considered to be a conservative, but not unrealistic scenario.

As a result of this approach, the tolerable doses derived for dermal and inhalation exposures are not required for the model. However, they are included in this report, for information, but also so that if additional data become available in the future that enable a more detailed model to be developed, these data are readily available to assist in such assessment.

17.1 Dermal Exposure Model

Krishnan (2004) and Krishnan and Carrier (2008) have developed a two-tiered approach to assessing dermal exposure.

The first tier determines whether dermal exposure contributes a significantly to total systemic exposure compared to oral exposure (considered to be >10% of a 1.5 litre daily intake)3. This assessment is made on the basis of the skin permeability co-efficient (Kp); if the Kp is greater than 0.024 cm/hour, absorption via the dermal route is considered to be significant.

In England and Wales, it is assumed that the average adult consumes 2 litres of water per day. Using this assumption, a Kp of 0.032 cm/hour can be calculated as a threshold value for significant exposure (>10% absorption).

If absorption is significant, a chemical undergoes the second tier of assessment. The second tier provides an estimate of dermal exposure as a litre-equivalent using the following equation:

Where:

 Kp is the skin permeability co-efficient (cm/hour);

 t is the duration of the shower or the bath (assumed to be 0.5 hours);

 Fabs is the fraction of the absorbed dose (assumed to be 0.7);

 A is the area of the skin exposed (assumed to be 18 000 cm² for adults;

 Cf is a factor to convert from cm² to litres (0.001)

Kp is calculated using the following equation (Bogen, 1994; WHO, 2000):

Where:

 MW is the molecular weight;

 Log Kow is the log of the octanol-water partition co-efficient

3 The average daily water intake in an adult in Canada is assumed to be 1.5 litres of water per day.

17.1.1 Dermal Exposure Results

The results of the Tier 1 assessment are provided in Table 17.1. These data indicate that, with the exception of urethane, dermal absorption is expected to be a significant route of exposure for all the VOCs. Therefore a Tier 2 assessment was conducted on all the VOCs except urethane. These data are presented in Table 17.2.

Table 17.1 Determination of skin permeability co-efficients for the fourteen VOCs.

Dermal Is dermal Molecular penetration absorption VOC log Kow Log Kp weight coefficient, significant? (i.e. is Kp (cm/h) Kp >0.032 cm/h?)

1,2-Dichloropropane 1.99 112.986 -0.7612144 0.173 Yes

Dichloromethane 1.3 84.9328 -0.89450112 0.127 Yes

Aniline 0.9 93.1283 -1.22613432 0.059 Yes

Benzyl chloride 2.3 126.5853 -0.71168712 0.194 Yes

1,3-Butadiene 1.99 54.0914 -0.14871056 0.710 Yes

1,1-Dichloroethane 1.79 98.9596 -0.73853984 0.183 Yes

Nitrobenzene 1.85 123.11 -0.952744 0.111 Yes

Oxirane methyl 0.03 58.0794 -1.39754576 0.040 Yes

1,2,3-Trichloropropane 2.27 147.432 -0.9469728 0.113 Yes

Urethane -0.15 89.0933 -1.83097032 0.015 No

Ethylene oxide -0.22 44.0526 -1.40566704 0.039 Yes

Formaldehyde 0.35 30.0258 -0.90866832 0.123 Yes o-Toluidine 1.4 107.155 -1.064012 0.086 Yes

Quinoline 1.88 129.1613 -0.99719752 0.101 Yes

Table 17.2 Dermal exposure to VOCs as a daily litre-equivalent

VOC Kp Dermal exposure (L-eq)

1,2-Dichloropropane 0.173 1.092

Dichloromethane 0.127 0.803

Aniline 0.059 0.374

Benzyl chloride 0.194 1.224

1,3-Butadiene 0.710 4.473

1,1-Dichloroethane 0.183 1.150

Nitrobenzene 0.111 0.702

VOC Kp Dermal exposure (L-eq)

Oxirane methyl 0.040 0.252

1,2,3-Trichloropropane 0.113 0.712

Ethylene oxide 0.039 0.248

Formaldehyde 0.123 0.777

o-Toluidine 0.086 0.544

Quinoline 0.101 0.634

17.2 Inhalation Exposure Model

Krishnan (2004) and Krishnan and Carrier (2008) have developed a two-tiered approach to assessing inhalation exposure.

The first tier determines whether inhalation exposure contributes a significant route of exposure compared to oral exposure (considered to be >10% of a 1.5 litre daily intake)4. This

assessment is made on the basis of the Fair:water ratio (the ratio of air-to-water VOC

concentration); if the Fair:water is greater than 0.00089, absorption via the inhalation route is considered to be significant.

In England and Wales, it is assumed that the average adult consumes 2 litres of water per

day. Using this assumption, a Fair:water of 0.001 can be calculated as a threshold value for significant exposure (>10% absorption).

If absorption is significant, a chemical undergoes the second tier of assessment. The second tier provides an estimate of inhalation exposure as a litre-equivalent using the following equation:

Where:

 Fair:water is the ratio of air-to-water VOC concentration;

 Qalv is the adult alveolar ventilation rate (assumed to be 675 l/hour);

 t is the duration of the shower or the bath (assumed to be 0.5 hours);

 Fabs is the fraction of the absorbed dose (assumed to be 0.7);

Fair:water is calculated using the following equation (Krishnan, 2004):

4 The average daily water intake in an adult in Canada is assumed to be 1.5 litres of water per day.

Where:

 Teff is the transfer efficiency of the VOC

 Kaw is the unitless Henry‘s Law constant at 25°C

 80.25 is the ratio of the volume of air in the average bathroom (6420 litres) to the average volume of water (80 litres) used whilst bathing or showering according to Krishnan (2004).

The transfer efficiency for a VOC should ideally be measured data, however, there are a general paucity of these data for the 14 VOCs examined in this report. Therefore, for many of the VOCs, the transfer efficiency from water to air has been calculated using the following formula (US EPA, 2000):

This formula appears to offer a good prediction for the transfer efficiency for chemicals within its domain range. For example, a transfer efficiency of 0.59 (i.e. 59%) has been reported for dichloromethane (Hamlin et al., 2009), while using this formula, a transfer efficiency of 0.53 (i.e. 53%) is estimated.

However, it should be noted that this formula is less appropriate for those chemicals with a dimensionless Henry‘s Law constant of <0.001, as this appears to have been to lowest experimental Henry‘s Law constant used in the derivation of the model upon which this formula is based (i.e. values below 0.001 are outside the domain range) In such cases, an arbitrary assumption of a 10% transfer efficiency has been applied to the chemical.

17.2.1 Inhalation Exposure Results

The results of the Tier 1 assessment are provided in Table 17.3. These data indicate that, only seven of the fourteen VOCs are likely to undergo any significant volatilisation from water to air during bathing and showering. Therefore a Tier 2 assessment was only conducted on the following VOCs: ,2-dichloropropane; dichloromethane; benzyl chloride; 1,3-butadiene; 1,1-dichloroethane; 1,2,3-trichloropropane and ethylene oxide . These data are presented in Table 17.4.

Table 17.3 Determination of the significance of volatilisation of VOCs during bathing and showering

Transfer Kaw Is transfer to air efficiency (dimensionless significant? (i.e. VOC Reference Fair:water from water to Henry's Law is Fair:water air constant) > 0.0012?)

1,2-Dichloropropane 0.52 estimated 0.115 0.00584 Yes

Dichloromethane 0.59 Hamelin et al., 2009 0.133 0.00672 Yes

Aniline 0.10 arbitrary value 0.0000826 0.00001 No

Benzyl chloride 0.38 estimated 0.0168 0.00269 Yes

1,3-Butadiene 0.76 estimated 3.01 0.00949 Yes

assuming the transfer efficiency is the same 1,1-Dichloroethane 0.61 as 1,2-dichloroethane 0.231 0.00721 Yes (McKone and Knezovich, 1991)

Nitrobenzene 0.16 estimated 0.000981 0.00015 No

Oxirane methyl 0.24 estimated 0.00285 0.00056 No

1,2,3-Trichloropropane 0.36 estimated 0.014 0.00239 Yes

Urethane 0.10 arbitrary value 0.00000263 0.000000 No

Ethylene oxide 0.30 estimated 0.00605 0.00122 Yes

Formaldehyde 0.10 arbitrary value 0.0000138 0.00000 No o-Toluidine 0.10 arbitrary value 0.000081 0.00001 No

Quinoline 0.10 arbitrary value 0.0000683 0.00001 No

Table 17.4 Inhalation exposure to VOCs as a daily litre-equivalent

VOC Fair:water Inhalation L-eq

1,2-Dichloropropane 0.005844 1.38

Dichloromethane 0.006722 1.59

Benzyl chloride 0.002687 0.63

1,3-Butadiene 0.009489 2.24

1,1-Dichloroethane 0.007212 1.70

1,2,3-Trichloropropane 0.002386 0.56

Ethylene oxide 0.001217 0.29

17.3 Multi-route Exposure

Using the litre-equivalent values calculated for dermal and inhalation exposure described in Sections 17.1 and 17.2 and assuming a typical drinking water consumption of 2 litres per day, a conservative total exposure can be calculated for each VOC. These data are presented in Table 17.5.

Table 17.5 Equivalent total daily intake of the fourteen VOCs following oral consumption of drinking water and one 30-minute bath or shower

Drinking Water Dermal Intake Inhalation Intake Total Intake VOC Intake (l/day) (l-eq/day) (l-eq/day) (l-eq/day)

1,2-Dichloropropane 2.00 1.09 1.38 4.47

Dichloromethane 2.00 0.80 1.59 4.39

Not a significant Aniline 2.00 0.37 2.37 source of exposure

Benzyl chloride 2.00 1.22 0.63 3.86

1,3-Butadiene 2.00 4.47 2.24 8.72

1,1-Dichloroethane 2.00 1.15 1.70 4.85

Not a significant Nitrobenzene 2.00 0.70 2.70 source of exposure

Not a significant Oxirane methyl 2.00 0.25 2.25 source of exposure

1,2,3- 2.00 0.71 0.56 3.28 Trichloropropane

Not a significant Not a significant Urethane 2.00 2.00 source of exposure source of exposure

Ethylene oxide 2.00 0.25 0.29 2.54

Not a significant Formaldehyde 2.00 0.78 2.78 source of exposure

Not a significant o-Toluidine 2.00 0.54 2.54 source of exposure

Not a significant Quinoline 2.00 0.63 2.63 source of exposure

17.4 Comparison of Exposure Data with Tolerable Daily Intakes

The estimated daily litre equivalent intakes for each VOC can be combined with the outputs of the model described in Section 16 to determine an estimated daily dose using the following formula:

Where:

 EDD is the estimated daily dose of a VOC (µg/kg bw/day);

 Dw is estimated drinking water concentration of the VOC according to the Combined Surface Water-Ground Water (Diffuse Source) Model (µg/l);

 leq is the total intake (i.e. the combined oral, dermal and inhalation intake) for a VOC (litre equivalents);

 60 is the assumed adult bodyweight (kg)

This estimated daily dose can then be compared with the Tolerable Daily Intake (TDI) for that specific VOC to determine the Risk Characterisation Ratio (RCR).

Where:

 EDD is the estimated daily dose of a VOC (µg/kg bw/day);

 TDI is the Tolerable Daily Intake of a VOC (µg/kg bw/day).

An RCR of less than 1 would indicate that there is unlikely to be a concern following exposure to that VOC in drinking water at the concentrations predicted by the model.

An RCR of greater than 1 indicates that it is not possible to preclude adverse health effects following exposure to a VOC in drinking water at the concentrations predicted in the model. This does not conclusively demonstrate that adverse health effects will occur, due to the inherent limitations of the model and use of large uncertainty factors derivation of many of the TDIs. However, it does highlight a potential for concern that may warrant a more detailed investigation of the occurrence of that VOC in drinking water.

As a worst-case scenario, the estimated drinking water concentration from the Combined Surface Water-Ground Water (Diffuse Source) Model using the poorest drinking water treatment removal technique has been applied to the equation described above to determine the estimated daily dose. This has then been compared to the TDI to develop RCRs for the 14 VOCs. These results are provided in Table 17.6. It should be noted that for the columns provided in this table, the values have been rounded for ease of reading, however, these figures have not been rounded in the model. Therefore, the RCRs presented in Table 17.6 may be slightly different from those that would be derived if using just the numbers presented in this table.

These data indicate that, even using the most conservative assumptions, none of the VOCs have RCRs >1; in fact, all are significantly below 1. Therefore, based on the assumption of a threshold level of toxicity for each of these VOCs, the predicted levels of VOCs in drinking water are not anticipated to be of concern to human health.

However, it should be noted that both 1,3-butadiene and ethylene oxide are considered to be Group 1 carcinogens (i.e. carcinogenic to humans) by the International Agency for Research on Cancer (IARC). Therefore, there is theoretically no ‗safe‘ level for these chemicals, and it would be appropriate to ensure that their concentrations in drinking water are as low as reasonably practicable. However, it would be expected that these chemicals would be manufactured and used in the UK under strictly controlled conditions. This is reflected in the occurrence data that have been located; no information on the occurrence of ethylene oxide in drinking water was located. Although, 1,3-butadiene has been detected in drinking water, its concentration has not been quantified, and has only been detected in surface water at a maximum concentration of 2 µg/l.

Table 17.6 Risk characterisation ratios for the fourteen VOCs using the combined surface water-ground water (diffuse source) model and assuming the poorest removal of the VOC during drinking water treatment

Estimated Tolerable Risk Total Estimated Concentration Daily Intake Characterisation VOC Intake Daily Dose in Drinking (TDI) Ratio (l-eq/day) (µg/kg bw/day)2 Water (µg/l)1 (µg/kg bw/day) (Daily Dose/TDI)

1,2-Dichloropropane 4.47 0.0142 0.0011 14 7.59 x10-5

Dichloromethane 4.39 0.1436 0.0105 6 1.75 x10-3

Aniline 2.37 0.2656 0.0105 7 1.5 x10-3

Benzyl chloride 3.86 0.0025 0.0002 6 2.64 x10-5

1,3-Butadiene 8.72 0.1431 0.0208 5 4.16 x10-3

1,1-Dichloroethane 4.85 0.0332 0.0027 475 5.66 x10-6

Nitrobenzene 2.70 0.549 0.0247 5 4.95 x10-3

Oxirane methyl 2.25 0.2792 0.0105 17 6.16 x10-4

1,2,3- 3.28 0.0043 0.0002 2 1.18 x10-4 Trichloropropane

Urethane 2.00 0.0029 0.0001 0.07 1.38 x10-3

Ethylene oxide 2.54 0.1362 0.0058 15 3.84 x10-4

Formaldehyde 2.78 0.4206 0.0195 86.73 2.25 x10-4 o-Toluidine 2.54 0.0108 0.0005 150 3.04 x10-6

Quinoline 2.63 0.0332 0.0015 25 5.82 x10-5

1. Based on the combined surface water-groundwater model and assuming the poorest performing drinking water treatment process is in operation, see Table 16.14.

3. Based on a conversion of the WHO tolerable concentration of 2.6 mg/l (2600 µg/l) to a bodyweight dose.

18. Conclusions

Table 18.1 Tolerable daily intakes for the fourteen VOCs

VOC Oral TDI Dermal TDI Inhalation TDI

1,2-Dichloropropane 14 µg/kg bw/day 14 µg/kg bw/day 66 µg/m³

Dichloromethane 6 µg/kg bw/day 30 µg/kg bw/day 124 µg/m³

Aniline 7 µg/kg bw/day No TDI derived 6 µg/m³

Benzylchloride 6 µg/kg bw/day No TDI derived 27 µg/m³

1,3-Butadiene 5 µg/kg bw/day No TDI derived 2 µg/m³

1,1-Dichloroethane 475 µg/kg bw/day 475 µg/kg bw/day 3620 µg/m³

Nitrobenzene 5 µg/kg bw/day 50 µg/kg bw/day 0.7 µg/m³

Oxirane methyl 17 µg/kg bw/day No TDI derived 71 µg/m³

1,2,3-Trichloropropane 2 µg/kg bw/day 2 µg/kg bw/day 1.7 µg/m³

Urethane 0.07 µg/kg bw/day 0.07 µg/kg bw/day 250 µg/m³

Ethylene oxide 15 µg/kg bw/day No TDI derived 0.60 µg/m³

Tolerable Concentration: Formaldehyde 1380 µg/kg bw/day 0.06 µg/m³ 2600 µg/l

o-Toluidine 150 µg/kg bw/day 8 µg/kg bw/day 67 µg/m³

Quinoline 25 µg/kg bw/day 25 µg/kg bw/day 19 µg/m³

Due to the gaps in these data, it was considered appropriate to assess the potential health concerns from exposure to VOCs from the consumption of drinking water and bathing and showering on the basis of the relative contribution of each route (oral, dermal and inhalation) compared to consumption on water in terms of litre-equivalents per day. This approach has previously been considered by the World Health Organization (WHO) in the assessment of chloroform, and by Health Canada in consideration of several of their Guidelines for Canadian Drinking Water Quality (GCDWQ).

For the purposes of this assessment, it has been assumed that a 60 kg adult will drink 2 litres of water per day and experience one 30 minute dermal and inhalation exposure (i.e. take either one shower or one bath of 30 minutes duration) each day. This is considered to be a

conservative, but not unrealistic scenario. The total intakes for each of the VOCs (as litre- equivalents per day) are presented in Table 18.2.

Table 18.2 Equivalent total daily intake of the fourteen VOCs following oral consumption of drinking water and one 30-minute bath or shower

Drinking Water Dermal Intake Inhalation Intake (l- Total Intake VOC Intake (l/day) (l-eq/day) eq/day) (l-eq/day)

1,2-Dichloropropane 2.00 1.09 1.38 4.47

Dichloromethane 2.00 0.80 1.59 4.39

Not a significant Aniline 2.00 0.37 2.37 source of exposure

Benzyl chloride 2.00 1.22 0.63 3.86

1,3-Butadiene 2.00 4.47 2.24 8.72

1,1-Dichloroethane 2.00 1.15 1.70 4.85

Not a significant Nitrobenzene 2.00 0.70 2.70 source of exposure

Not a significant Oxirane methyl 2.00 0.25 2.25 source of exposure

1,2,3-Trichloropropane 2.00 0.71 0.56 3.28

Not a significant Not a significant Urethane 2.00 2.00 source of exposure source of exposure

Ethylene oxide 2.00 0.25 0.29 2.54

Not a significant Formaldehyde 2.00 0.78 2.78 source of exposure

Not a significant o-Toluidine 2.00 0.54 2.54 source of exposure

Not a significant Quinoline 2.00 0.63 2.63 source of exposure

Data have also been collated on the physico-chemical properties, use and occurrence of these VOCs in wastewater, environmental water and drinking waters. These data have been used to develop a series of models to estimate the concentrations of these VOCs in drinking water in England and Wales.

Three types of models have been developed to estimate the concentrations in drinking water:

 A ‗surface water model‘ that consists of release of VOCs to wastewater, removal during wastewater treatment, discharge to surface water and abstraction and treatment at a drinking water treatment works;

 A ‗diffuse source‘ groundwater model that assumes release of VOCs to soil over a wide area, migration to groundwater and abstraction and treatment at a drinking water treatment works; and

 A ‗point source‘ groundwater model that assumes release of VOCs in a small area and migration to groundwater and abstraction and treatment at a nearby drinking water treatment works.

The results from the surface water model and the diffuse source groundwater model were combined to provide a worst-case scenario of a drinking water treatment works that is supplied by a river that also receives a significant proportion of flow from a groundwater source.

These models produced estimated concentrations in final drinking water that were applied to the total daily intake (litre-equivalents per day) to calculate an estimated daily dose. These estimated daily doses were then compared to the oral TDIs to determine the Risk Characterisation Ratios (RCRs) for each of the VOCs. The RCRs presented in Table 18.3 are based on poorest performing drinking water treatment process modelled (either air stripping, ozone or GAC) and indicate in none of the VOCs considered are anticipated to be of concern to human health.

However, it should be noted that these TDIs are the assumption of a threshold level for an adverse health effect; 1,3-butadiene and ethylene oxide are considered to be Group 1 carcinogens (i.e. carcinogenic to humans) by the International Agency for Research on Cancer (IARC). Therefore, there is theoretically no ‗safe‘ level for these chemicals, and it would be appropriate to ensure that their concentrations in drinking water are as low as reasonably practicable.

Table 18.3 Risk characterisation ratios for the fourteen VOCs using the combined surface water-ground water (diffuse source) model

1,2-Dichloropropane 4.47 0.0142 0.0011 14 7.59 x10-5

Dichloromethane 4.39 0.1436 0.0105 6 1.75 x10-3

Aniline 2.37 0.2656 0.0105 7 1.5 x10-3

Benzyl chloride 3.86 0.0025 0.0002 6 2.64 x10-5

1,3-Butadiene 8.72 0.1431 0.0208 5 4.16 x10-3

1,1-Dichloroethane 4.85 0.0332 0.0027 475 5.66 x10-6

Nitrobenzene 2.70 0.549 0.0247 5 4.95 x10-3

Oxirane methyl 2.25 0.2792 0.0105 17 6.16 x10-4

1,2,3-Trichloropropane 3.28 0.0043 0.0002 2 1.18 x10-4

Urethane 2.00 0.0029 0.0001 0.07 1.38 x10-3

Ethylene oxide 2.54 0.1362 0.0058 15 3.84 x10-4

Formaldehyde 2.78 0.4206 0.0195 86.73 2.25 x10-4 o-Toluidine 2.54 0.0108 0.0005 150 3.04 x10-6

Quinoline 2.63 0.0332 0.0015 25 5.82 x10-5

1. Based on the combined surface water-groundwater model and assuming the poorest performing drinking water treatment process is in operation, see Table 16.14.

3. Based on a conversion of the WHO tolerable concentration of 2.6 mg/l (2600 µg/l) to a bodyweight dose.

19. Suggestions for Further Research

However, there are several gaps within the data that, if addressed, may aid in the validation of the model and in supporting this conclusion.

Therefore, the following recommendations and considerations are made based on this report:

 The most significant data gap within this report is a lack of information on the current volumes of VOCs manufactured and used within England and Wales and the conditions under which they are manufactured and used. While commercial considerations may mean that this information is unlikely to be forthcoming, such information would significantly aid in the development of the model, and may even allow for the development of models to assess the risks to drinking water supplies in specific regions.

 The available data on the occurrence of these VOCs are also limited. As such, while the results from the model have generally been consistent with reported environmental and drinking water concentrations, it has not been possible to fully validate the results of the model. Further information on the occurrence of these VOCs in the environment, particularly those that are most likely to be of health-concern (such as those considered Group 1 carcinogens by IARC and those with the highest RCRs) would aid in the determination of the reliability of the model predictions.

 It has not been possible to derive tolerable daily intakes via all three of the routes considered within this project (oral, dermal and inhalation) for a number of chemicals due to a lack of relevant toxicological data. To overcome this, route-to-route extrapolations have been applied to derive as many tolerable daily intakes as possible, and a multi-route exposure approach, on the basis of the relative contribution of each route compared to consumption on water in terms of litre-equivalents per day, has been applied. This approach has previously been considered by the World Health Organization (WHO) in the assessment of chloroform, and by Health Canada in consideration of several of their Guidelines for Canadian Drinking Water Quality (GCDWQ). This is considered to be a conservative approach, however, if additional toxicological data were to become available, it would be appropriate to reconsider this approach in light of these additional data.

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Oberst, F.W., Hackley, E.B. and Comstock, C.C. (1956) Chronic toxicity of aniline vapor (5 ppm) by inhalation. Arch. Ind. Health 13, 379-384.

OECD (1979) OECD Expert group degradation/accumulation. Final report Vol.1. Organisation for economic cooperation and development. Berlin: Umweltbundesamt, 141.

OECD (2000) Screening Information Data Set (SIDS) Initial Assessment Report for Aniline. Organisation for Economic Co-operation and Development.

Ott, M.G. and Langner, R.R. (1983) A mortality survey of men engaged in the manufacture of organic dyes. J. Occ. Med. 25: 763-768.

Parodi, S., Sala, M., Russo, P., Zunino, A., Balbi, C., Albini, A., Velerio, F., Cimberle, M.R. and Santi, L. (1982) DNA damage in liver, kidney, bone marrow, and spleen of rats and mice treated with commercial and purified aniline as determined by alkaline elution assay and sister chromatid exchange induction. Cancer Research, 42: 2277-2283.

Pauluhn, J. (2004) Subacute inhalation toxicity of aniline in rats: Analysis of time-dependence and concentration-dependence of hematotoxic and splenic effects. Toxicological Sciences, 81: 198-215.

Pitter, P. (1976) Determination of biological degradability of organic substances. Water Research, 10: 231-5.

Podluzhnyi, P.A. (1979) Gig. i Sanit. No. 1, 44 (1979); cited in Barlow, S. M., F. M. Sullivan (Eds.): Reproductive Hazards of Industrial Chemicals, p 55, Academic Press, London, 1982.

Price, C.J., Tyl, R.W., Marks, T.A., Paschke, L.L., Ledoux, T.A. and Reel, J.R. (1985) Teratologic and postnatal evaluation of aniline hydrochloride in the Fischer 344 rat. Toxicology and Applied Pharmacology, 77: 465-478.

Razo-Fores , E. et al. (1996) Water Science and Technology, 33: 47-57.

Razo-Fores , E. et al. (1997) FEMS Microbiol. Review 20: 525-38.

Richardson, M. (1985) Nitrification inhibition in the treatment of sewage. Royal Society of Chemistry.

Roudabush, R.L., Terhaar, C.J., Fassett, D.W. and Dziuba, S.P. (1965) Comparative acute effects of some chemicals on the skin of rabbits and guinea pigs. Toxicology and Applied Pharmacology, 7: 559-563.

RSC (1991) Chemical Safety Datasheets, Volume 4a: Toxic Chemicals (A-L), RSC. Royal Society of Chemistry.

Ruder, A.M., Ward, E.M., Roberts, D.R., Teass, A.W., Brown, K.K., Fingerhut, M.A. and Stettler, L.E. (1992) Response of the National Institute for Occupational Safety and Health to an occupational health risk from exposure to ortho-toluidine and aniline. Scandinavian Journal of Work, Environment & Health, 18(2): 82-84.

Sarasa J., Roche M.P., Ormad M.P., Gimeno E., Puig A. and Ovelliero J.L. (1998) Treatment of a wastewater resulting from dye manufacturing with ozone and chemical coagulation. Water Research 32(9):2721-2727.

Sauleda R. and Brillas E. (2001) Mineralization of aniline and 4-chlorophenol in acidic solution by ozonation catalysed with Fe2+ and UVA light. Applied Catalysts B: Environmental 29(2):135-145.

Schefer, W. and Waelchli, O. (1980) Testing of the biological degradability of organic chemical wastewater components. Z. Wasser Abwasser Forsch. 13: 205-9.

Short, C.R., King, C., Sistrunk, P.W. and Kerr, K.M. (1983) Subacute toxicity of several ring- substituted dialkylanilines in the rat. Fundamental and Applied Toxicology, 3: 285-292.

Sorahan, T., Hamilton, L. and Jackson, J.R. (2000) A further cohort study of workers employed at a factory manufacturing chemicals for the rubber industry, with special reference to the chemicals 2-mercaptobenzothiazole (mbt), aniline, phenyl-β-naphthylamine and o- toluidine. Occupational and Environmental Medicine, 57(2): 106-115.

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SRC (2013b) Biodeg Chemical Search. Syracuse Research Corporation. Available from http://www.syrres.com/esc/biodeg.htm

Symons, J.M., Mckinney, R.E., Smith, R.M. and Donovan, E.J. (1961) Degradation of nitrogen containing organic compounds by activated sludge. International Journal of Air and Water Pollution, 4, 115.

Sziza, M. and Podhragyai, L. (1957) Toxikologische Untersuchung Einiger in der Ungarischen Industrie zur Anwendung Gelangender Aromatischer Amidoverbindungen. Arch. Gewerbe- Pathol. Gewerbehyg. 15, 447-456.

Tomlinson, T., Boon, A. and Trobman, C. (1966) Inhibition of nitrification in the activated sludge process of sewage disposal. Journal of Applied Bacteriology, 29, 266-291.

Topham, J.C. (1980a) The detection of carcinogen-induced sperm head abnormalities in mice. Mutation Research, 69: 379-387.

Topham, J.C. (1980b) Do induced sperm head abnormalities in mice specifically identify mammalian mutagens rather than carcinogens? Mutation Research, 69: 379-387.

US EPA (1985) Health and Environmental Effects Profile for Aniline. Office of Solid Waste and Emergency Response, Washington, DC, Environmental Criteria and Assessment Office, Cincinnati, OH. United States Environmental Protection Agency. ECAO-CIN-P136.

US EPA (1994) Office of Pollution Prevention and Toxics Chemical Fact Sheets, Aniline Fact Sheet: Support Document. United States Environmental Protection Agency.

US EPA (2012) US EPA Region 5 Superfund. Duell and Gardner landfill: EPA ID# MID980504716. Available from http://www.epa.gov/R5Super/npl/michigan/MID980504716.html

Vasilenko, N.M., Khizhnyakova, L.N., Zvezdai, V.I., Manfanovskii, V.V., Anatovskaya, V.S., Krylova, E.V., Sonkin. I.S., Gnezdilova, A.I., Voskoboinikova, N.A. and Gubina, N.F. (1972a) Clinico-hygienic parallels in the effects of aniline on the body Vrach Delo, 8: 132-134.

Vasilenko, N.M., Volodchenko, V.A., Khizhnyakova, L.N., Zvezdai, V.I., Manfanovskii, V.V., Anatovskaya, V.S., Krylova, E.V., Voskoboinikova, N.A., Gnezdilova, A.I. and Sonkin. I.S. (1972b) Information to substantiate a decrease of the maximum permissible concentration of aniline in the air of working zones Gigiena i Sanitarija, 37: 31-35.

Vlachos, D.A. (1989) Mouse Bone Marrow Micronucleus Assay of Aniline. Du Pont HLR 263- 89, Unpublished.

Wangenheim, J. and Bolcsfoldi, G. (1988) Mouse lymphoma L5178Y thymidine kinase locus assay of 50 compounds. Mutagenesis, 3: 193-205.

Ward, E., Carpenter, A., Markowith, S., Roberts, D. and Halperin, W. (1991) Excess number of bladder cancers in workers exposed to ortho-toluidine and aniline. Journal of the National Cancer Institute, 83: 501-506.

Wegman, R.C.C. and DeKorte, G.A.L. (1981) Water Research 15: 391-4.

Westmoreland, C. and Gatehouse, D.G. (1991) Effects of aniline hydrochloride in the mouse bone marrow micronucleus test after oral administration. Carcinogenesis, 12: 1057-1059.

Wilmer, J.L., Kligerman, A.D. and Erexson, G.L. (1981) Sister chromatid exchange induction and cell cycle inhibition by aniline and its metabolites in human fibroblasts. Environm. Mutag. 3, 627-638.

Wilmer, J.L., Erexson, G.L. and Kligerman, A.D. (1984) The effect of erythrocytes and haemoglobin on sister chromatid exchange induction in cultured human lymphocytes exposed to aniline Hcl. Basis Life Science 293, 561-567.

Wu, Z., Liu, Y., Liu, H., Xia, Y., Shen, W., Hong, Q., Li, S. and Yao, H. (2012) Characterization of the nitrobenzene-degrading strain Pseudomonas sp. a3 and use of its immobilized cells in the treatment of mixed aromatics wastewater. World Journal of Microbiology and Biotechnology 28 (8): 2679-2687

Xu, Z. et al. (1990) Huanjing Kexue 11: 29-31.

Yoshimi, N., Sugie, S., Iwata, H., Niwa, K., Mori, H., Hashida, C. and Shimizu, H. (1988) The Genotoxicity of a variety of aniline derivates in a DNA repair test with primary cultured rat hepatocytes. Mutation research, 206: 183-191.

Yoshioka, Y., Nagase, H., Ose, Y. and Sato, T. (1986) Evaluation of the test method activated sludge, respiration inhibition test. proposed by the OECD. Ecotoxicology and Environmental Safety. 12: 206-212.

Zahn, R. and Wellens, H. (1980) Examination of biological degradability in static tests – Further experience and new application possibilities (German) Z. Wasser Abwasser Forsch. 13: 1-7.

20.4 Benzylchloride

Aranyi, C., O'Shea, W., Graham, J.A. and Miller, F. (1986) The Effects Of Inhalation Of Organic Chemical Air Contaminants On Murine Lung Host Defenses. Fundam. Appl. Toxicol., 6, 713-720.

Ashby, J., Gaunt, C. and Robinson. (1982) Carcinogenicity bioassay of 4-chloromethyl biphenyl (4CMB), 4-hydroxymethyl (4HMB) and benzyl chloride (BC) on mouse skin. Interim 7-month report. Mutat. Res. 100 (1-4):399-401.

Booth, S.C., Mould, A.J., Shaw, A. and Garner, R.C. (1983) The biological activity of 4- chloromethylbiphenyl, benzyl chloride and 4-hydroxymethylbiphenyl in four short-term tests

for carcinogenicity. A report of an individual study in the U.K.E.M.S. genotoxicity trial 1981. Mutat Res 119: 121–133.

ChemID (2013) ChemID Plus Advanced. US National Library of Medicine. Available from http://chem.sis.nlm.nih.gov/chemidplus/

Coombs, M.M. (1982a) Attempts to initiate skin tumours in mice in the 2-stage system using 4-chloromethylbiphenyl (4CMB), 4-hydroxymethylbiphenyl (4HMB), and benzyl chloride (BC) Report of the experiment at 10 months. Mutat. Res. 100:403-405.

Coombs, M.M. (1982b) The UKEMS Genotoxicity Trial: A Summary of the assays for skin tumor induction in mice, the subcutaneous implant test and the sebaceous gland suppression test. Mutat. Res. 100:407-409.

De Ceaurriz, J., Desiles, J.P., Bonnet, P., Marignac, B., Muller and, Guenier, J.P. (1983) Concentration-dependent behavioral changes in mice following short-term inhalation exposure to various industrial solvents. Toxicol Appl Pharmacol. 67(3):383-9.

ECHA (2013) REACH registration dossier for α-chlorotoluene.

EFSA (2011) EFSA Panel on Food Additives and Nutrient Sources added to Food (ANS); Scientific Opinion on the reevaluation of butylated hydroxyanisole–BHA (E 320) as a food additive. EFSA Journal 2011; 9(10):2392.

Environment Canada and Health Canada (2009) Screening Assessment for the Challenge Benzene, (chloromethyl)-(Benzyl chloride) Chemical Abstracts Service Registry Number 100- 44-7.

Fahmy, M.J. and Fahmy, O.G. (1982) Genetic activities of 4-chloromethylbiphenyl, the 4- hydroxy derivative and benzyl chloride in the soma and germ line of Drosophila melanogaster. Mutat Res 100: 339–344.

Fukuda, K., Matsushita, H., Sakabe, H. and Takemoto, K. (1981) Carcinogenicity of benzyl chloride, benzal chloride, benzotrichloride and benzoyl chloride in mice by skin application.

Holmberg, B. and Malmfors, T. (1974) The effect of industrial solvents on adrenergic transmitter mechanisms. Wenner-Gren Cent. Ist. Symp. Ser., 22, 191-200.

HSDB (2013) Hazardous Substances Databank. US National Library of Medicine. Available from http://toxnet.nlm.nih.gov/cgi-bin/sis/htmlgen?HSDB

IARC (1999) Summaries & Evaluations. Volume 71: α-Chlorinated Toluenes and Benzoyl Chloride. International Agency for Research on Cancer.

IUCLID (2000) alpha-Chlorotoluene. International Uniform Chemical Information Dataset. European Chemicals Bureau. European Commission.

Jacobson, S.N. and Alexander, M. (1981) Appl. Environ. Microbiol. 42: 1062-6.

James, R.H. et al. (1984) Journal of Air Pollution Control Association, 35: 959-61.

Kitano, M. (1978) Biodegradation and Bioaccumulation Test on Chemical Substances. OECD Tokyo Meeting Reference Book Tsu-No. 3.

Lijinsky, W. (1986) Chronic Bioassay of Benzyl Chloride in F344 Rats and (C57BL/6JXBALB/C)F1 Mice. J. Natl. Cancer Inst. 76(6): 1231-1236

OECD (1998) Benzyl chloride. SIDS Initial Assessment Report for 8th SIAM, France, October 28-30, 1998. Organization for Economic Cooperation and Development.

McGregor, D.B., Brown, A., Cattanach. P., Edwards, I., McBride, D. and Caspary, W.J. (1988) Responses of the L5178Y tk+/tk- mouse lymphoma cell forward mutation assay to coded chemicals. II. 18 coded chemicals. Environ. Molec. Mutagen. 11:91-118.

Mirzayans, F., Davis, P.J., Parry, J.M. 1982. The cytotoxic and mutagenic effects of 4CMB, BC and 4HMB in V79 Chinese hamster cells. Mutat Res 100: 239–244.

Rittfeldt, L., Ahlberg, M.S., Zingmark, P.A. and Santesson, J. (1983) Occupational exposure to benzyl chloride and benzal chloride due to contaminated butyl benzyl phthalate. Scand J Work Environ Health 9:367-368.

Rudnev, M.I. et al. (1979) Cited in Benzychlorid, Bericht NR. 48, Berufsgenossenschaft der chemischen Industrie (1990)

Sakabe, H., Matsushita, H. and Koshi, S. (1976) Cancer among benzoyl chloride manufacturing workers. Ann. NY. Acad. Sci. 271:67-70.

Sakabe, H. and Fukuda, K. (1977) An updating report on cancer among benzoyl chloride manufacturing workers. Ind. Health. 15(3-4):173-174.

Sasaki, S. (1978) Aquatic Pollutants Transformations and Biological Effects. NY, NY: Pergamon Press, 283-98.

Schmuck, G., Lieb, G., Wild, D., Schiffmann, D. and Henschler, D. (1988) Characterization of an in vitro micronucleus assay with Syrian hamster embryo fibroblasts. Mutat Res 203: 397– 404.

Scott, K. and Topham, J.C. (1982) Sperm head abnormality test. Mutat Res. 100(1-4):345-50.

Sheldon, L.S. and Hites, R.A. (1978) Environ Sci Technol 12: 188-94.

Skowronski, G. and Abdel-Rahman, S. (1986) Teratogenicity of benzyl chloride in the rat. J. Toxicol. Environ. Health, 17, 51-56.

Solveig Walles, S.A. (1981) Reaction of benzyl chloride with hemoglobin and DNA in various organs of mice. Toxicol Lett 9: 379–387

Sorahan, T., Waterhouse, J.A., Cooke, M.A., Smith, E.M., Jackson, J.R. and Temkin, L. (1983) A mortality study of workers in a factory manufacturing chlorinated toluenes. Ann Occup Hyg. 27(2):173-82.

Sorahan, T. and Cathcart, M. (1989) Lung cancer mortality among workers in a factory manufacturing chlorinated toluenes: 1961-84. British Journal of Industrial Medicine 46:425- 427.

SRC (2013a) Physprop Database. Syracuse Research Corporation. Available from http://www.syrres.com/what-we-do/databaseforms.aspx?id=386

US EPA (1994) Integrated Risk Information System (IRIS) online. Benzyl chloride (CASRN 100-44-7). National Center for Environmental Assessment, Washington, D.C.

US EPA (2013) Chemical Contaminant List 3 (CCL 3) United States Environmental Protection Agency. Available from http://www.epa.gov/ogwdw/ccl/ccl3.html

Vernot, E.H., MacEwen, J.D., Haun, C.C. and Kinkead, E.R. (1977) Acute toxicity and skin corrosion data for some organic and inorganic compounds and aqueous solutions. Toxicol. Appl. Pharmacol. 42:417-423

Wong, O. (1988) A cohort mortality study of employees exposed to chlorinated chemicals. Am. J. Ind. Med., 14, 417-431.

Zeiger, E., Anderson, B., Haworth, S., Lawlor, T., Mortelmans, K. and Speck, W. (1987) Salmonella mutagenicity tests. III. Results from the testing of 225 chemicals. Environ. Mutagen. Vol 9 (Suppl 9):1-109

Zissu, D. (1995) Histopathological changes in the respiratory tract of mice exposed to ten families of airborne chemicals. Journal of Applied Toxicology, 15(3): 207-213

20.5 1,3-Butadiene

Abdel-Rahman, S.Z., Ammenheuser, M.M. and Ward, J.B. Jr. (2001) Human sensitivity to 1,3- butadiene: role of microsomal epoxide hydrolase polymorphisms. Carcinogenesis, 22(3): 415–423

Abdel-Rahman, S.Z., El-Zein, R.A., Ammenheuser, M.M., Yang, Z., Stock, T.H., Morandi, M. and Ward, J.B. Jr. (2003) Variability in human sensitivity to 1,3-butadiene: Influence of the allelic variants of the microsomal epoxide hydrolase gene. Environ Mol Mutagen, 41(2): 140– 146

Adler, I-D., Cao, J., Filser, J.G., Gassner, P., Kessler, W., Kliesch, U., Neuhäuser-Klaus, A. and Nüsse, M. (1994) Mutagenicity of 1,3-butadiene inhalation in somatic and germinal cells of mice. Mutat Res. 309: 307-314.

Ahlberg, R.W., Sorsa, M., Pfäffli, P., Mäki-Paakkanen, J. and Viinanen, R. (1992) Genotoxicity and exposure assessment in the manufacture of 1,3-butadiene. In: Occupational Health in the Chemical Industry. WHO: Copenhagen, Denmark. pp 199-204.

Altshuller, A.P., Klosterman, D.L., Leach, P.W., Hindawi, I.J. and Sigsby, J.E. Jr. (1966) Products and biological effects from irradiation of nitrogen oxides with hydrocarbons or aldehydes under dynamic conditions. Air and Wat Pollut Int J. 10; 81-98.

Anderson, D., Edwards, A.J. and Brinkworth, M.H. (1993) Male-mediated F1 effects in mice exposed to 1,3-butadiene. In: Sorsa et al. (eds) Butadiene and styrene: Assessment of health hazards. IARC Scientific Publications. IARC: Lyon, France. pp 171-181.

Anderson, D., Hughes, J.A., Brinkworth, M.H., Peltonen, K. and Sorsa, M. (1996) Levels of ras oncoproteins in human plasma from 1,3-butadiene-exposed workers and controls. Mutat Res. 349: 115-120.

Arce, G.T., Vincent, D.R., Cunningham, M.J., Choy, W.N. and Sarrif, A.M. (1990) In vitro and in vivo genotoxicity of 1,3-butadiene and metabolites. Environ Health Persp. 86; 75-78.

ATSDR (2012) Toxicological profile for 1,3-butadiene. Agency for Toxic Substances and Disease Registry. U.S. Department of health and human services, Public Health Service.

Au, W.W., Bechtold, W.E., Whorton, E.B. Jr. and Legator, M.S. (1995) Chromosome aberrations and response to γ-ray challenge in lymphocytes of workers exposed to 1,3- butadiene. Mutat Res. 334: 125-130.

Autio, K., Renzi, L., Catalan, J., Albrecht, O.E. and Sorsa, M. (1994) Induction of micronuclei in peripheral blood and bone marrow erythrocytes of rats and mice exposed to 1,3-butadiene by inhalation. Mutat Res. 309: 315-320.

Carpenter, C.P., Shaffer, C.B., Weir, C.S. and Smyth, H.F. (1944) Studies on the inhalation of 1,3-butadiene; with a comparison of its narcotic effect with benzol, toluol and styrene, and a note on the elimination of styrene by the human. J Ind Hyg Toxicol 26: 69-78.

Carpenter, C.P. and Smyth, H.R. Jr. (1946) Chemical burns of the rabbit cornea. Amer. J. Ophthalmol. 29:1363-1372.

ChemID (2013) ChemID Plus Advanced. US National Library of Medicine. Available from http://chem.sis.nlm.nih.gov/chemidplus/

Cheng, H., Sathiakumar, N., Graff, J., Matthews, R. and Delzell, E. (2007) 1,3-Butadiene and leukemia among synthetic rubber industry workers: Exposure–response relationships. Chem- Biol Interact., 166, 15–24.

Chou, M.S., Huang, B.J., Chang, H.Y. (2005) Decomposition of gas phase 1,3-butadiene by ultraviolet/ozone processes. Journal of the Air and Waste Management Association, 55(7):919-929.

Cowles, S.R., Tsai, S.P., Snyder, P.J. and Ross, C.E. (1994) Mortality, morbidity, and haematologic results from a cohort of long-term workers involved in 1,3-butadiene monomer production. Occup Environ Med. 51; 323-329.

Cunningham, M.J., Choy, W.N., Arce, G.T., Rickard, L.B., Vlachos, D.A., Kinney, L.A. and Sarrif, A.M. (1986) In vivo sister chromatid exchange and micronucleus induction studies with 1,3-butadiene in B6C3F1 mice and Sprague-Dawley rats. Mutagen. 1: 449-452.

de Meester, C., Poncelet, F., Roberfroid, M. and Mercier, M. (1980) The mutagenicity of butadiene towards Salmonella typhimurium. Toxicol Lett 6(3):125-130.

Delzell, E., Sathiakumar, N., Graff, J., Macaluso, M., Maldonado, G. and Matthew, R. (2006) An Updated Study of Mortality among North American Synthetic Rubber Industry Workers, Boston, MA, Health Effects Institute.

Divine, B.J. and Hartman, C.M. (2001) A cohort mortality study among workers at a 1,3- butadiene facility. Chem.-biol. Interact., 135–136, 535–553

Environment Canada (2000) 1,3-Butadiene. Priority Substances List Assessment Report. Canadian Environmental Protection Act, 1999. Health Canada.

EU (2002) 1,3-Butadiene. EU Risk Assessment Report. Final Report. Institute for Health and Consumer Protection. European Chemicals Bureau.

Ewing, B.B. et al (1977) Monitoring to Detect Previously Unrecognized Pollut. in Surface Waters. Washington, DC: USEPA-560/6-77-015.

Gostinsky, V.D. (1965) The toxicity of isoprene and the maximum permissible concentration of its vapour in the atmosphere of industrial premises. Gig Tr Prof Zabol. 9: 36-42

Hackett, P.L., Mast, T.J., Brown, M.G., Clark, M.L., Evanoff, J.J., Rowe, S.E., McClanahan, B.J., Buschbom, R.L., Decker, J.R., Rommereim, R.L. and Westerberg, R.B. (1988) Dominant lethal study in CD-1 mice following inhalation exposure to 1,3-butadiene. Report No. NIH-Y01- ES-70153. Pacific Northwest Laboratory: Washington, USA.

Hayes, R.B., Xi, L., Bechtold, W.E., Rothman, N., Yao, M., Henderson, R., Zhang, L., Smith, M.T., Zhang, D., Wiemels, J., Dosemeci, M., Yin, S. and O‘Neill, J.P. (1996) hprt Mutation frequency among workers exposed to 1,3-butadiene in China. Toxicology 113: 100-105.

Hayes, R.B., Zhang, L., Yin, S., Swenberg, J.A., Xi, L., Wiencke, J., Bechtold, W.E., Yao, M., Rothman, N., Haas, R., O‘Neill, J.P., Zhang, D., Wiemels, J., Dosemeci, M., Li, G. and Smith, M.T, (2000) Genotoxic markers among butadiene polymer workers in China. Carcinogenesis 21: 55-62.

Hazleton Laboratories, Europe, Ltd. (1981) The Toxicity and Carcinogenicity of Butadiene Gas Administered to Rats by Inhalation for Approximately 24 Months, Report No. 2653-522/2. International Institute of Synthetic Rubber Producers.

Hazleton Laboratories, Europe, Ltd. (1982) 1,3-Butadiene: Inhalation Teratogenicity Study in the Rat, Report No. 2653-522/2; HSE-81-0040/3, November 1; HSE-81-0040, Addendum 1, March 1982. International Institute of Synthetic Rubber Producers.

IARC (2008) World Health Organization International Agency for Research on Cancer. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans. Volume 97: 1,3-Butadiene, Ethylene Oxide and Vinyl Halides (Vinyl Fluoride, Vinyl Chloride and Vinyl Bromide).

IARC (2012) World Health Organization International Agency for Research on Cancer. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans. Volume 97, 100F 1,3- Butadiene.

Irons, R.D., Smith, C.N., Stillman, W.S., Shah, R.S., Steinhagen, W.H. and Leiderman, L.J. (1986a) Macrocytic-megaloblastic anemia in male B6C3F1 mice following chronic exposure to 1,3-butadiene. Toxicol Appl Pharmacol. 83: 95-100.

Irons, R.D., Smith, C.N., Stillman, W.S., Shah, R.S., Steinhagen, W.H. and Leiderman, L.J. (1986b) Macrocytic-megaloblastic anemia in male NIH Swiss mice following repeated exposure to 1,3-butadiene. Toxicol Appl Pharmacol. 85: 450-455.

Irons, R.D., Oshimura, M. and Barrett, J.C. (1987) Chromosome aberrations in mouse bone marrow cells following in vivo exposure to 1,3-butadiene. Carcinogenesis. 8: 1711-1714.

IUCLID (2000) 1,3-Butadiene. International Uniform Chemical Information Dataset. European Chemicals Bureau. European Commission.

Jelitto, B., Vangala, R.R. and Laib, R.J. (1989) Species differences in DNA damage by butadiene: Role of diepoxybutane. Arch Toxicol Suppl 13:246-249.

Kelsey, K.T., Wiencke, J.K., Ward, J., Bechtold, W. and Fajen, J. (1995) Sister-chromatid exchanges, glutathione S-transferase θ deletion and cytogenetic sensitivity to diepoxybutane in lymphocytes from butadiene monomer production workers. Mutat Res. 335: 267-273.

Larionov, L.F. et al. (1934) The physiological action of butadiene, butene-2 and isoprene. Kazanskii Meditsinkii Zhurnal. 30; 440-445 (HSE translation no. 10855).

Legator, M.S., Au, W.W., Ammenheuser, M. and Ward, J.B. Jr. (1993) Elevated somatic cell mutant frequencies and altered DNA repair responses in nonsmoking workers exposed to 1,3- butadiene. IARC Sci Publ. 1993;(127):253-63.

Mäki-Paakkanen, J., Nylund, L., Hayashi, M., Norppa, H., Sorsa, M. (1993) Cytogenetic studies with 1,3-butadiene, vinyl acetate, and styrene in rats and mice. (No further details available).

NTP (1984) NTP Technical Report on the Toxicology and Carcinogenesis of 1,3-butadiene

(CAS No. 106-99-0) in BCF3F1 Mice (Inhalation Studies). National Toxicology Program, Technical Report Series No. 288, NIH Publication No. 84-2544.

NTP (1987a) Inhalation Developmental Toxicology Studies: Teratology Study of 1,3-butadiene (CAS No. 106-99-0) in Mice. National Toxicology Program study TER85144.

NTP (1987b) Inhalation Developmental Toxicology Studies: Teratology Study of 1,3-butadiene (CAS No. 106-99-0) in the rat. National Toxicology Program study TER89041.

NTP (1993) NTP Technical Report on the Toxicology and Carcinogenesis Studies of 1,3- Butadiene (CAS No. 106-99-0) in B6C3F1 Mice (Inhalation Studies). National Toxicology Program, Technical Report Series No. 434, NIH Publication No. 93-3165.

OECD (1996) Screening Information Data Set (SIDS). Initial Assessment Report for 1,3- butadiene. Organisation for Economic Co-operation and Development.

Przygoda, R.T., Bird, M.G., Whitman, F.T., Wojcik, N.C. and McKee, R.H. (1993) Induction of micronuclei in mice and hamsters by 1,3-butadiene. Environ Mut. 56.

Recio, L., Osterman-Golkar, S., Csanady, G.A., Turner, M.J., Myhr, B., Moss, O. and Bond, J.A. (1992) Determination of mutagenicity in tissues of transgenic mice following exposure to 1,3-butadiene and N-ethyl-N-nitrosourea. Toxicol Appl Pharmacol. 117; 58-64.

Ripp, G.K. (1969) Toxicohygienic characteristics of 1,3-butadiene. Tr Omsk Med Inst. 88; 10- 18 (NTC translation 75 –13092-06T).

Sasiadek, M., Järventaus, H. and Sorsa, M. (1991a) Sister-chromatid exchanges induced by 1,3-butadiene and its epoxides in CHO cells. Mutat Res. 263: 47-50.

Sasiadek, M., Norppa, H. and Sorsa, M. (1991b) 1,3-butadiene and its epoxides induce sister- chromatid exchanges in human lymphocytes in vitro. Mutat Res. 261: 117-121.

Sasiadek, M. (1992) Cytogenetic studies of workers from the rubber industry. Mutat Res. 279: 195-198.

Sernau, R., Cavagnaro, J. and Kehn, P. (1986) 1,3-butadiene as an S9 activation-dependent gaseous positive control substance in L5178Y cell mutation assays. Environ Mut. 8 (suppl 6): 75-76.

Shugaev, B. (1969) Concentrations of hydrocarbons in tissues as a measure of toxicity. Arch. Environ. Health, 18, 878-882.

Sills, R.C., Hong, H.L., Boorman, G.A., Devereux, T.R. and Melnick, R.L. (2001) Point mutations of K-ras and H-ras genes in forestomach neoplasms from control B6C3F1 mice and following exposure to 1,3-butadiene, isoprene or chloroprene for up to 2-years. Chem Biol Interact 135-136:373-386.

Sorsa, M., Autio, K., Demopoulos, N.A., Järventaus, H., Rössner, P., Srám, R.J., Stephanou, G. and Vlachodimitropoulòs, D. (1994) Human cytogenetic biomonitoring of occupational exposure to 1,3-butadiene. Mutat Res. 309: 321-326.

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Tates, A.D., van Dam, F.J., de Zwart, F.A., van Teylingen, C.M.M. and Natarajan, A.T. (1994) Development of a cloning assay with high cloning efficiency to detect induction of 6- thioguanine-resistant lymphocytes in spleen of adult mice following in vivo inhalation exposure to 1,3-butadiene. Mutat Res. 309: 299-306.

Tates, A.D., van Dam, F.J., de Zwart, F.A., Darroudi, F., Natarajan, A.T., Rössner, P., Peterková, K., Peltonen, K., Demopoulos, N.A., Stephanou, G., Vlachondimitropoulos, D. and Srám, R.J. (1996) Biological effect monitoring in industrial workers from the Czech Republic exposed to low levels of butadiene. Toxicol. 113: 91-Toronto (2013).

Tsai, S.P., Wendt, J.K. and Ransdell, J.D. (2001) A mortality, morbidity, and hematology study of petrochemical employees potentially exposed to 1,3-butadiene monomer. Chem-Biol Interact, 135–136, 555–567

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Victorin, K., Busk, L., Cederberg, H. and Magnusson, J. (1990) Genotoxic activity of 1,3- butadiene and nitrogen dioxide and their photochemical reaction products in Drosophila and in the mouse bone marrow micronucleus assay. Mutat Res. 228: 203-209.

Vincent, D.R., Arce, G.T. and Sarrif, A.M. (1986) Genotoxicity of 1,3-butadiene. Assessment by the unscheduled DNA synthesis assay in B6C3F1 mice and Sprague-Dawley rats in vivo and in vitro. Environ Mut. 8 (suppl 6): 88.

Ward, E.M., Fajen, J.M., Ruder, A.M., Rinsky, R.A., Halperin, W.E. and Fesler-Flesch, C.A. (1995) Mortality study of workers in 1,3-butadiene production units identified from a chemical workers cohort. Environ Health Perspect. 103; 598-603.

Ward, J.B. Jr., Ammenheuser, M.M., Whorton, E.B. Jr., Bechtold, W.E., Kelsey, K.T. and Legator, M.S. (1996a) Biological monitoring for mutagenic effects of occupational exposure to butadiene. Toxicol. 113; 84-90.

Ward, E.M., Fajen, J.M., Ruder, A.M., Rinsky, R.A., Halperin, W.E. and Fesler-Flesch, C.A. (1996b) Mortality study of workers employed in 1,3-butadiene production units identified from a large chemical workers cohort. Toxicol. 113; 157-168.

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Xiao, Y. and Tates, A.D. (1995) Clastogenic effects of 1,3-butadiene and its metabolites 1,2- epoxybutene and 1,2,3,4-diepoxybutane in splenocytes and germ cells of rats and mice in vivo. Environ Mol Mutagen. 26: 97-108.

Zlobina, M. and Dueva, I. (1974) Allergenic action of polybutadiene Latex SKOP and its volatile products. Chem. Abstr. 1978481 :34129m.

20.6 1,1-Dichloroethane

American Conference of Governmental Industrial Hygienists, (1991) Inc. Documentation of the Threshold Limit Values and Biological Exposure Indices. 6th ed. Volumes I, II, III. Cincinnati, OH: ACGIH

ATSDR (1990) Toxicological profile for 1,1-dichloroethane. Agency for Toxic Substances and Disease Registry.

ATSDR (1994) Toxicological profile for 1,1-Dichloroethane. Agency for Toxic Substances & Disease Registry. U.S. Department of Health and Human Services.

Baehr, A.L. et al. (1999) Water Res 35: 127-36.

Bronzetti, G., Galli, A., Vellosi, R. et al. (1987) Genetic activity of chlorinated ethanes. Eur J Cancer Clin Oncol 23:1737-1738.

Browning, E. (1965) Toxicity and metabolism of industrial solvents. New York, NY: Elsevier Science Publishing Co., Inc., 247-252.

Burmaster, D.E. (1982) Environment 24: 6-13, 33-6.

Callander, R. and Sheard, C. (1992) 1,1-dichloroethane: Genetic Toxicology Screening. Ames test. CTL/L/4471.

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Chen, W.R., Sharpless, C.M., Linden, K.G. and Suffet, I.H. (2006) Treatment of volatile organic chemicals on the EPA contaminant candidate list using ozonation and the O3/H2O2 advanced oxidation process. Environmental Science and Technology 40(8):2734-2739.

Colacci, A., Arfellini, G., Mazzullo, M. et al. (1985) Genotoxicity of 1,1-dichloroethane. Research Communications in Chemical Pathology and Pharmacology, 49:243-254.

Cole, R.H. et al. (1984) J. Water Pollution Control Fed. 56: 898-908.

Dobbs R. and Cohen J. (1980) Carbon adsorption isotherms for toxic organics. Municipal Environmental Research Laboratory, Office of Research and Development, US Environmental Protection Agency. EPA-600/8-80-023.

Dow Chemical Company. Unpublished data. Midland, Michigan, 48640.

Dow Chemical Company. (1971) Unpublished data. American Industrial Hygiene Association: Hygienic Guide- 1,1-dichloroethane.

Dyksen, J.E. and Hess, A.F. (1982) Journal of American Water Works Association 74: 394- 403.

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ECHA (2013) REACH registration dossier for nitrobenzene.

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Fronk, C.A. (1987) Destruction of volatile organic contaminants in drinking water by ozone treatment. Ozone Science & Engineering 9(3):265-288.

Fusillo, T.V. et al. (1985) Ground Water 23: 354-60

GLWQB (1983) Great Lakes Water Quality Board; An Inventory Of Chemical Substances . The Great Lakes Ecosystem Vol 1. Windsor Ontario, Canada.

Grayson, M. (1978) Kirk-Othmer encyclopedia of chemistry and technology. Vol. 5, 3rd ed. New York, NY: John Wiley and Sons, Inc.

Hatch, G.G., Mamay, P.D., Ayer, M.L., Casto, B.C. and Nesnow, S. Chemical enhancement of viral transformation in Syrian hamster embryo cells. Cancer Research, 43: 1945-1950.

Hofmann, H.T., Birnstiel, H. and Jobst, P. (1971) Inhalation toxicity of 1,1 and 1,2- dichloroethane. Archiv für Toxikologie, 27(3): 248-265.

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Kelley, R.D. (1985) Synthetic Organic Compounds Sampling Survey of Public Water Supplies p. 38 NTIS PB85-214427.

Klaunig, J.E., Ruth, R.J. and Pereira, M.A. (1986) Carcinogenicity of chlorinated methane and ethane compounds administered in drinking water to mice. Environmental Health Perspectives, 69: 89-95.

Klinkead, E.R. and Leaky, H.F. (1987) Evaluation of the Acute Toxicity of Selected Groundwater Contaminants. Report No. AD-A180-198.

Kolpin, D.W. et al. (1997) Groundwater in the Urban Environment. Chilton et al, eds. Rotterdam, The Netherlands: Blakema, 469-74.

Krill, R.M. and Sonzogni, W.C. (1986) J. Amer. Water Works Assoc. 78: 70-5.

Lesage, S. (1991) Characterisation of groundwater contaminants using dynamic thermal stripping and absorption/thermal desorption-GC-MS. Fresenius Journal of Analytical Chemistry 339: 516-527.

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Milman, H.A., Story, D.L., Riccio, E.S., et al. (1988) Rat liver foci and in vitro assays to detect initiating and promotory effects of chlorinated ethanes and ethylenes. Ann. NY Acad. Sci., 534: 521-530.

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NCI. (1977) Bioassay of 1,1-dichloroethane for possible carcinogenicity. Technical Report Series No. 66, Department of Health and Human Services, Public Health Service, National Cancer Institute. National Institutes of Health.

Nohmi, T., Miyata, R., Yoshikawa, K. and Ishidate, M. (1985) Mutagenicity tests on organic chemical contaminants in city water and related compounds. I. Bacterial Mutagenicity Tests. Bull. Natl. Inst. Hyg. Sci., 103: 60-64.

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ORVWSC (1982) Ohio River Valley Sanit. Comm. (1982) Assessment of Water Quality Conditions Ohio River Mainstream 1980-81.

Otson, R. et al. (1982) J. Assoc. Offic. Analyt. Chem. 65: 1370-4.

Parr-Dobrzanski, R.J. (1992) 1,1-Dichloroethane: 4-day preliminary inhalation toxicity study in the rat. CTL/L/4931.

Pattys Industrial Hygiene and Toxicology 3rd edition (1981) Wiley and Sons, New York.

Piet, G.J. and Morra, C.F. (1983) Artificial Groundwater Recharge. Huisman, L. and Olsthorn, T.N. (editors); Pitman Pub.

Qi, S., Snoeyink, V.L., Beck, E.A., Koffskey, W.E. and Lykins, B.W. (1992) Using Isotherms To Predict GACs Capacity for Synthetic Organics. Journal of American Water Works Association, 84(9):113-120.

Riccio, E., Griffin, A., Mortelmans, K., et al. (1983) A comparative mutagenicity study of volatile halogenated hydrocarbons using different metabolic activation systems. Environ Mut 5:472.

RTECS. (1988) Registry of Toxic Effects of Chemical Substances. National Library of Medicine, National Toxicology Information Program, Bethesda, MD. December 9, 1988.

Sabel, G.V. and Clark, T.P. (1984) Waste Management Research 2: 119-30.

Schwetz, B.A., Leong, B.K.J. and Gehring, P.J. (1974) Embryo- and fetotoxicity of inhaled carbon tetrachloride, 1,1-dichloroethane and methyl ethyl ketone in rats. Toxicology and Applied Pharmacology, 28: 452-464.

Simmon, V.F., Kauhanen, K., Mortelmans, K., Tardiff, R.G. (1977) Mutagenic activity of chemicals identified in drinking water. Mutation Research, 53(2): 261-262.

Smyth, H.F. Jr. (1956) Improved communication: Hygienic standards for daily inhalation. American Industrial Hygiene Association Quarterly., 129-195.

Speth T.F. and Miltner R.J. (1990) Technical Note: Adsorption capacity of GAC for synthetic organics. Journal of American Water Works Association 82(2):72-75.

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WHO (2003) WHO 1,1-Dichloroethane in Drinking-water. Background document for development of WHO Guidelines for Drinking-water Quality. World Health Organization. WHO/SDE/WSH/03.04/19

Young, D.R. et al. (1983) pp. 871-84 in Water Chlorination Vol 4 Book 2.

20.7 Nitrobenzene

Abrams, E.F. et al. (1975) Springfield, VA: Versar Inc.

Agency for Toxic Substances and Disease Registry (ATSDR) (1990) Toxicological profile for 1,1-dichloroethane. Agency for Toxic Substances and Disease Registry (ATSDR). U.S. Public Health Service.

Akiyama, T. et al. (1980) J UOEH 2: 285-300.

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Biodynamics Inc (1984) An Inhalation Teratology Study in Rabbits with Nitrobenzene. EPA Document No. 40-8424492, Fiche No. OTS0510651; submitted to Nitrobenzene Assoc., Wilmington, Delaware.

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Bringmann, G. and Kühn, R. (1980) Comparison of the toxicity thresholds of water pollutants to bacteria, algae, and protozoa in the cell multiplication inhibition test. Water Res, 14: 231- 241.

Burns, L.A., Bradley, S.G., White, K.L. Jr., McCay, J.A., Fuchs, B.A., Stern, M., Brown, R.D., Musgrove, D.L., Holsapple, M.P., Luster, M.I. and Munson, A.E. (1994) Immunotoxicity of Nitrobenzene in Female B6C3F1 Mice. Drug and Chemical Toxicology, 17(3):271-315

Canton, J.H., Slooff, W., Kool, H.J. et al. (1985) Toxicity, biodegradability, and accumulation of a number of Cl/N-containing compounds for classification and establishing water quality criteria. Regul Toxico1 Pharmacol 5:123-131.

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Chou, W.L., Speece, R.E. and Siddiqi, R.H. (1978) Acclimation and degradation of petrochemical wastewater components by methane fermentation. Biotechnol Bioeng Symp 8:391-414.

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Dodd, D.E., Fowler, E.H., Snellings, W.M., Pritts, I.M., Tyl, R.W., Lyon, J.P., O‘Neal, F.O. and Kimmerle, G. (1987) Reproduction and Fertility Evaluations in CD Rats following Nitrobenzene Inhalation. Fundamental and Applied Toxicology 8:493-505.

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Feltes, J. and Levsen, K. (1990) Journal of Chromatography 518: 21-40.

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Gatermann, R. et al. (1995) Marine Pollution Bullet 30: 221-27.

Giger, W. and Roberts, P.V. (1978) Water Pollution Microbiology, Mitchell R ed, Wiley: NY, NY 2: 135-75.

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Gomolka, E. and Gomolka, B. (1979) Ability of activated sludge to degrade nitrobenzene in municipal wastewater. Acta Hydrochemistry and Hydrobiology, 7(6): 605-622.

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Kubota, Y. (1979) Ecotoxicology and Environmental Safety 3: 256-68.

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Malaney, G.W. and Mckinney, R.E. (1966) Oxidation abilities of benzene-acclimated activated sludge. Water Sewage Works, 113: 302-9.

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Meijers, A.P. and Vanderleer, R.C. (1976) Water Research 10: 597-604.

Mirsalis, J.C., Tyson, C.K. and Butterworth, B.E. (1982) Detection of genotoxic carcinogens in the in vivo-in vitro hepatocyte DNA repair assay. Environ Mutag 4: 553-562.

Mitsumori, K., Kodama, Y., Uchida, O., Takada, K., Saito, M., Naito, K., Tanaka, S., Kurokawa, Y., Usami, M., Kawashima, K., Yasuhara, K., Toyoda, K., Onodera, H., Furukawa, F., Takahashi, M. and Hayashi, Y. (1994) Confirmation Study, Using Nitrobenzene, of the Combined Repeat Dose and Reproductive/Developmental Toxicity Test Protocol Proposed by the Organization for Economic Cooperation and Development (OECD). The Journal of Toxicological Sciences, 19:141-149.

Morgan, K.T., Gross, E.A., Lyght, O. and Bond, J.A. (1985) Morphologic and biochemical studies of a nitrobenzene-induced encephalopathy in rats. Neurotoxicology 6:105-116.

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Johannsen, F.R., Levinskas, G.J., Rusch, G.M., Terril, J.B. and Schroeder, R.E. (1988) Evaluation of the Subchronic and Reproductive Effects of a Series of Chlorinated Propanes in the Rat. I. Toxicity of 1,2,3-Trichloropropane. J. Toxicol. Environ. Health. 25(3):299-315.

Keith, L.H. et al. (1976) Ident. Anal. Org. Pollut. Water, Keith LH Ed Ann Arbor Press, Ann Arbor MI, pp. 327-73.

Kier, L.D. (1982) Ames/Salmonella Mutagenicity assays of 1,2,3-Trichloropropane, 1,2,2,3- Tetrachloropropane and 1,1,2,2,3-Pentachloropropane. Monsanto Company.

Kolpin, D.W. et al. (2000) Ground Water 38: 858-63.

Konnecker and Schmidt (2003) Environmental risk assessment for 1,2,3-trichloropropane – is there a risk for the aquatic environment? Fresenius Environmental Bulletin, Vol 12 (12).

Kuo, M.C.T. et al. (2000) Bull. Environ. Contam. Toxicol. 65: 654-59.

La, D.K., Schoonhoven, R., Ito, N. and Swenberg, J.A. (1996) The Effect of Exposure Route on DNA Adduct Formation and Cellular Proliferation by 1,2,3-Trichloropropane.

Lag, M., Soderland, E.J., Omichinski, J.G., Brunborg, G., Holme, J.A., Dahl, J.E., Helson, S.D. and Dybing, E. (1991) Effect of Bromine and Chlorine Positioning in the Induction of Renal and Testicular Toxicity by Halogenated Propanes. Department of Environmental Medicine, National Institute of Public Health, Olso, Norway.

Leistra, M. and Boesten, J.J.T.I. (1989) Agric. Ecosys. Environ. 26: 369-89.

Lucas, S.V. (1984) GC/MS Analysis of Organics in Drinking Water Concentrates and Advanced Waste Treatment Concentrates: Vol 3 USEPA-600/1-84-020 (NTIS PB85-128247).

Merrick, B.A., Robinson, M. and Condie, L.W. (1991) Cardiopathathic Effect of 1,2,3- Trichloropropane after Subacute and Subchronic Exposure in Rats. Journal of Applied Toxicology, 11(3):179-187.

Miller, R.R., Quast, J.F. and Gushow, T.S. (1986a) 1,2,3-Tichloropropane: 2-Week Vapor Inhalation Study in Rats and Mice, May 27, 1986. The Dow Chemical Company.

Miller, R.R., Quast, J.F. and Momany-Pfruender, J.J. (1986b) 1,2,3-Tichloropropane: 2-Week Vapor Inhalation Study to Determine the No Adverse-Effect Level in Rats and Mice, May 13, 1986. The Dow Chemical Company.

NTP (1993) Toxicology and Carcinogenesis Studies of 1,2,3-Trichloropropane (CAS No. 96- 18-4) in F344/N Rats and B6C3F1 Mice. Technical Report Series No. 384. National

Toxicology Program. US Department of Health and Human Services. Public Health Services. National Institutes of Health.

OECD (2004) 1,2,3-Trichloropropane. SIDS Initial Assessment Report for SIAM 18, Paris, 20- 23 April 2004. Organization for Economic Cooperation and Development.

Oki, D.S. and Giambelluca, T.W. (1987) Ground Water 25: 693-702.

Rusch, G.M. and Rinehart, W.E. (1979) A 13-Week Inhalation Toxicity Study of 1,2,3- Trichloropropane in the Rat. The Monsanto Company.

Saito-Suzuki, R., Teramoto, S. and Shirasu, Y. (1982) Dominant lethal studies in rats with 1,2- dibromo-3-chloropropane and its structurally related compounds. Mutation Research, 101:321-327.

Sawin, V.L. and Hass, B.S. (1982) Assay of 1,2,3-Trichloropropane (1,2,3-TCP) for Gene Mutation in Mouse Lymphoma Cells. Shell Development Company.

Schafer, R.B., von der Ohe, P.E., Kuhne, R., Schuurmann, G. and Liess, M. (2011) Occurrence and Toxicity of 331 organic pollutants in large rivers of North Gemany over a decade (1994-2004). Environmental Science and Technology, 45: 6167-6174.

Schroeder R.E. and Rinehart W.E. (1980) A One Generation Reproduction-Fertility Study of 1,2,3-Trichloropropane in Rats.

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Tafazoli, M. and Kirsch-Volders, M. (1996) In vitro mutagenicity and genotoxicity study of 1,2- dichloroethylene, 1,1,2-trichloroethylene, 1,3-dichloropropane, 1,2,3-trichloropropane and 1,1,3-trichloropropane, using the micronucleus test and the alkaline single cell gel electrophoresis technique (comet assay) in human lymphocytes. Mutation Research, 37:185- 202.

Terrill, J.B. and Daly, I.W. (1983) A 13-Week Inhalation Toxicity Study of 1,2,3- Trichloropropane in the Rat. The Monsanto Company.

Tratnyek, P.G., Sarathy, V. and Fortuna, J.H. (2008) Fate and Remediation of 1,2,3- Trichloropropane. Remediation of Chlorinated and Recalcitrant Compounds - 2008. Proceedings of the Sixth International Conference on Remediation of Chlorinated and Recalcitrant Compounds.

US EPA (1983) Health and Environmental Effects Profile for Trichloropropane Isomers (External Review Draft) EACO-CIN-P010. U.S. Environmental Protection Agency.

US EPA (2009) Toxicological Review Of 1,2,3-Trichloropropane (Cas No. 96-18-4) In Support of Summary Information on the Integrated Risk Information System (IRIS). September 2009 U.S. Environmental Protection Agency.

Villeneuve, D.C., Chu, I., Secours, V.E., Cote, M.G., Plaa, G.L and Valli, V.E. (1985) Results of a 90-day Toxicity Study on 1,2,3- and 1,1,2-Trichloropropane Administered Via Drinking Water.

Wakeham, S.G. et al. (1983) Can. J. Fish. Aq. Sci. 40: 304-21.

WHO (2003) 1,2,3-Trichloropropane. Concise International Chemical Assessment Document 56. World Health Organization.

Yamamoto, K. et al. (1997) Environmental Pollution 95: 135-43.

Zebarth, B.J. et al. (1998) Water Qual. Res. J Can 33: 31-50.

20.10 Urethane

Agrelo, C. and Amos, H. (1981) DNA repair in human fibroblasts. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. I: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science

Allen, J.W., Langenbach, R., Nesnow, S., Sasseville, K., Leavitt, S., Campbell, J., Brock, K. and Sharief, Y. (1982) Comparative genotoxicity studies of ethyl carbamate and related chemicals: further support for vinyl carbamate as a proximate carcinogenic metabolite. Carcinogenesis, 3(12): 1437-1441.

Altmann, H.-J., Dusemund, B., Goll, M. and Grunow, W. (1991) Effect of ethanol on the induction of lung tumours by ethyl carbamate in mice. Toxicology, 68: 195-201.

Amacher, D.E. and Turner, G.N. (1982) Mutagenic evaluation of carcinogens and noncarcinogens in the L5178Y/TK assay utilizing postmitochondrial fractions (S9) from normal rat liver. Mutation Research, 97: 49-65.

Bateman, A.J. (1967) A failure to detect any mutagenic action of urethane in the mouse. Mutation Research, 4: 710-712.

Barale, R., Scapoli, C., Falezza, A., Ventura, L., Bernacchi, F., Loprieno, N. and Barrai, I. (1992) Skin cytogenetic assay for the detection of clastogens-carcinogens topically administered to mice. Mutation Research, 271(3): 223-230.

Berenblum, I. and Haran-Ghera, N. (1955) The initiating action of ethyl carbamate (urethane) on mouse skin. British Journal of Cancer, 9: 453-456.

Bridges, B.A., MacGregor, D. and Zeiger, E. (1981) Summary report on the performance of bacterial mutation assays. Part I. Background and Summaries. In: de Serres, F.J. and Ashby, J. Eds. Progress in Mutation Research, Vol. I: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science

Burke, D.A., Wedd, D.J., Herriott, D., Bayliss, M.K., Spalding, D.J.M. and Wilcox, P. (1994) Evaluation of pyrazole and ethanol induced S9 fraction in bacterial mutagenicity testing. Mutagenesis, 9(1): 23-29.

CCRIS (2013) Chemical Carcinogenesis Research Information System Available from http://toxnet.nlm.nih.gov/

Cheng, M., Conner, M.K. and Alarie, Y. (1981a) Multicellular in vivo sister-chromatid exchanges induced by urethane. Mutation Research, 88: 223-231.

Cheng, M., Conner, M.K. and Alarie, Y. (1981b) Potency of some carbamates as multiple tissue sister chromatid exchange inducers and comparison with known carcinogenic activities. Cancer Research, 41: 4489-4492.

ChemID (2013) ChemID Plus Advanced. US National Library of Medicine. Available from http://chem.sis.nlm.nih.gov/chemidplus/

Conner, M.K. and Cheng, M. (1983) Persistence of ethyl carbamate-induced DNA damage in vivo as indicated by sister chromatid exchange analysis. Cancer Research, 43: 965-971.

Crebelli, R., Bellincampi, D., Conti, G., Conti, L., Morpurgo, G. and Carere, A. (1986) A comparative study on selected chemical carcinogens for chromosome malsegregation, mitotic crossing-over and forward mutation induction in Aspergillus nidulans. Mutation Research, 172(2): 139-149.

Csukás, I., Gungl, E., Fedorcsák, I., Vida, G., Antoni, F., Turtóczky, I. and Solymosy, F. (1979) Urethane and hydroxyurethane induce sister-chromatid exchanges in cultured human lymphocytes. Mutation Research, 67(4): 315-319.

Csukás, I., Gungl, E., Antoni, F., Vida, G. and Solymosy, F. (1981) Role of metabolic activation in the sister chromatid exchange-inducing activity of ethyl carbamate (urethane) and vinyl carbamate. Mutation Research, 89(1): 75-82.

Dahl, G.A., Miller, J.A. and Miller, E.C. (1978) Vinyl carbamate as a promutagen and a more carcinogenic analog of ethyl carbamate. Cancer Research, 38: 3793-304.

De Jong, G., Van Sittert, N.J. and Natarajan, A.T. (1988) Cytogenetic monitoring of industrial populations potentially exposed to genotoxic chemicals and of control populations. Mutation Research 204, 451-464.

Director, A.E., Tucker, J.D., Ramsey, M.J. and Nath, J. (1998) Chronic ingestion of clastogens by mice and the frequency of chromosome aberrations. Environmental and Molecular Mutagenesis, 32: 139-147.

Dogan, E.E., Yesilada, E., Ozata, L. and Yologlu, S. (2005) Genotoxicity testing of four textile dyes in two crosses of Drosophila using wing somatic mutation and recombination test. Drug and Chemical Toxicology, 28: 289-301.

Dunlap, W.J. et al. (1976a) Ident. Anal. Org. Pollut. Water 453-77.

Dunlap, W.J. et al. (1976b) Organic Pollutants Contributed to Ground Water by a Landfill. Ada, OK: USEPA. USEPA-600/9-76-004.

Evans, E.L. and Mitchell, A.D. (1981) Effects of 20 coded chemicals on sister chromatid exchange frequencies in cultured Chinese hamster cells. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. I: Evaluation of Short-Term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science

FAO/WHO (2005) Joint FAO/WHO Expert Committee on Food Additives. Sixty-fourth meeting Rome, 8-17 February 2005; JECFA/64/SC. Food and Agriculture Organisation and World Health Organization.

FSA (2001) Food Advisory Committee, Annual Report 2000. Food Standards Agency. Crown Copyright, UK.

Frölich, A. and Würgler, F.E. (1990) Genotoxicity of ethyl carbamate in the Drosophila wing spot test: dependence on genotype-controlled metabolic capacity. Mutation Research, 244: 201-208.

Graf, U. and Van Schaik, N. (1992) Improved high bioactivation cross for the wing somatic mutation and recombination test in Drosophila melanogaster. Mutation Research, 271: 59-67.

Hoffler, U., Dixon, D., Peddada, S. and Ghanayem, B.I. (2005) Inhibition of urethane-induced genotoxicity and cell proliferation in CYP2E1-null mice. Mutation research, 572: 58-72.

HSDB (2013) Hazardous Substances Databank. US National Library of Medicine. Available from http://toxnet.nlm.nih.gov/cgi-bin/sis/htmlgen?HSDB

Hübner, P., Groux, P.M., Weibel, B., Sengstag, C., Horlbeck, J., Leong-Morgenthaler, P.-M. and Lüthy, J. (1997) Genotoxicity of ethyl carbamate (urethane) in Salmonella, yeast and human lymphoblastoid cells. Mutation Research, 390(1-2): 11-19.

IARC. (1974) Monographs Volume 7. Some anti-thyroid and related substances, nitrofurans and industrial chemicals. International Agency for Research on Cancer.

IARC (2010) Monographs Volume 96. Alcohol consumption and ethyl carbamate. International Agency for Research on Cancer.

Inai, K., Arihiro, K., Takeshima, Y. et al. (1991) Quantitative risk assessment of carcinogenicity of urethane (ethyl carbamate) on the basis of long-term oral administration to B6C3F1 mice. Jpn J Cancer Res, 82: 380–385.

Itoh, A. and Matsumoto, N. (1984) Organ-specific susceptibility to clastogenic effect of urethane, a trial application of whole embryo culture to testing system for clastogen. The Journal of Toxicological Sciences, 9: 175-192.

Iversen, O.H. (1991) Urethan (ethyl carbamate) is an effective promoter of 7,12- dimethylbenz[a]anthracene-induced carcinogenesis in mouse skin two-stage experiments.Carcinogenesis, 12: 901-903.

Jagannath, D.R., Vultaggio, D.M. and Brusick, D.J. (1981) Genetic activity of 42 coded compounds in the mitotic gene conversion assay using Saccharomyces cerevisiae strain D4. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. I: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science

JECFA (2006) Safety Evaluation of Certain Contaminants in Food, WHO Food Additives Series 55. Joint FAO/WHO Expert Committee on Food Additives.

Jotz, M.M., Mitchell, A.D. (1981) Effects of 20 coded chemicals on the forward mutation frequency at the thymidine kinase locus in L5178Y mouse lymphoma cells. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. I: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science

Kamiguchi, Y. and Tateno, H. (2002) Radiation- and chemical-induced structural chromosome aberrations in human spermatozoa. Mutation Research, 504: 183-191.

Keith, L.H. (1976) Environ Sci Tech 10: 555-64.

Knaap, A.G.A.C. and Kramers, P.G.N. (1982) Absence of synergism between mutagenic treatments, given one generation apart, in Drosophila melanogaster. Mutation Research, 92: 117-121.

Leonskaya, G.I. (1980) Gig. Naselen Mest., 19: 40-43.

Lewis, R.J. (1996) Sax's Dangerous Properties of Industrial Materials. 9th ed. Volumes 1-3. New York, NY: Van Nostrand Reinhold.

Martin, C.N. and McDermid, A.C. (1981) Testing of 42 coded compounds for their ability to induce unscheduled DNA repair synthesis in HeLa cells. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. I: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science.

Martin, C.N., McDermid, A.C. and Garner, R.C. (1978) Testing of known carcinogens and noncarcinogens for their ability to induce unscheduled DNA synthesis in HeLa cells. Cancer Res, 38: 2621–2627.

Matsushima, T., Hayashi, M., Matsuoka, A., Ishidate, M., Miura, K.F., Miura, K.F., Shimizu, H., Suzuki, Y., Morimoto, K., Ogura, H., Mure, K., Koshi, K. and Sofuni, T. (1999) Validation study of the in vitro micronucleus test in a Chinese hamster lung cell line (CHL/IU). Mutagenesis, 14(6): 569-580.

McCann, J., Choi, E., Yamasaki, E. and Ames, B.N. (1975) Detection of carcinogens as mutagens in the Salmonella/microsome test: assay of 300 chemicals. Proceedings of the National Academy of Sciences, USA, 72: 5135-5139.

Nomura, T., Hayashi, T., Masuyama, T. et al. (1990) Carcinogenicity of sublimed urethane in mice through the respiratory tract. Jpn J Cancer Res, 81: 742–746.

NTP. (1996) Urethane in drinking water and urethane in 5% ethanol, Toxicity Report Series No. 52. National Toxicology Program. U.S. Department of Health and Human Services, Public Health Service, National Institutes of Health.

Osaba, L., Aguirre, A., Alonso, A. and Graf, U. (1999) Genotoxicity testing of six insecticides in two crosses of the Drosophila wing spot test. Mutation Research, 439: 49-61.

Osswald, H. (1959) On the question of mitosis-inhibition by polyvalent carbamic acid esters in Ehrlich‘s carcinoma. Arzneimittelforschung, 9: 595-598.

Oster, I.I. (1958) Interactions between ionizing radiation and chemical mutagens. Z Indukt A bstamm Vererbungsl, 89: 1-6.

Otto, H. and Plötz, D. (1966) Experimentelle Tumorinduktion mit Urethan-anaerosolen. Z. Krebsforsch., 68: 284-292.

Pai, V., Bloomfield, S.F. and Gorrod, J.W. (1985) Mutagenicity of N-hydroxylamines and N- hydroxycarbamates towards strains of Escherichia coli and Salmonella typhimurium. Mutation Research, 151: 201-207.

Parry, J.M. and Sharp, D. (1981) Induction of mitotic aneuploidy in the yeast strain D6 by 42 coded compounds. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. I: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science

Pereira, M.A., Khoury, M.M., Glauert, H.P. and Davis, R.A. (1991) Screen of five alkyl carbamates for initiating and promoting activity in rat liver. Cancer Letters, 57: 37-44.

Pero, R.W., Bryngelsson, T., Widegren, B., Hogstedt, B., Welinder, H. (1982) A reduced capacity for unscheduled DNA synthesis in lymphocytes from individuals exposed to propylene oxide and ethylene oxide. Mutation Research, 104, 193-200.

Pero, R.W., Osterman-Golkar, S., Hogstedt, B. (1985) Unscheduled DNA synthesis correlated to alkylation of haemoglobin in individuals occupationally exposed to propylene oxide. Cell Biology and Toxicology, 1, 309-314.

Perry, P.E. and Thompson, E.J. (1981) Evaluation of the sister chromatid exchange method in mammalian cells as a screening system for carcinogens. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. I: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science

Popescu, N.C., Turnbull, D. and DiPaolo, J.A. (1977) Sister chromatid exchange and chromosome aberration analysis with the use of several carcinogens and noncarcinogens. Journal of the National Cancer Institute, 59: 289–293.

Port, R., Schmahl, D. and Wahrendorf, J. (1976) Some examples of dose-response studies in chemical carcinogenesis. Oncology, 33: 66-71.

Richardson, M.L. et al. (1994) The Dictionary of substances and their effects. Vol. 1-7. The Royal Society of Chemistry. Thomas Graham House, The Science Park, Cambridge, CB4 4WF.

Roberts, G.T. and Allen, J.W. (1980) Tissue-specific induction of sister chromatid exchanges by ethyl carbamate in mice. Environ Mutagen, 2: 17–26.

RTECS. (1992) Registry of Toxic Effects of Chemical Substances. National Institute of Occupational Safety and Health, USA.

Salaman, A.H. and Roe, F.J.C. (1953) Incomplete carcinogens: Ethyl carbamate (urethane) as an initiator of skin tumour formation in the mouse. British Journal of Cancer, 7(4): 472-481.

Santini, P., Moretton, J. and d‘Aquino, M. (1985) Detection of the genetic toxicity of diethyl pyrocarbonate using bacterial systems. Rev. Latinoam. Microbiol., 27(2): 157-162.

Schlatter, J. and Lutz, K. (1990) The carcinogenic potential of ethyl carbamate (urethane): risk assessment at human dietary exposure levels. Food and Chemical Toxicology, 28(3): 205- 211.

Simmon, V.F. (1979a) In vitro mutagenicity assays of chemical carcinogens and related compounds with Salmonella typhimurium. Journal of the National Cancer Institute, 62: 893- 899.

Simmon, V.F. (1979b) In vitro assays for recombinogenic activity of chemical carcinogens and related compounds with Saccharomyces cerevisiae D3. J Natl Cancer Inst, 62: 901–909.

Simmon, V.F., Kauhanen, K., Mortelmans, K., Tardiff, R.G. (1977) Mutagenic activity of chemicals identified in drinking water. Mutation Research, 53(2): 261-262.

Sina, J.F., Bean, C.L., Dysart, G.R., Taylor, V.I. and Bradley, M.O. (1983) Evaluation of the alkalin elution/rat hepatocyte assay as a predictor of carcinogenic/mutagenic potential. Mutation Research, 113(5): 357-391

Sirica, A.E., Hwang, C.G., Sattler, G.L. and Pitot, H.C. (1980) Use of primary cultures of adult rat hepatocytes on collagen gel-nylon mesh to evaluate carcinogen-induced unscheduled DNA synthesis. Cancer Research, 40: 3259-3267.

Sofuni, T., Honma, M., Hayashi, M., Shimada, H., Tanaka, N., Wakuri, S., Awogi, T., Yamamoto, K.I., Nishi, Y. and Nakadate, M. (1996) Detection of in vitro clastogens and spindle poisons by the mouse lymphoma assay using the micro well method: interim report of an international collaborative study. Mutagenesis, 11(4): 349-355.

Sotomayor, R.E., Sega, G.A. and Kadlubar, F. (1994) Induction of DNA damage by urethane in mouse testes: DNA binding and unscheduled DNA synthesis. Environ Mol Mutagen, 24: 68–74.

SRC (2013) Physprop Database. Syracuse Research Corporation. Available from http://www.syrres.com/what-we-do/databaseforms.aspx?id=386

SYKE (2013) Databank of environmental chemicals, Urethane. Finnish Environment Institute. Available from http://wwwp.ymparisto.fi/scripts/Kemrek/Kemrek.asp

Takaori, S., Tanabe, K. and Shimamoto, K. (1966) Developmental abnormalities of skeletal system induced by ethylurethan in the rat. Jpn J Pharmacol, 16: 63–73.

Toth, B., Della Porta, G. and Shubik, P. (1961a) The occurrence of malignant lymphomas in urethane-treated Swiss mice. British Journal of Cancer, 15: 322-326.

Toth, B., Tomatis, L. and Shubik, P. (1961b) Multipotential carcinogenesis with urethan in the Syrian golden hamster. Cancer Research, 21: 1537-1541.

Topaktas, M., Rencüzoğullari, E. and Ila, H.B. (1996) In vivo chromosomal aberrations in bone marrow cells of rats treated with Marshal. Mutation Research, 371: 259-264.

US EPA (2013) Chemical Contaminant List 3 (CCL 3) United States Environmental Protection Agency. Available from http://www.epa.gov/ogwdw/ccl/ccl3.html

Vogt, M. (1948) Mutationsauslösung bei Drosophila durch Athylurethan. Experientia, 4: 68– 69.

Von der Hude, W., Kalweit, S., Engelhardt, G., McKiernan, S., Kasper, P. and Slacik-Erben, R. (2000) In vitro micronucleus assay with Chinese hamster V79 cells- results of a collaborative study with in situ exposure to 26 chemical substances. Mutation Research, 468: 137-163.

Watanabe, K., Sakamoto, K. and Sasaki, T. (1998) Comparisons on chemically-induced mutation among four bacterial strains, Salmonella typhimurium TA102 and TA2638, and Escherichia coli WP2/pKM101 and WP2 uvrA/pKM101: Collaborative study II. Mutation Research, 412(1): 17-31.

Westmoreland, C., Plumstead, M. and Gatehouse, D. (1991) Activity of urethane in rat and mouse micronucleus tests after oral administration. Mutation Research, 262: 247-251.

Wild, D. (1978) Cytogenetic effects in the mouse of 17 chemical mutagens and carcinogens evaluated by the micronucleus test. Mutation Research, 56: 319-327.

Wyrobek, A.J. and Bruce, W.R. (1975) Chemical induction of sperm abnormalities in mice. Proceedings of the National Academy of Science, USA, 72: 4425-4429.

Zeiger, E., Anderson, B., Haworth, S., Lawlor, T. and Mortelmans, K. (1992) Salmonella mutagenicity tests. V. Results from the testing of 311 chemicals. Environmental Molecular Mutagenesis, 19(21): 2-141.

Zimmermann, F.K. and Scheel, I. (1981) Induction of mitotic gene conversion in strain D7 of Saccharomyces cerevisiae by 42 coded compounds. In: de Serres, F.J. and Ashby, J, eds, Progress in Mutation Research, Vol. 1: Evaluation of Short-term Tests for Carcinogens. Report of the International Collaborative Program, New York, Elsevier Science.

20.11 Ethylene oxide

Agurell, E., Cederberg, H., Ehrenberg, L., Lindahl-Kiessling, K., Rannug, U. and Torngvist, M. (1991) Genotoxic effects of ethylene oxide and propylene oxide: a comparative study. Mutation Research, 250: 229-237.

Anand, V.P., Cogdill, C.P., Klausner, K.A., Lister, L., Barbolt, T., Page, B.F., Urbanski, P., Woss, C.J. and Boyce, J. (2003) Reevaluation of ethylene oxide hemolysis and irritation potential. Journal of Biomedical Materials Research Part A, 64(4): 648-654.

ARTG (2013) Database Austrailian Register of Therapeutic Goods Database Website. Available from http://www.tga.gov.au/industry/artg.htm

ATSDR (1990) Toxicological profile for ethylene oxide. U.S. Department of Health and Human Services. Agency for Toxic Substances & Disease Registry.

Australia Pesticides and Veterinary Medicines Authority (2013) Public Chemical Registration Information System Search. Available from https://portal.apvma.gov.au/pubcris.

Bird, M.l. (1952) Chemical production of mutations in Drosophila: comparison of techniques. J. Genet., 50; 480-485.

Bogyo, D.A., Lande, S.S., Meylan, W.M., Howard, P.H. and Santodonato, J. (1980) Investigation of selected potential environmental contaminants: Epoxides. Final technical report. New Jersey (NJ): Syracuse Research Corporation, Center for Chemical Hazards Assessment. Report No.: EPA-560/11-80-005.

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ChemID (2013) ChemID Plus Advanced. US National Library of Medicine. Available from http://chem.sis.nlm.nih.gov/chemidplus/

De Flora, S. (1991) Study of 106 organic and inorganic compounds in the Salmonella/microsome test. Carcinogenesis, 2: 283-298.

Deichmann, W.B. (1969) (ed.) Toxicology of drugs and chemicals, Acad. Press N.Y. p. 258.

Dellarco, V.L. et al. (1990) Environmental and Molecular Mutagenesis, 16: 85-103.

Dunkelberg, H. (1982) Carcinogenicity of ethylene oxide and 1,2-propylene oxide upon intragastric administration to rats. Br. J. Cancer, 46, 924-933.

EC (2000) International Uniform ChemicaL Information Database (CD-ROM - 2000). Ethylene Oxide. European Chemicals Bureau, European Commission - JRC, Environment Institute, Ispra, Italy.

EU (2013) EU Pesticides Database. Available from http://ec.europa.eu/sanco_pesticides /public/index.cfm?event=activesubstance.detail.

Embree, L.W, Lyon, l.P. and Hine, C.H. (1977) The mutagenic potential of ethylene oxide using the dominant-Iethal assay in rats. Toxicol. appl. Pharmacol., 40, 261-267.

Entomological Society of America (1978) Special Publication, Vol. 78-1, 17.

Environment Canada/Health Canada (2001) Priority Substances List Assessment Report Canadian Environmental Protection Act, 1999, Ethylene Oxide.

Fahmy, O.G. and Fahmy, M.J. (1970) Gene elimination in carcinogenesis: reinterpretation of the somatic mutation theory. Cancer Res., 30, 195-205.

Farooqi, Z., Törnqvist, M., Ehrenberg, L. and Natarajan, AT (1993) Genotoxic effects of ethylene oxide and propylene oxide in mouse bone marrow cells. Mutat. Res., 288, 223-228.

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20.12 Formaldehyde

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Feron, V.J., Bruyntjes, J.P., Woutersen, R.A., Jmmel, H.R. and Appelman, L.M. (1987) Nasal tumours in rats after short-term exposure to a cytotoxic concentration of formaldehyde, Zeist, Netherlands, CIVO Institute TNO, 17 pp (Report No. V87.167/130347).

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LaVoie, E.J., Adams, E.A., Shigematsu, A. and Hoffmann, D. (1983) On the metabolism of quinoline and isoquinoline: possible molecular basis for differences in biological activities. Carcinogenesis, 4(9): 1169-1173.

LaVoie, J.E., Shigematsu, A., Adams, E.A., Rigotty, J. and Hoffmann, D. (1984) Tumor- initiating activity of quinoline and methylated quinolines on the skin of Sencar mice. Cancer letters 22:269-273.

LaVoie, E.J. et al. (1988) Quinolines and benzoquinolines: studies related to their metabolism, mutagenicity, tumour-initiating activity and carcinogenicity. In: polynuclear aromatic hydrocarbons: a decade of progress, Cooke M. and Daniels A.J. (eds), 10th International Symposium, October 21-23. Battelle Press.

Lewis, R.J. (1996) Sax‘s Dangerous Properties of Industrial Materials, 9th Edition. Volumes 1- 3. Van Nostrand Reinhold.

Liao, P., Yuan, S., Zhang, W., Tong, M. and Wang, K. (2012) Mechanistic aspects of nitrogen- heterocyclic compound adsorption on bamboo charcoal. Journal of Colloid and Interface Science, 382(1): 74-81.

Matsumoto, T.. et al. (1978) Agric. Biol. Chem. 42: 861-864.

Matsuoka, A., Hayashi, M. and Ishidate, M. (1979) Chromosomal aberration tests on 29 chemicals combined with S9 mix in vitro. Mutat Res. 66(3): 277-90.

McFee, A.F. (1989) Genotoxic Potency of Three Quinoline Compounds Evaluated In Vivo in Mouse Marrow Cells. Environmental and Molecular Mutagenesis, 13: 325-331.

McGregor, D.B., Brown, A., Cattanach, P., Edwards, I., McBride, D., Riach, C. and Caspary, W.J. (1988) Responses of the L5178Y tk+/tk- mouse lymphoma cell forward mutation assay: III. 72 coded chemicals. Environ Mol Mutagen. 12(1): 85-154.

Meijers, A.P. and VanderLeer, R.C. (1976) Water Research 10: 597-604.

Miyata, Y., Fukushima, S., Hirose, M., Masui, T. and Ito, N. (1985) Short-term screening of promoters of bladder carcinogenesis in N-butyl-N-(4-hydroxybutyl)nitrosamine initiated, unilaterally ureter-ligated rats. Jpn J Cancer Res (Gann), 76:828-834.

Nagao, M., Yahagi, T., Seino, Y., Sugimura, T. and Ito, N. (1977) Mutagenicities of quinoline and its derivatives. Mutat. Res. 42, 335-342.

Neuwoehner, J., Reineke, A-K., Hollender, J., Eisentraeger, A. (2009) Ecotoxicity of quinoline and hydroxylated derivatives and their occurrence in groundwater of a tar-contaminated field site. Ecotoxicology and Environmental Safety 72: 819–827.

Ondrus, M.G. and Steinheimer, T.R. (1990) Journal of Chromatography Science 28: 324-30.

Sandhowe-Grote D. (1979) Ermittlung der LD50 von Produckt 110831 (Chinolin, rein) an maennlichen und weiblichen Ratten nach oraler Applikation, Auftragsnummer 337a, Huntingdon Research Centre, Deutschland, Muenster.

San Sebastian J.R. and Hsie A.W. (1982) Environm. Mutagen. 4, 395.

Sasaki, S. (1978) The scientific aspects of the chemical substance control law in Japan. Aquatic Pollutants: Transformation and biological effects. Hutzinger, O., Von Letyoeld, L. H. and Zoeteman, B. C. J. (Editors) Oxford: Pergamon Press, 283-98.

Schehrer, L., Regan, J.D. and Westendorf, J. (2000) UDS induction by an array of standard carcinogens in human and rat hepatocytes: effects of cryopreservation.

Seixas, G.M., Andon, B.M., Hollingshead, P.G. and Thilly, W.G. (1982) The aza-arenes as mutagens for Salmonella typhimurium. Mutat. Res. 102: 201-212.

Shinohara, Y., Ogiso, T., Hananouchi, M., Nakanishi, K., Yoshimura, T. and Ito, N. (1977) Effect of various factors on the induction of liver tumors in animals by quinoline. Gann, 68: 785-796.

Sideropoulos, A.S. and Specht, S.M. (1984) Evaluation of microbial testing methods for the mutagenicity of quinoline and its derivatives. Current Microbiology, 11: 59-65

Sina J.F., Bean C.L., Dysart G.R., Taylor V.I. and Bradley M.O. (1983) Evaluation of the alkaline elution/rat hepatocyte assay as a predictor of carcinogenic/mutagenic potential. Mutat Res. 113(5): 357-91.

Singer, P.C. et al (1979) Treatability and Assessment of Coal Conversion Wastewater: Phase 1 USEPA-600/7-79-249.

Smyth H.F. et al. (1951) Arch. Ind. Hyg. Occup. Med. 4: 119-122.

SRC (2013) Physprop Database. Syracuse Research Corporation. Available from http://www.syrres.com/what-we-do/databaseforms.aspx?id=386

SRC (2013) Biodeg Database. Syracuse Research Corporation. Available from http://www.srcinc.com/what-we-do/databaseforms.aspx?id=382

Stuermer, D.H. et al (1982) Environmental Science and Technology 16: 582-7.

Suzuki, T., Miyata, Y., Saeki, K., Kawazoe, Y., Hayashi, M. and Sofuni, T. (1998) In vivo mutagenesis by the hepatocarcinogen quinoline in the lacZ transgenic mouse: evidence for its in vivo genotoxicity. Mutation Research 412: 161-166.

Suzuki, T., Takeshita, K., Saeki, K., Kadoi, M., Hayashi, M. and Sofuni, T. (2007) Clastogenicity of quinoline and monofluorinated quinolones in Chinese hamster lung cells. Journal of Health Science, 53(3): 325-328.

Suzuki H., Takasawa H., Kobayashi K., Terashima Y., Shimada Y., Ogawa I., Tanaka J., Imamura T., Miyazaki A. and Hayashi M. (2009) Evaluation of a liver micronucleus assay with 12 chemicals using young rats (II): a study by the Collaborative Study Group for the Micronucleus Test/Japanese Environmental Mutagen Society-Mammalian Mutagenicity Study Group. Mutagenesis, 24(1): 9-16

US EPA (2001) Toxicological Review of Quinoline (CAS No. 91-22-5) In Support of Summary Information on the Integrated Risk Information System (IRIS). U.S. Environmental Protection Agency.

Wang, X., Huang, X., Zuo, C. and Hu, H. (2004) Kinetics of quinoline degradation by O3/UV in aqueous phase. Chemosphere, 55(5): 733-741.

Willems, M.I., Dubois, G., Boyd, D.R., Davies, R.J., Hamilton L., McCullough, J.J. and van Bladeren, P.J. (1992) Comparison of the mutagenicity of quinoline and all monohydroxyquinolines with a series of arene oxide, trans-dihydrodiol, diol epoxide, N-oxide and arene hydrate derivatives of quinoline in the Ames/Salmonella microsome test. Mutat. Res. 278: 227-236.

Williams, G.M., Mori, H. and McQueen, C.A. (1989) Structure-activity relationships in the rat hepatocyte DNA-repair test for 300 chemicals. Mutat Res. 221(3): 263-86.

Worstmann, W. (1981) Akute Toxizitaet von Produkt-Nr. 110.831 (Chinolin, rein) an maennlichen und weiblichen Ratten nach dermaler Applikation, Auftragsnummer 875/4/81, Huntingdon Research Centre, Deutschland, Muenster.

Zhang, W.B., An, T.C., Xiao, X.M., Fu. J.M., Sheng G.Y. and Cui M.C. (2003) Photochemical degradation performance of quinoline aqueous solution in the presence of hydrogen peroxide. Journal of Environmental Science and Health - Part A Toxic/Hazardous Substances and Environmental Engineering, 38(11): 2599-2611.

20.15 Surface and Groundwater Model Development

ATSDR (1990) Toxicological profile for 1,1-dichloroethane. Agency for Toxic Substances and Disease Registry.

ATSDR (1999) Toxicological Profile for Formaldehyde. Agency for Toxic Substances and Disease Registry.

ATSDR (2000) Toxicological Profile for Methylene Chloride. Agency for Toxic Substances and Disease Registry.

Bellamy, P.H., P.J., Loveland, R.I., Bradley, R.M., Lark and G.J.D. Kirk. (2005) Carbon Losses from all soils across England and Wales 1978-2003. Nature, Vol. 437, p245-248.

Benson, B. (2003) Concise International Chemical Assessment Document 51 for 1,1- dichloroethene (vinylidene chloride) United States Environmental Protection Agency (USEPA) report prepared under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

Centres for Disease and Control of Pollution (1978) Criteria for a Recommended Standard: Occupational Exposure to Benzyl Chloride. DHHS (NIOSH) Publication Number 78-182.

Chen WR, Sharpless CM, Linden KG and Suffet IH, (2006) Treatment of volatile organic chemicals on the EPA Contaminant Candidate List using ozonation and the O3/H2O2 advanced oxidation process. Environ Sci Technol, 40, (8), 2734-2739.

Dezham, P. (2012) Regional Sewerage System Pretreatment Program Annual Report Fiscal Year 2011-2012. Inland Empire Utilities Agency, PO Box 9020, Chino Hills, California.

Dewalle, F., and Chian, E. (1979) Presence of priority pollutants in sewage and their removal in sewage treatment plants, First Annual Report of the Municipal Environmental Research Laboratory , US Environmental Protection Agency, Cincinnati .

EC (2013) Manufacturing statistics - NACE Rev. 2. Data from April 2013. Statistical classification of economic activities in the European Community (NACE) European Commission. Available from http://epp.eurostat.ec.europa.eu/statistics_explained/index.php/Manufacturing_statistics_- _NACE_Rev._2

ECHA (2002) Methyloxirane (Propylene oxide). CAS No: 75-56-9 INECS No: 200-879-2. Summary Risk Assessment Report. Special Publication I.02.129.

Environment Agency (2006) Remedial Targets Methodology – Hydrogeological Risk Assessment for Land Contamination. Product code: GEHO0706BLEQ-E-E Published by Environment Agency, Bristol.

Environment Agency (2011) H1 Annex D-Basic Surface water discharges Report GEHO0810BSXL-E-E v2.2 published by the Environment Agency, Horizon House, Deanery Road Bristol BS1 5AH.

Environment Agency (2013) Groundwater Protection: Principles and Practice (GP3). Part 4 – Position Statements and Legislation. Revision 4 – Draft for Consultation. 2 Published by the Environment Agency, Horizon House, Deanery Road Bristol BS1 5AH.

Environment Australia (1999) Emission estimation techniques for sewage and wastewater treatment, National Pollutant Inventory, Queensland Environmental Protection Agency Report on behalf of the Commonwealth Government.

Environment Canada/Health Canada (2001) Priority Substances List Assessment Report: Ethylene Oxide. Canadian Environmental Protection Act.

Environment Canada and Health Canada (2009) Screening Assessment for the Challenge Benzene, (chloromethyl)-(Benzyl chloride). Chemical Abstracts Service Registry Number 100- 44-7.

EPA, (2009) Water treatment technology feasibility support document for chemical contaminants for the second six-year review of national primary drinking water regulations. USEPA Report EPA 815-B-09-007.

EPA, (2013) Technical Fact Sheet - 1,2,3-Trichloropropane (TCP). Available from http://www.epa.gov/fedfac/pdf/technical_fact_sheet_tcp_january2013.pdf

Environmental Protection Agency (1979) Sources of toxic pollutants found in influents to sewage treatment plants - V. Hartford Water Pollution Control Plant, Hartford, Connecticut.

EU (2002a) 1,3-Butadiene. EU Risk Assessment Report. Final Report. Institute for Health and Consumer Protection. European Chemicals Bureau.

EU (2002b) Methyloxirane (Propylene oxide). EU Risk Assessment Report. Final Report. Institute for Health and Consumer Protection. European Chemicals Bureau.

European Chemicals Bureau (ECB) (2004) European Union Risk Assessment for aniline. European Commission Report EUR 21092

FAO/WHO (2005) Joint FAO/WHO Expert Committee on Food Additives. Sixty-fourth meeting Rome, 8-17 February 2005; JECFA/64/SC. Food and Agriculture Organisation and World Health Organization.

Forrest. M.J. (1999) Chemical Characterisation of Polyurethanes. RAPRA Technology Ltd.

FSA (2001) Food Advisory Committee, Annual Report 2000. Food Standards Agency. Crown Copyright, UK.

HSDB (2013) Hazardous Substances Databank. US National Library of Medicine. Available from http://toxnet.nlm.nih.gov/cgi-bin/sis/htmlgen?HSDB

Huling SG and Pivetz BE, (2006) Engineering Issue: In-Situ Chemical Oxidation. USEPA Report EPA 600-R-06-072.

IPCS (2002) Concise International Chemical Assessment Document 40 Formaldehyde. World Health Organization.

Kielhorn, J., Könnecker, G., Pohlenz-Michel, C., Schmidt, S., and Mangelsdorf, I. (2003) Concise International Chemical Assessment for 1,2,3-trichloropropane. Document 56 prepared by the Fraunhofer Institute of Toxicology and Aerosol Research, Drug Research and Clinical Inhalation, Hanover, Germany under the joint sponsorship of the United

Kuo, C.H. (1985) Reactions of ozone with organics in aqueous solution. USEPA Report EPA/600/3-85/031.

Langlais, B., Reckhow, D.A. and Brink, D.R. (Eds), (1991) Ozone in Water Treatment. Lewis Publishers, Chelsea, Michigan.

Lawrence, S.J. (2006) Description, properties, and degradation of selected volatile organic compounds detected in ground water — A Review of Selected Literature: Atlanta, Georgia, U. S. Geological Survey, Open-File Report 2006-1338, 62 p., a Web-only publication available from http://pubs.usgs.gov/ofr/2006/1338/.

Löffler, F.E., J.E. Champimi, K.M. Ritalahti, S.J. Sprague, and J.M. Tiedje. (1997) Complete reductive dechlorination of 1,2-dichloropropane by anaerobic bacteria. Appl. Environ. Microbiol. 63(7), 2870-2875.

Marine Programs (2010) Saanich Peninsula Treatment Plant, Wastewater and Marine Environment Program, Annual Report 2009. Report prepared on behalf of Capital Regional District, 625 Fisgard Street, Victoria, British Columbia, Canada, V8W 2S6.

Melcer, H., Bell, J. and Thompson, D. (1992) Predicting the fate of volatile organic compounds in municipal wastewater treatment plants. Water Science and Technology, 25(4- 5), 383-389.

Meyer, S. (2013) South Bay water reclamation plant & ocean outfall annual pretreatment report from 1 January to 31 December, 2012. Environmental Monitoring and Technical Services Public Utilities Department, March 2013.

Middala, S., Campbell, S., Olea, C., Scruggs, A. and Hasson, A.S. (2011) Kinetics and mechanism of the reaction of propylene oxide with chlorine atoms and hydroxy radicals. Int J Chem Kin, 43, (9), 507-521.

Munter, R., Preis, S., Kamenev, S. and Siirde, E. (1993) Methodology of ozone introduction into water and wastewater treatment. Oz Sci Eng, 15, 149-165.

National Center for Biotechnology Information (NCBI) (2004a) Quinoline - Compound Summary (CID 5641). National Center for Biotechnology Information, National Library of Medicine, Building 38A Bethesda, MD 20894.

National Center for Biotechnology Information (NCBI) (2004b) Urethane - Compound Summary (CID 5641). National Center for Biotechnology Information, National Library of Medicine, Building 38A Bethesda, MD 20894.

Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

NERC (2007) Countryside Survey. http://www.countrysidesurvey.org.uk

NTP (2011) Urethane. Report on Carcinogens, Twelfth Edition. US National Toxicology Program.

OECD (2003) 1,2-Dichloropropane. SIDS Initial Assessment Report for SIAM 17, Italy, 11-15 November 2003. Organization for Economic Cooperation and Development.

OECD (2004) SIDS Initial Assessment Report on o-Toluidine CAS No: 95-53-4. Organization for Economic Cooperation and Development.

ONS (2013) Office for National Statistics, UK. Available from http://www.statistics.gov.uk/hub/index.html

Owens, T.W. (2012) Annual report for wastewater treatment works / wastewater collection system fiscal year 2011-2012 (Caroline Beach). North Carolina Department of Environment and Natural Resources., Division of Water Quality, Compliance and Enforcement Unit, Raleigh, North Carolina.

Ramanand, K. (2011) Large Scale Ex-Situ Biotreatment of Source Material at a Superfund Site Impacted with Chlorinated and Non-Chlorinated Chemicals. Preprint, Battelle's 2011 Bioremediation and Sustainable Environmental Technologies Symposium, Reno, NV, June 27-30, 2011.

Reungost, J., Pic, J.S., Manero, M.H. and Debellefontaine, H. (2010) Oxidation of nitrobenzene by ozone in the presence of faujasite zeolite in a continuous flow gas-liquid-solid reactor. Wat Sci Tech, 62, (5), 1076-83.

Scientific Committee on Toxicity, Ecotoxicity and the Environment (SCTEE) (2003) Opinion on the results of the Risk Assessment of Aniline - Environmental part CAS No: 62-53-3 EINECS No: 200-539-3. Adopted by the CSTEE during the 38th plenary meeting of 12 June 2003.

Slavich, F.E., Tetla, R.A., and Zimmer, A.T. (1988) Hazardous waste and wastewater characterisation survey, Columbus AFB MS, USA FOEHL Report, 88-076EQ0040FHH.

Smit, C.E. (2010) Environmental risk limits for benzyl chloride and benzylidene chloride. The Netherlands National Institute for Public Health and the Environment (Dutch: Rijksinstituut voor Volksgezondheid en Milieu (RIVM) Report 601714016/2010.

Stepek, J. (2009) Groundwater information sheet 1,2,3Trichloropropane (TCP) State Water Resources Control Board, Division of Water Quality, GAMA Program.

Stover, E.L. and Kincannon, D.F. (1983) Biological treatability of specific organic compounds found in chemical industry wastewaters. Journal of the Water Pollution Control Federation, 55, 97-105.

Sun, Z., Ma, J., Wang, L. and Zhao, L. (2005) Degradation of nitrobenzene in aqueous solution by ozone-ceramic honeycomb. J Environ Sci, 17, (5), 716-721.

Tabak, H., Quave, S.A., Mashni, C.I, and Barth, E.F. (1981) Biodegradability studies with organic priority pollutant compounds. Journal of the Water Pollution Control Federation, 53:1503–1518.

Takatsuki, M., Hayashi, K., Miura, C., Nosaka, M., Funahashi, N., and Yamane, S. (2004) Hazard assessment for ethylene oxide. Report, Version 1.0, Number 36 compiled by Chemicals Evaluation and Research Institute (CERI) Japan, and National Institute of Technology and Evaluation (NITE), Japan on behalf of New Energy and Industrial Technology Development Organisation (NEDO), published by Daiichi Hoki, Tokyo, 2002.

UK National Ecosystem Assessment (2011) The UK National Ecosystem Assessment: Synthesis of the Key Findings. UNEP-WCMC, Cambridge.

United Nations Environment Programme (UNEP) (1998) Screening Information Data Sets (SIDS). Initial Assessment Report for Benzyl Chloride. UNEP Publications on behalf of Organisation for Economic Co-operation and Development (OECD).

United Nations Environment Programme (UNEP) (2004) Screening Information Data Sets (SIDS) Initial Assessment Report for o-Toluidine. UNEP Publications on behalf of Organisation for Economic Co-operation and Development (OECD).

U.S. Environmental Protection Agency USEPA (2010) Screening level hazard characterisation. Hazard Characterization Document.

US EPA/SRC (2011) EpiSuite version 4.1. US Environmental Protection Agency Office of Pollution Prevention Toxics and Syracuse Research Corporation.

Verschueren, K. (1996) Handbook of Environmental Data on Organic Chemicals, Van Nostrand Reinhold, 3rd Edition.

von Gunten, U. (2003) Ozonation of drinking water: Part I. Oxidation kinetics and product formation. Wat Res, 37, (7), 1443-1467.

Wang, X., Huang, X., Zuo, C. and Hu, H. (2004) Kinetics of quinolone degradation by O3/UV in aqueous phase. Chemosphere, 55, (5), 733-741.

WHO (1998) Concise International Chemical Assessment Document 7 - o-Toluidine, World Health Organization

WHO (2003a) Environmental Health Criteria 230 Nitrobenzene. Published by International Programme on Chemical Safety (IPCS) and Inter-Organization Programme for the Sound Management of Chemicals (IOMC).

WHO (2003b) Environmental Health Criteria 54 Ethylene oxide. Published by International Programme on Chemical Safety (IPCS) and Inter-Organization Programme for the Sound Management of Chemicals (IOMC).

WHO (2003c) Environmental Health Criteria 40 Formaldehyde. Published by International Programme on Chemical Safety (IPCS) and Inter-Organization Programme for the Sound Management of Chemicals (IOMC).

Young, D. (2012) Salmon Creek Wastewater Treatment Plant Annual Pretreatment Report, Submittal Satisfies Section S6 F. of Waste Discharge Permit #WA-002363-9 for Clark County, Clark Regional Wastewater District and the City of Battle Ground

Zhang, J. (2012) Control of emerging contaminants by granular activated carbon and the impact of natural organic matter. University of Toronto. Available from https://tspace.library.utoronto.ca/bitstream/1807/32643/3/Zhang_Juan_201206_MASc_Thesis .pdf

20.16 Exposure Modelling

Bogen, K.T. (1994) Models based on steady-state in vitro dermal permeability data underestimate short-term in vivo exposure to organic chemicals in water. J Expos. Anal. Environ. Epidemiol. 4:457-476

Hamelin, G., Charest-Tardif, G. and Tardif, R. (2009) Physiologically-based pharmacokinetic (PBPK) modeling for liver tumour development from human drinking water exposure to dichloromethane. Cited in Health Canada (2011).

Health Canada (2011) Guidelines for Canadian Drinking Water Quality. Guideline Technical Document. Dichloromethane. Available from www.healthcanada.gc.ca

Krishnan, K. (2004) Development of a two-tier approach for evaluating the relevance of multi- route exposures in establishing drinking water goals for volatile organic chemicals. Cited in Health Canada (2011).

Krishnan, K. and Carrier, R. (2008) Approaches for evaluating the relevance of multiroute exposures in establishing guideline values for drinking water contaminants. J. Environ Sci. Health C. 26:300-316

McKone, T.E. and Knezovich, J.P. (1991) The transfer of trichloroethylene from shower to indoor air: Experimental measurements and implications. J. Air Waste Manage. Assoc. 41: 832-837

US EPA (2000) Volatilization Rates from Water to Indoor Air Phase II. National Center for Environmental Assessment Washington Office. Office of Research and Development. U.S. Environmental Protection Agency.

WHO (2000) Chapter 4: Chemical Hazards. In: Guidelines for Safe Recreational-water Environments. Vol 2: Swimming Pools, Spas and Similar Recreations-water Environments. World Health Organization.