Microbial and photolytic degradation of benzothiazoles in ... · Microbial and photolytic...

Microbial and photolytic degradation of benzothiazoles in water and wastewater

vorgelegte Dissertation von

M.Sc. Hafida Kirouani-Harani

Aus Algerien

von der Fakultät III der Technischen Universität Berlin

zur Erlangung des akademischen Grades

-Doktorin der Naturwissenschaften-

- Dr. rer. nat. –

genehmigte Dissertation

Promotionsausschuss: Vorsitzender: Prof. Dr. Möser Gutachter: Prof. Dr.-Ing. M. Jekel Gutachter: Prof. Dr. rer. nat. W. Rotard Gutachter: PD Dr. T. Reemtsma

Tag der wissenschaftlische Aussprache: 06. Juni 2003

Berlin 2003

D 83

2

ACKNOWLEDGMENTS

Professor Dr. Martin Jekel is gradually acknowledged for his agreement to supervise this

project.

I do sincerely acknowledge the help and assistance given to me by Dr.-Ing. Thorsten Reemtsma

as supervisor while elaborating this project, for support and participation in various parts of

this study and critical discussion of this work.

The financial support by DAAD/GTZ, the German Academic Exchange Service (Deutscher

Akademischer Austausch Dienst and the German Agency for Technical Cooperation (Deutsche

Gesellschaft für Technische Zusammenarbeit), are gradually acknowledged particularly Mrs.

Martina Labrens and Mr. Wiegand.

Jutta Jakobs deserves special thanks for her help, her moral support and encouragement

during down periods.

The reparation of my Computer by Hans Rietdorf was greatly appreciated.

I would like to thank all colleagues of the department of water quality control, Technical

University of Berlin (Fachgebiet Wasserreinhaltung, TU Berlin) for the harmonic work

atmosphere namely J. Jakobs, Dr. A. Putschew, K. von Nordheim, U. Förster, J. Friedrich, T.

Storm and all those who have contributed to the fulfilment of the present project.

My husband and my children Randa and Mourad are acknowledged for her comprehension.

The last but not the least to be thanked is Dr. Anke Putschew for helpful discussions during the

course of this work, for her German linguistic supervising and for her being present whenever

I needed her.

3

ABSTRACT

In the present study, an analytical method for the determination and extraction of benzothiazoles in water and wastewater was developed. Because bacteria are important agents in transforming most organic compounds, their specific roles in biotransformation of benzothiazole compounds in water was explored. The biodegradability of these compounds by using aerobic and anaerobic conditions was investigated. Additionally, the photochemical degradation of benzothiazoles was studied. The polar benzothiazole compounds such as benzothiazole sulfonic acid (BTSA) and hydroxybenzothiazole (OHBT) are well recovered from the polar polymer phases. The other analytes show good recoveries on the silicagel-cartridges. Based on their high capacity, RP 18 and EN were chosen for a sequential solid phases-extraction of benzothiazoles. The recovery rates of the solid-phase extraction (SPE) from sewage treatment plants effluents and influents are slightly decreased compared to those obtained with tap water but remained acceptable and are full satisfied. Aerobic mixed cultures are capable of utilising benzothiazoles as a sole source of carbon, nitrogen and energy. Benzothiazole (BT), hydroxybenzothiazole (OHBT), mercaptobenzothiazole (MBT) and methylthiobenzothiazole (MTBT) were investigated in separate batch tests. Benzothiazole and hydroxybenzothiazole can be microbially degraded in water in a relatively short time, whereas methylthiobenzothiazole was proved to be stable under aerobic conditions. Mercaptobenzothiazole was methylated to methylthiobenzothiazole. By using the Sequencing Batch Reactor (SBR), mercaptobenzothiazole was transformed to benzothiazole sulfonic acid (BTSA), methylthiobenzothiazole (MTBT) and traces of dithiobisbenzothiazole (MBTS). Anaerobic treatment revealed no transformation of hydroxybenzothiazole whereas, methylthiobenzothiazole was partly degraded. Mercaptobenzothiazole was partly metabolised to BT and BTSA was partially transformed to OHBT, which was not observable earlier than 5 weeks together with other unknown substances that were not identified. Concerning thiocyanomethylthiobenzothiazole (TCMTB), BT is beside MBT an important metabolite. Under photolytic conditions in water and wastewater, both mercaptobenzothiazole and methylthiobenzothiazole revealed the formation of two stable photoproducts: benzothiazole (BT) and hydroxybenzothiazole (OHBT).

ZUSAMMENFASSUNG

Im Rahmen dieser Arbeit wurde eine analytische Methode für den Nachweis von Benzothiazolen entwickelt. Da Bakterien oft eine große Rolle bei der Transformation von organischen Stoffen in Wasser spielen, wurde der Bio-Abbau (anaerob/aerob) der

4

Verbindungen untersucht. Desweitern wurde das photolytische Abbauverhalten der Verbindungen untersucht. Die polaren Benzothiazole wie Benzothiazol Sulfonsäure (BTSA) und Hydroxybenzothiazol (OHBT) können mit guten Wiederfindungsraten über polare Polymer-Phasen (EN-Phase) extrahiert werden. Die andere Analyten zeigten gute Wiederfindungen auf Silicagel-Kartuschen. Aufgrund der hohen Kapazität wurden RP 18 und EN Material für eine sequentielle Fest-Phasen-Extraktion der Benzothiazole gewählt. Geringfügig niedrigere Wiederfindungsraten wurden für die Extraktion der Benzothiazole aus komplexer Matrix (Kläranlagenzu- und ablauf) festgestellt. Mischkulturen aerober Bakterien können Benzothiazole als Kohlenstoff-, Stickstoff- und Energiequelle nutzen. In separaten Batch-Tests mit Mischkulturen aerober Bakterien wurde das Verhalten von Benzothiazol (BT), Hydroxybenzothiazol (OHBT), Mercaptobenzothiazol (MBT) und Methylthiobenzothiazol (MTBT) untersucht. BT und OHBT können in wässrigen Medien mikrobiell in verhältnismäßig kurz Zeit abgebaut werden. MTBT erwies sich unter analogen Bedingungen als stabil. MBT wird unter aeroben Bedingungen zu MTBT methyliert. In einem Sequencing Batch Reactor (SBR) konnten Benzothiazol Sulfonsäure (BTSA), Methylthiobenzothiazol (MTBT) und geringe Mengen an Dithiobisbenzothiazol (MBTS) nachgewiesen werden. Anaerobe Behandlungen zeigten keine Transformation von Hydroxybenzothiazol, während Methylthiobenzothiazol zum Teil abgebaut wurde. Mercaptobenzothiazol wurde zum Teil metabolisiert zu Hydroxybenzothiazol. Benzothiazol Sulfonsäure wurde zum Teil umgewandelt zu Hydroxybenzothiazol. OHBT konnte erst nach 5 Wochen zusammen mit anderen unbekannten Substanzen nachgewiesen werden. Bezüglich Thiocyanomethylthiobenzothiazol (TCMTB), ist OHBT neben MBT somit ein wichtiger Metabolit. Unter photolytischen Bedingungen sind Mercaptobenzothiazol und methylthiobenzothiazole nicht stabil. In Wasser und Abwasser konnten zwei stabile Photoprodukte, Benzothiazol (BT) and Hydroxybenzothiazol (OHBT), nachgewiesen werden.

5

Contents

page ACKNOWLEDGEMENTS 2 ABSTRACT 3 TABLE OF CONTENTS 5 LIST OF FIGURES 7 LIST OF TABLES 9 1.INTRODUCTION 10 2. THEORETICAL BACKGROUND 13 2.1 General considerations 13 2. 1.1 Provenance toxicity and use and of benzothiazoles 13 2.1.1.1 Provenance 13 2.1.1.2 Toxicity 18 2.1.1.3 Use 20 2.2 Benzothiazoles in aquatic system 22 2.2.1 Quantification and identification 22 2.2.1.1 Gas chromatography, GC 22 2.2.1.2 High performance liquid-chromatography, HPLC 23 2.2.1.3 Others 24 2.2.2 Extraction of benzothiazoles from water 26 2.2.2.1 Liquid-liquid extraction, LLC 26 2.2.2.2 Solid-phase extraction, SPE 27 2.2.3 Microbial transformations of benzothiazoles 27 2.2.3.1 Aerobic transformations 28 2.2.3.2 Anaerobic transformations 31 2.2.3.3 Sequencing Batch Reactor, SBR 34 2.2.4 Photochemical processes in waste water 34 2.2.4.1 Some basic principles of photochemistry 36 2.2.4.2 Photolysis rates and sunlight quantum yield 36 3. EXPERIMENTAL 39 3.1 Chemicals and materials 39 3.2 Analytical Apparatus and procedures 39 3.2.1 HPLC analysis 39 2.2.2 UV-absorbance 41 2.2.3 Dissolved Organic Carbon, DOC 41 2.2.4 Photolysis 41 3.3 Identification 42 3.4 Reagents and Sample preparations 43 3.4.1 HPLC-samples 43 3.4.2 Aerobic degradation of benzothiazoles 43

6

3.4.2.1 Substrate with biomass and glucose 43 3.4.2.2 Substrate without biomass (abiotic)/Substrate without glucose (not activated) 44 3.4.3 Anaerobic degradation 44 3.4.4 Sequencing Batch Reactor, SBR 45 3.4.5 Photolytic transformations 47 3.4.6 Solid Phase Extraction, SPE 48 4. RESULTS AND DISCUSSION 50 4.1 Development of a method for the benzothiazole extraction, SPE 50 4.1.1 Search for a suitable phase for the benzothiazole extraction 50 4.1.2 Sequential solid-phase extraction, SPE 54 4.1.3 Other eluents 55 4.1.4 pH influence on the extraction 57 4.1.4 Application on wastewaters 58 4.2 Aerobic degradation of some benzothiazoles in water samples 62 4.2.1 Substrate with biomass (sludge) and in the presence of glucose 62 4.2.2 Substrate without biomass (abiotic) 69 4.2.3 Substrate in the absence of glucose (not active) 69 4.2.4 Biodegradation of mercaptobenzothiazole (MBT) in the Sequencing-Batch Reactor (SBR) 73 4.2.4.1 Benzothiazoles in influent 1 75 4.2.4.2 Benzothiazoles in effluent 1 (reactor 1) 77 4.2.4.3 Benzothiazoles in effluent 2 (reactor 2) 79 4.3 Anaerobic degradation of some benzothiazoles 82 4.3.1 Mixture of benzothiazole substances 82 4.3.2 Individual substances 83 4.4 Photolytic degradation of some benzothiazoles 90 4.4.1 MTBT photolysis in pure water 90 4.4.2 MBT photolysis in pure water 92 4.4.3 Application in wastewater 102 4.4.3.1 MTBT photolysis in wastewater (biotic) 102 4.4.3.2 MBT photolysis in wastewater (biotic) 103 4.4.3.3 OHBT photolysis in wastewater 106 5. CONCLUSIONS 108 7. REFERENCES 112

7

LIST OF FIGURES Figure 2.1: General structure of benzothiazoles 13 Figure 2.2: Tautomeric forms of MBT and OHBT 16 Figure 2.3: Metabolic pathways proposed by Reemtsma et al. (1995) 30 Figure 2.4: Partial pathway of TCMTB photolysis and desinfection (Brownlee et al., 1992) 36 Figure 3.1: UV maximum absorption of MBT at pH 6.8 and 8.2 40 Figure 3.2: Photolysis instrument 42 Figure 3.3: Schematic diagram of the SBR units 46 Figure 4.1: UV- chromatograms of a standard mixture of benzothiazoles in aqueous solutio 50 Figure 4.2: SPE on different phases by pH 7 53 Figure 4.3 a: : Chromatogram example of extraction (M1)on RP18 /MeOH 53 Figure 4.3: Chromatogram examples of extraction on different sorbents 54 Figure 4.4: Sequential extraction of benzothiazoles 55 Figure 4.5: Chromatogram examples (M1) with different eluents on RP18 56 Figure 4.6: Recovery of sequential extraction and sequential elution of benzothiazole compounds at different pH- values 57 Figure 4.7: UV-Chromatograms example of spiked effluent samples on EN-phase eluted with methanol 59 Figure 4.8: UV-Chromatogram example of spiked influent sample on RP 18-cartridge

eluted with methanol 59 Figure 4.9: Recovery in influent and effluent samples 61 Figure 4.10: BT aerobic degradation test 63 Figure 4.11: OHBT aerobic degradation test 63 Figure 4.12: Degradation of BT and OHBT under aerobic conditions forming

NH4+ and SO4 65

Figure: 4.13: MTBT aerobic degradation test 66 Figure: 4.14 a: MBT aerobic degradation test (µg/L) 66 Figure 4.14 b: MBT aerobic degradation test (µmol/L) 67 Figure 4.15: Aniline aerobic degradation test 69 Figure 4.16: BT aerobic degradation test in absence of glucose and under abiotic

conditions 70 Figure 4.17: OHBT aerobic degradation test in absence of glucose and under abiotic conditions 70 Figure 4.18: MBT aerobic degradation test in absence of glucose and under abiotic conditions 71 Figure 4.19: MBT aerobic degradation test in presence of Antraquinone-sulphonate and

biomass 72 Figure 4.20: MBT aerobic degradation test with antraquinone sulfonate (abiotic essay) 72 Figure 4. 21: pH-value of SBR experiment 74 Figure 4. 22: DOC-SBR experiment 74 Figure 4.23: UV-absorption of SBR 75 Figure 4.24: HPLC-results of influent 1 76 Figure 4.25: Concentrations in effluent 1 (reactor 1 77 Figure 4.26: Concentration in effluent 2 (reactor 2) 79 Figure 4.27: SBR transformations in Reactor 1 and 2 80 Figure 4.28: Anaerobic behaviour of benzothiazole mixture 82

8

Figure 4.29: Anaerobic biotransformation of OHBT 83 Figure 4.30: Anaerobic biotransformation of MTBT 84 Figure 4.31: Anaerobic biotransformation of MBT 84 Figure 4.32: Anaerobic biotransformation of BTSA 85 Figure 4.33: Anaerobic biotransformation of TCMTB 86 Figure 4.34: Observed anaerobic transformations of benzothiazoles 87 Figure 4.35: UV-transformations of MTBT under photolysis 91 Figure 4.36: HPLC-results of MTBT-photolysis 91 Figure 4.37: UV-absorption of MTBT at 223 and 280 during photolysis 92 Figure 4.38: UV-transformations of MBT under photolysis 93 Figure 4.39: HPLC-results of MBT-photolysis 94 Figure 4.40: MBT photolysis in pure water 95 Figure 4.41: HPLC-results of MBT photolysis 96 Figure 4.42: MBT solution in the absence of photolysis 97 Figure 4.43: Comparison of MBT behavior at 320 nm under and in absence of photolytic conditions 97 Figure 4.44: UV-Comparison of MBT at 320 nm under and in absence of photolysis 98 Figure 4.45: UV-irradiation of MBT in ethanol (Párkányi and Abdelhamid, 1985) 99 Figure 4.46: UV-irradiation of MBT in ethanol (Abdou et al., 1987) 100 Figure 4.47: MTBT behaviour in wastewater under and in the presence of photolysis 103 Figure 4.48: MBT behaviour in wastewater without photolysis 104 Figure 4.49: MBT photolysis in wastewater 104 Figure 4.50: MBT in wastewater under photolysis 105 Figure 4.51: Comparison of MBT in wastewater with and without photolysis 106 Figure 4.52: OHBT behaviour in wastewater under photolytic conditions 106 Figure 5.1: Observed aerobic transformations 109 Figure 5.2: Observed SBR transformations 109 Figure 5.3: Observed anaerobic transformations 110 Figure 5.4: Observed photolytic transformations 111

9

LIST OF TABLES Table 2.1: Concentration of some benzothiazoles in different waters 15

Table 2.2: Structural formulas, chemical names, abbreviations and some properties of studied

benzothiazoles 16

Table 2.3: Toxicity of studied benzothiazoles 19

Table 2.4: Summary of a few benzothiazoles identification and quantification

by some authors 25

Tale 2.5: Degradation results of benzothiazoles under aerobic conditions 31

Table 2.6: Aerobic and anaerobic degradation tests of some benzothiazoles in aquatic

environment 33

Table 3.1: Experimental UV- maximum absorption 41

Table 3.2: Growth medium 44

Table 3.3: Photolysis conditions of some benzothiazoles 47

Table 4.1: Recovery rates average (%) for the SPE-extraction (neutral pH) 52

Table 4.2: Recovery of extraction with different eluents 56

Table 4.3: Concentration of benzothiazole in effluent (10 µl injection) 60

Table 4.4: Concentration of benzothiazole in influent (10µl injection) 60

Table 4.5: Carbon balance prior and after MBT addition 75

Table 4.6: Influent 1 between 3 and 16 weeks 77

Table 4.7: Balance of benzothiazoles found in the effluent 1 78

Table 4.8: Transformation of benzothiazoles under aerobic treatment 87

Table 4.9: Quantum yield (Φ) of studied benzothiazoles in water and wastewater 102

10

1. Introduction Expanding commercial applications of chemicals in industrial production, agricultural

practices, and human health care generate large quantities of hazardous chemical waste that

often result in environmental pollution. Approximately 80 billion pounds of hazardous wastes

are produced annually in the United States alone (U.S. International Trade Commission, 1982;

Leisinger, 1983). Water pollution is most commonly associated with the discharge of effluents

from sewers or sewage treatment plants, drains and factories. The organic pollution of

wastewater is present in soluble, colloidal and particulate forms and is mostly converted by

microorganism into biomass, carbon dioxide, water, organic by-products and metabolic organic

matter.

There is growing concern in recent years regarding possible harmful effects to man and other

living organisms of polluting organic chemical released into the environment as a consequence

of man’s activities. This concern arises primarily from the fact that some organic compounds

have been shown to be carcinogenic to mammals (Chaudhry, 1994).

Benzothiazoles are one of the more significant classes of polluting chemicals and have become

of global concern as serious environmental pollution, which give rise to this concern. They are

increasingly finding application in the field of industrial processes and as consequence

potential pollutants of the aquatic environment in which they are susceptible to be microbially

and / or photochemically degraded. A wide variety of benzothiazoles are available and some

of them are well known in the rubber industry.

Because of their desirable physical and chemical properties, benzothiazoles gained widespread

use in a variety of application, such as rubber additives and anticorrosive agents. Additionally

it is a substituent of a wide range of chemicals, including drugs and herbicides. Owing to its

very varied applications, benzothiazoles are produced and used in large quantities in many

countries. The production of benzothiazoles in the United States for the rubber processing, the

major use, was > 30,000 tons in 1981 (U.S. International Trade Commission, 1982). The total

production of benzothiazole derivatives in the United States in 1981 was 16.3 x 104 Tons (U.S.

International Trade Commission, 1982).

Benzothiazole are an important class of industrial chemicals. Many benzothiazoles are utilized

as vulcanization accelerators in the manufacture of rubber, as pesticides, as corrosion inhibitors

in antifreeze, and as photosensitizers in photography (Brownlee et al., 1992).

11

Several benzothiazoles have been reported in the environment: in production plant waste water

(Jungclaus et al., 1976; Puig et al.,1996 ), in river (Brownlee et al., 1992; Meijers et al., 1976;

Jungclaus et al., 1978) and drinking water (Frensch et al., 1987; Coleman et al., 1980 ), street

runroff (Spies et al., 1987), tire plant waste water (Jungclaus et al., 1976; Coleman et al.,

1980), sediment (Spies et al., 1987), ground water (Barker et al., 1988), starry flounder liver

(Spies et al., 1987), in street dust (Rogge et al., 1993), atmospheric aerosols (Kim et al., 1990),

in milk and other foods (Taylor,1986) and in surface water (Dietrich et al., 1988).

N-,O-, and S-heterocycles (NOSHs) comprising benzothiazoles were also found in polar and

asphaltic fractions of crude petroleum (Warshawsky , 1992; Beiko et al.,1987; Catallo, 1990).

Toxic and ecotoxic effects of benzothiazoles have been described by Hinderer et al. (1983),

Makoto Ema et al. (1989); Hendriks et al. (1994), De Wever, (1994), and Reemtsma et al.

(1995).

Although there is an extensive literature on the production and use of benzothiazoles

(Encyclopedia of chemical technology, 1982; Ullmann, Encyclopedia of industial Chemistry,

1993), there are considerably fewer reports on their environmental occurrence and fewer still

on their environmental fate. Because bacteria are important agents in transforming most

organic compounds, exploring their specific roles in biotransformation of benzothiazole

compounds holds great interest (De Wever et al., 1997; De Wever et al., 1998; Reemtsma et

al., 1995). Only a few benzothiazole-degrading organisms are reported in the literature and the

information on benzothiazole degradation pathways is scarce. However, the fact, that these

compounds can be observed at concentrations in the marine environment similar to other well

known contaminants justify further studies.

The objectives of this work were to study the biodegradation of benzothiazoles, in order to

understand the nature and pathways of BT degradation and to investigate their

photodegradability in water and wastewater. Additionally, a development of their extraction

was necessary.

Therefore, and within this framework, the main objective of the present work is an attempt to

clarify some questions concerning:

12

• the biodegradability of benzothiazoles in water and wastewater using aerobic

conditions with mixed cultures, with the microbial breakdown of the benzothiazole

skeleton.

• The anaerobic degradation of benzothiazoles. In the complete absence of oxygen, few

studies have shown that the natural environments contain microorganisms able to

degrade benzothiazoles (Reemstma et al., 1995).

• The determination and identification of metabolites in case of aerobic degradation as

well as in the anaerobic case.

For this purpose the biodegradation of such compounds was investigated in fed-batch systems.

Another aim of this study is to investigate the photodegradability of some benzothiazoles and

to discover if all studied benzothiazoles were photodegradable.

Owing to the deficiency of reliable analysis procedures, relatively few reports have appeared in

the literature on the specific extraction of benzothiazole compounds from aqueous media. In

this thesis, sensitive methods for extraction and analysis of benzothiazoles were developed and

selected. In order to determine these compounds in effluents and wastewater, a method was

required for the parallel extraction and analysis of several benzothiazole compounds from

wastewater. We developed a suitable method for these purpose.

Initially a method of extraction and chromatographic analysis of nine benzothiaole compounds

(BTSA, BT, OHBT, ABT, MBT, MTBT, MeBT, TCMTB, MBTS) from water and wastewater

was developed. An analytical quantitative high-pressure liquid chromatography (HPLC)

method using UV detection was developed for simultaneous determination of all studied

benzothiazoles to achieve this aim.

13

2. Theoretical background

2.1 General considerations

2.1.1 Provenance, toxicity and use of benzothiazoles

Benzothiazoles form a part of xenobiotic, heterocyclic, molecular structures comprising a

benzene ring fused with a thiazole ring. Their general structure is shown in Figure 2.1.

S

NR

Figure 2.1: General structure of benzothiazoles

2.1.1.1 Provenance

Benzothiazole enter the environment from a variety of sources such as the leaching of rubber

products, but particularly by routes associated with the manufacture and use of

mercaptobenzothiazole (MBT) and MBT-based rubber additives (Jungclaus et al., 1976), fine

particles of automobile tires, and antifreeze. Therefore, several benzothiazoles have been

reported in aquatic environments: Benzothiazole (BT) in river and drinking water (Carey et al.,

1983; Rennie, 1988). Additionally, this compound has been identified as surface water

pollutant (Braunstein et al., 1989) and in ground waters. MBT in plant waste water (Coleman

et al., 1980; Jungclaus et al., 1976), and methylthiobenzothiazole (MTBT) in drinking and

river water (Carey et al., 1983; Coleman et al., 1980). The presence of benzothiazoles in

atmospheric deposition (Welch and Watts, 1990) was also reported. This observed atmospheric

benzothiazoles can be explained by traffic- or wind-induced resuspension of street dust with

subsequent atmospheric transport . Rogge et al. (1993) found a BT concentration of 4.4 mg/L

in the road dust from Pasadena. It interesting to note that benzothiazole was also identified in

fish samples (Runge and Steinhart, 1990).

BT and MTBT were frequently found in both effluents of wastewater treatment plants (Paxeus,

1996) and surface water. MBT was removed from effluents in industrial effluent treatment

plants of rubber additives manufactures (De Wever and Verachtert, 1994). The same authors

reported the production of benzothiazole-2-sulfonic acid (BTSA) and, particularly at high

MBT concentrations, of a dimer of MBT (dithiobisbenzothiazole (MBTS)), which is insoluble

and accumulated in the activated sludge.

14

Benzothiazoles were also found in tannery wastewater: BT, MTBT and MBT (Reemtsma et al.,

1995) with a dominance of MBT at a concentration of 3.3 - 6.9 µmol/L. These compounds

have been detected in the waters by other authors (Fiehn et al. 1994; Liska et al. 1992; Rivera

et al., 1987). 655 µg/L MBT, 10.5µg/L BT, and 39 µg/L MTBT were detected in tannery

wastewater samples by Fiehn et al. (1994).

In Japan, not only 0.5 µg/L of MBTS was identified in water and a concentration ranging from

0.05 to 0.17 mg/kg in sediment tests, which were not near the industrial places, but also

concentrations from 12 to 82 ng/g of 2-(-4morpholino)benzothiazole (24MoBT) in the street

dust samples and 373 ng/g in the soil collected from the edge of a highway in San Francisco

have been detected (Kumata et al., 1996). Additionally, and in the same country (Japan),

MBTS was proved in different medications, which were packed with rubber seal. This MBTS

migrated into the aqueous phase. MBT was found also in human liver.

High quantities of MBT (1.1 mg/L), BT (20 mg/L) and MBTS (1.8 mg/L) were detected in

wastewater resulting from the production of rubber additives (Puig et al., 1996). Previous

studies found that benzothiazoles BT, OHBT, and 24MoBT were the major organic

compounds that leach from crumb rubber material (CRM) and asphalt containing 1-3 % rubber

(Wright et al., 1996). The compounds BT and OHBT are breakdown products in some of the

vulcanisation accelerators and antioxidants added to rubber during manufacturing (Santodonato

et al., 1976).

Some of benzothiazoles, particularly 2-mercaptobenzothiazole (2-MBT), are also used as

fungicides and in machine coolants. Thus the diffusion of small molecules from rubber into

surrounding fluids has also been shown to transmit mercaptobenzothiazole compounds into

milk and other foods (Taylor,1986).

Methylbenzothiazole (MeBT) was identified in water of some estuaries. Concentrations of this

compound were determined (Spies et al., 1987; Theobald et al., 1995). BT and MTBT were

detected in marine water during an ecotoxicological experiment (Bester et al., 1995) at

Helgoland in the central German Bight. A concentration of 55 ng/L of MTBT in the Elbe river

were found by Bester et al. (1995), while in water from the North sea (Germany) the

concentration of MTBT ranged from 0.04 to 1.37 ng/L. The values for MTBT varied from 0.4

to 1.4 ng/L in 1993 by the same authors (Bester et al., 1993a and 1993b). As well as the

15

presence of benzothiazoles in two European rivers (Hendriks et al., 1994; Franke et al., 1995)

was mentioned but the concentration about these compounds was lacking.

Some benzothiazoles such as MBT (0.4 mg from 537 mg extract), MeBT (1.8 mg from 537 mg

extract) and OHBT (0.5 mg from 537 mg extract) have been isolated from fermentation culture

extracts of Micrococcus sp, which is a marine bacterium obtained from tissues of the sponge

Tadania ignis (Stierle and Cardellina, 1991).

Benzothiazole was previously found in all of the extracts of waters collected from four creeks

which flowed through urban areas into Hamilton Harbour (Scott et al. 1995). This indicated

that benzothiazole in Hamilton Harbour was not necessarily directly from an industrial activity,

but a by-product of urban life entering the receiving water from a number of tributaries. Scott

at al. (1996) detected a concentration ranging from 1 to 10 ng/L of benzothiazoles in water

samples from the great lakes drainage basin (Ontario, Canada). 124.3 mg benzothiazole/Kg in

tire waste and 4.4 mg benzothiazole/Kg in road dust were found. By the TCMTB production in

Leverkusen, ca. 250 kg MBT/a were registered in wastewater.

Benzothiazoles enter antifreeze from different ways. For example, they can be leached from

the rubber hoses in the cooling systems of automobiles. In antifreeze BT and OHBT were 29.6

and 20.1 mg/L respectively (Reddy and Quinn, 1997).

Table 2.1 shows the existence of some benzothiazoles in different waters. Structural formulas,

chemical names, abbreviations, water solubility and octanol / water partition coefficients of

tested benzothiazoles are given in Table 2.2.

Table 2.1: Concentration of some benzothiazoles in different waters

country Sample place benzothiazole concentration Germany atmosphere BT qualitative detectable USA Sewage from rubber industry BT 20 and 60 µg/L USA River water BT 2,3 and 23µg/L USA drinking water MTBT qualitative detectable Antarctic sea water MTBT 17 ng/L Antarctic sea water BT 6-37 ng/L USA Sewage from rubber industry MBT 0.54-0.71 mg/L USA Sewage from MBTS production MBT 0.5-1 mg/L USA Sewage from MBTS production MBTS 35 µg/L

16

Table 2.2: Structural formulas, chemical names, abbreviations and some properties of studied

benzothiazoles

R = formula Chemical names Abrevia tion

Mole cular weight

Water Solubili ty (mg/L) 25°C

Octanl/water (log10 Kow)

SO3H OH NH2 SH H CH3 SCH3 SCH2SCN S-S-BT

C7H5N O3S2 C7H5NOS C7H6 N2S C7H5 N S2 C7H5NS C8H7NS C8H7 N S2 C9H6 N2S3 C14H8 N2S4

Benzothiazol Sulfonic acid Hydroxybenzothiazole Aminobenzothiazole Mercaptobenzothiazole Benzothiazole Methylbenzothiazole Methylthiobenzothiazole Thiocyanomethylthiobenzothiazole Dithiobisbenzothiazole

BTSA OHBT ABT MBT BT MeBT MTBT TCMTB MBTS

215 151 150 167 135 149 181 238 332

nf 2354 c 310.3 c 120 *,e

4300 a 366.3 c 125 *,5 125 *, e 10 a

-0.99 2.35 b 2.00 d 2.86 f 2.17 b 2.72 d 3.22 g 3.12 a

4.66 d

* 24°C, nf = not found, a CHEM INSPECT TEST INST (1992), b Hansch, C et al. (1995), c Meylan, WM et al.

( 1996), d Meylan, WM &Howard, PH (1995), e Brownlee, BG et al. (1992), f TSCATS, g Platford, RF (1983).

Both MBT and OHBT can exist in two tautomeric forms (Fig. 2.2).

S

NSH

S

N HS

��

2-Mercaptobenzothiazole 2(3H)-benzothiazolethion

MBT (thiol form) (thion form)

��

S

N H

S

NOH O

2-Hydroxybenzothiazole 2(3H)-Benzothiazolone

OHBT (thiol form) OBT (keto form)

Figure 2.2: Tautomeric forms of MBT and OHBT

17

All benzothiazole used are solids at room temperature with the exception of benzothiazole

(BT), which is liquid. Table 2.2 shows that BT possesses a high solubility (4300 mg/L), this is

probably due to its high polarity and the fact that it is a liquid at room temperature. BT is also

considered as volatile with a vapour pressure 0.0143 mm Hg (25°C).

OHBT has a good solubility of 2354 mg/L in water. MBT, MTBT and TCMTB are moderately

soluble with respective solubilities of 120 mg/L, 125 mg/L and 125 mg/L. These values of

solubility and their respective vapour pressure (0.000464, 0.00026, and 3.12 10-7 mm Hg)

show that these benzothiazoles are be considered as not volatile from aqueous solutions (or

volatile with difficulty).

MeBT and ABT are quite soluble with 366.3 mg/L and 310.3mg/L respectively. MBTS is

poorly soluble with a solubility of 10 mg/L and not volatile (2.54 10-10 mm Hg) from aqueous

solution. Brownlee et al. (1992) studied the solubility of some of benzothiazoles at different pH

values and temperatures in water and found that the solubility of MBT was enhanced and

showed a pH dependence indicating that the ionised form is more soluble, but that of TCMTB

was not. TCMTB hydrolyses quantitatively to MBT, the speed of hydrolysis is highly

depending on the pH-value, the temperature and the matrix.

Based on octanol/water partition coefficients given in table 2.2, we notice that the studied

benzothiazoles cover a wide range of polarity from pKow –0.99 (BTSA) to pkow 4.66 (MBTS).

BTSA is the least hydrophobic with the lowest pKow. ABT, OHBT, BT, and MeBT are

intermediate and have relatively low pKow coefficients ranging from 2 to 2.72. Based on their

log Kow values these benzothiazoles can be bioaccumulated (Metcalfe et al., 1988). Because the

structures of BT and OHBT do not have hydrolysable or reactive moieties like in the case of an

ester bond, abiotic chemical transformations should not occur.

MTBT and TCMTB are considered as hydrophobic with pKow of 3.12 and 3.22, respectively,

which is consistent with their structures (see Table 2.2), but TCMTB is the most hydrophobic.

MBTS with a pKow = 4.66 means that this compound has a high potential of bioaccumulation.

MBT exists in three forms in neutral solution (pH 6-8): the unionised thione and thiol

tautomers (with a thione predominance) (Fig. 2.2) and the ionised thiolate form. Therefore, this

large range of solubility led to the difficulty in their identification and separation. De Wever et

18

al. (1994) showed that, according to its octanol/water partition coefficient, MBT might interact

with membrane-bound system and might solubilize in bacterial membranes. Consequently,

MBT affected bacterial membranes and interfered with the aerobic respiratory chain. The

partition of benzothiazoles in octanol indicated that especially MBT and OHBT, at the

physiological pH values, are quite hydrophobic and thus may distribute in hydrophobic

bacterial cell membranes (De Vos et al. 1993). This was confirmed through experiments

showing that bacterial cell respiration was inhibited using lactate or succinate as substrates.

This is very likely due to their existence in tautomeric keto forms (Halassa and Smith, 1973).

Because some benzothiazoles are water soluble, depending of their Kow, it is unlikely that they

will sorb to particles, or be bioaccumulated.

Based on these specification, we can say that the analysis of these compounds on reversed-

phase chromatography may be difficult and requires quite different conditions. We note that

BTSA is poorly retained on C18-phase, while the others are from moderately to well retained

on C18-phase. The MBTS is the latest to be eluted on this phase.

2.1.1.2 Toxicity

Some authors have shown that MBT is mainly responsible for toxic effects in MBT production

activated sludge (De Wever et al., 1997; De Vos et al., 1993). MBT has been also shown to

induce tumors (Gold et al., 1993), to be toxic (at 600 mmol L-1) to aquatic organisms

(Yoshioka and Ose, 1993), and may also hamper waste treatment (De Wever et al., 1994).

MBT and MBTS have been reported as one of the most frequent allergens causing shoe

dermatitis (Storrs, 1986; Jung et al., 1988). Hinderer et al. (1983) proved that MBTS induced

genetic damage to mammalian cells. MBT interfered with the nitrification processes and

exhibited biocidal effects (Czekowski and Roomoor 1981; Usov et al. 1987; Shuto et al.

1989).

Toxicity of MBT towards bacterial growth and respiration has been studied by Santodonato et

al. (1976), Usov et al. (1987) and by De Wever et al. (1994). MBT has an approximate fatal

concentration of 2 mg /L in 48 h (Sitting, 1981) and a LD50 (in rat by oral injection) of 100

mg/kg (Sax, 1984). The toxicity of MBT towards several bacteria was confirmed and was

situated at around 100 mg/L.

19

The aquatic toxicity of TCMTB to some freshwater organisms has been reported (Nishiuchi,

1976). Nikl and Farrel (1993) studied the effect of TCMBT on the reduced swimming

performance and gill structures in juvenile salmonids exposed to TCMBT. This later has been

also reported to be mutagenic when tested by the salmonella typhimurium system as test

microorganisms (Jeans and Li, 1978). The oral LD50 of TCMTB for rats was 1.59 g/Kg.

Additionally, TCMBT has some effects on the microfungal populations and some biological

properties of soils (Dacre, 1981).

Toxic and ecotoxic effects of derivatives of benzothiazoles have been described by Reemtsma

et al. (1995) concerning bacteria (Vibro fischeri) and inhibition of nitrification. Luminescence

inhibition (EC50) of V. fisheri was measured by Reemtsma et al. (1995) and was 6 µg/L for

TCMTB and 3 mg/L benzothiazole. The same group found also an inhibition of nitrification in

sediment columns at concentrations of 0.1-0.3 mg/L for benzothiazole and MBT as well as

another derivatives, while Hendriks et al. (1994) tested MTBT and BT for acute toxicity on

Daphnia magna and found EC50 values of 10 and 50 mg /L, respectively. Benzothiazole has a

mouse LD50 of 900 mg /Kg (oral) (Sax, 1984).

Table 2.3: Toxicity of studied benzothiazoles (EC50 and LD50)

benzothiazole CAS Registry No.

Acrute oral LD50 (mg/Kg)

EC50 (mg/L) LC50 ( mg/L)

MBT [149-30-4] 3,000 /100 0.12 (20h) V. fischeri 19 (48h) D. magna MBTS [120-78-5] >5,000 82 (48h) D. magna 66 (96h) Salmo

gairdneri BT [95-16-9] 900 4.4 (20h) V.fischeri 110 (48h) Oryzias

latipes OHBT [934-34-9] - - - ABT [136-95-8] - - - MeBT [120-75-2] - - - MTBT [615-22-5] - - - TCMTB [21564-17-0] 1,590 - -

20

2.1.1.3 Use

Benzothiazoles represents an important class of vulcanization accelerators. The production,

subsequent oxidation, and hydrolysis reactions of MBT result in a salt-rich waste-water

containing MBT, benzothiazole (BT), 2-hydroxybenzothiazole (OHBT) and benzothiazole-2-

sulfonate (BTSA) as the most important pollutants. Benzothiazoles in estuarine sediments are

potential indicators of sediment pollution by street runoff (Spies et al. 1987).

Benzothiazole (BT)

In rubber production, catalysts are used for the cross-linking of unsaturated polymers made

from monoenes. Among the accelerators used, there are several benzothiazoles. Primary

products are 2-mercaptobenzothiazole (MBT) (Ullmann´s Encyclopedia, 1993), 2-

hydroxybenzothiazole (OHBT), and, of course, benzothiazole (BT) itself.

The compound benzothiazole (BT) is also produced as a by-product in the industrial products

and processes and finds its way into a range of industrial effluents and is thus commonly found

in the aquatic environment (Brownlee et al. 1981). It is also produced as a result of wear on

vehicle tyres (Santodonato et al. 1976; Anonym. 1985). However, the presence of BT in river

water has been proposed as an indicator of the presence of road run-off in rivers (Spies et al.

1987). In California BT and MTBT were used as tracer for street runoff water (Spies et al.,

1987). BT is also mentioned as a fungicide (Liska et al., 1992).

Mercaptobenzothiazole (MBT)

MBT is the most important member of the benzothiazole group of heterocyclic aromatic

compounds. In fact, its discovery in ca. 1920 led to the major use in the production of rubber

additive chemicals but predominately, as vulcanization accelerator in rubber industry (Engels

et al., 1993; Days et al., 1993; Santodonato et al. 1976; Anon. 1985; Encyclopedia of chemical

technology, 1982). MBT is also applied for various purposes, such as bio-corrosion inhibitor

(Hahn, 1993; Reichert et al., 1993; Mylari et al., 1991; Wotersdorf et al., 1989) in industrial

cooling and in the galvanic industry, and coating agent of metallic surfaces (Rennie, 1988).

Some authors (Santodonato et al., 1976) talked about MBT use for its fungicidal properties, or

its antimicrobial effects (Foltinova et al., 1978; Czechowsky and Rossmoore, 1981; Ambro and

Otepka, 1982; De Wever et al., 1997). It is also used as an external chemotherapeutic and

antifungal drug in medical application (Bujdaková et al., 1993). The metal (mainly Zinc) salt of

21

mercaptobenzothiazole is used mainly in latex compounds (Ulltmann´s Encyclopedia of

industrial chemistry, 1993). Its metal chelating properties may be directed towards metal ions

in enzymes and MBT is an effective inhibitor of lactate dehydrogenase (Czechowsky and

Rossmoore, 1981).

2,2-dithiobisbenzothiazol (MBTS)

MBTS, also called Dibenzthiazyl disulphid formed by oxidation of 2-mercaptobenzothiazole is

employed extensively and exclusively as vulcanisation accelerator at concentration of 0.05-2 %

(weight) in the rubber industry. From the total amount of benzothiazole accelerators, ca. 53%

are used in tire production and 47% in the manufacture of technical rubber (seals, tubes,

footwear, tyres, belts, etc…).

The chemicals used as a rubber accelerator have a high reactive nature and are used to obtain

shorter cure times for the reaction between sulphur and rubber during vulcanisation. With

MBT, vulcanisation starts rapidly, with MBTS it is somewhat delayed.

Thiocyanomethyltiobenzothiazole (TCMTB)

The compound thiocyanomethylthiobenzothiazole (TCMBT) has been shown by Daniels and

Swan (1987) to possess useful antifungal activity. TCMBT showed considerable promise and

was a replacement for the chlorinated phenols on surface-treated lumber (Dacre, 1984; Cserjesi

and Johnson, 1982). In the United States, TCMTB has been shown to be an effective treatment

against stain on Southern pine lumber (Eslyn and Cassens, 1983).

TCMTB is an active ingredient in several formulations of pesticides. Furthermore, it is gaining

in popularity as antimicrobial agents in the paper and leather industries (Parbery and Taylor,

1989). The effectiveness of TCMTB as a fungicide for leather has been confirmed by various

workers as well as in ship and boat antifouling paints in many countries (Russell et., 1985;

Russell and Pinchuck, 1985; Galloway, 1973).

The oxidised derivatives of TCMTB are also used as a seed dressing against the diseases of

cereals, mainze, rice, cotton, sugar, cotton, sugar beet and legumes. TCMBT has also been

shown to bind more strongly to soils with higher organic content (Hanssen et al., 1991).

22

2-Aminobenzothiazole (ABT)

Aminobenzothiazole (ABT) is used in the manufacturing of some disperse azo dyes, which are

prepared by diazotization of aminobenzothiazoles followed by coupling with N,N-

dialkylaniline and for the preparation of pesticides (Shuttleworth et al., 1990; Towns, 1999;

Pesticide Chemistry, 1998). ABT was the first type to be exploited and has gone on to become

a source of technically important red dye-stuffs for polyesters and continues to attract industrial

interest (Hu et al., 1987). However, the synthesis of benzothiazole-based azo-disperse dyes has

increasingly become available, the volume of material pertaining to the properties and

application of colorants of this kind has grown markedly.

2.2 Benzothiazoles in aquatic system

2.2.1 Quantification and identification

Relatively few reports have been appeared in the literature on the chromatography of

benzothiazole. Therefore, only a restricted number of studies of the chromatographic behaviour

of benzothiazoles in water is available.

2.2.1.1 Gas chromatography, GC

The determination of benzothiazoles, in particular 2-(4-Morpholinyl)benzothiazole, in various

types of environmental samples (e.g., river sediments, water, river water) by gas

chromatography equipped with a flame photometric detector (FPD) has been studied by

Kumata et al (1996). This study led to a limit of detection of 0.2 ng. Gas chromatography has

been also used by Wallnöfer (1976), Christopher and Quinn (1997) and Saxton (1987), Jung et

al. (1988), and Dhindsa et al. (1990). Chowdhury et al. (1979) used gas chromatography for the

separation and identification of alkyl-and bisbenzothiazoles on SE-30 and Carbowax 20M

columns. In this case both type of benzothiazoles showed satisfactory separation on SE-30

column but the Carbowax column proved to be unsuitable for these compounds.

Gas chromatography was also performed by Fiehn et al. (1994) with a flame ionization

detector for the analysis of BT and MTBT from industrial wastewater. The authors found that

these compounds were well detectable with a detection limit ranging from 1 to 2 ng, which

were compared with those of HPLC results by the same authors and proved to be slightly lower

23

than the latter one. Nevertheless, HPLC analysis was favourable in wastewater, since TCMTB

and MBTS were not detectable by means of GC.

In our case, we tried to analyse all cited benzothiazoles including TCMBT, therefore gas

chromatography cannot be used to analyse this compound (TCMBT) due to its instability at

high temperatures (Daniels and Swan, 1987). Thus, HPLC has been used to analyse the

benzothiazole derivatives.

Another method for the determination of benzothiazoles with (GC / MS) has been used by

Headley (1987), Dietrich et al. (1988) and Barker et al. (1988).

2.2.1.2 High performance liquid-chromatography, HPLC

High performance liquid chromatography (HPLC) has made a large impact on the analysis of

compounds since its beginning in the late 1950s and throughout its continuous development to

the present. Consequently, it has great potential for use in water analyses where many

dissolved substances are either ionic or polar and therefore exhibit low volatility. For detailed

theoretical background material on the HPLC, a number of books are available on HPLC

(Snyder, 1979; Simpson, 1982; Krstulovic and Brown, 1982).

Analysis of some benzothiazole compounds, which are more polar or less thermally stable such

as TCMTB by reversed-phase liquid chromatography has been reported (Daniels and Swan,

1987; Fowler et al., 1987; Elfarra and Hwang, 1990; Brownlee et al., 1992; Stolkova and

al.,1999). Reversed-phase LC on C18 phases has also been previously employed for the

quantitative analysis of TCMTB (Parbery and Taylor, 1989; Bugby et al., 1990; Kennedy,

1986). MBT, BT, TCMTB, and MTBT were separated on C18 reversed-phase by Brownlee et

al. (1992) in less than 10 min with acetonitrile/water (50/50) as mobile phase. For analysis of

MBTS, an eluent of methanol-water (90/10) was used.

The determination of MBTS and TCMTB in industrial and municipal wastewaters was

performed by Warner et al. (1985) using a HPLC-method. Benzothiazoles were determined by

HPLC-UV (De Vos et al., 1993; De Wever et al. 1998) on a lichrosorb RP18 column with a

mobile phase of 35:65 acetonitrile-water and a commercially ion pair reagent (Pic A) was

added.

24

An analytical quantitative high pressure liquid chromatography (HPLC) method was also

developed by Hansson and Agrup (1993) for simultaneous determination of four benzothiazole

derivatives (MBT, MBTS, MMBT (morpholinyl mercaptobenzothiazole) and CBS (N-

cyclohexyl-2-benzothiazyl sulfenamide)) in a mix-batch testing standard.

Fiehn et al. (1994) used HPLC for the analysis of TCMTB, MBT and MTBT from industrial

wastewater (RP18 column and UV-detection) and suggested that the method was not limited to

the analysis of industrial wastewater.

All these HPLC-analysis coupled to UV-detection proved to be suitable to detect the

benzothiazoles mentioned but the sensitivity of UV detection was limited.

2.2.1.3 Others

The structural identification of some benzothiazoles, mainly BTSA and MBT by NMR; MS-

and IR spectroscopy was performed by Fiehn et al. (1997) in addition to HPLC/UV and

GC/MS for their analysis. The GC-MS was also developed by Wilson et al. (1991) to identify

some benzothiazoles in urine extracts.

Louter et al. (1994) used the on-line solid-phase extraction-Gas chromatography-mass selective

detection (SPE-GS-MSD) to determine benzothiazoles in river water. This method was shown

to be a rapid technique for the trace-level determination and provisional identification of

organic pollutants in surface water. The recovery of benzothiazole with SPE-GS-MSD was 80

± 12 % at a level of 1 µg/l, while Reemstma (2000) developed a method for the analysis of 2-

substituted benzothiazoles from wastewaters by reversed phase liquid chromatography /

electrospray ionization tandem mass spectrometry (LC/ESI-MS/MS). It was found later that

the method was well applicable for the direct analysis of wastewaters, and the detection limits

ranged from 10 ng/l for ABT to 200 ng/l for OHBT and 2.5 µg/l for BT.

Ultraviolet Spectrophotometry (Jones and Woodcook, 1975), and also capillary electrophoresis

(Ching-Erh Linet al.,1996) have been also used for the determination of benzothiazoles.

Some alkyl, aryl- and bisbenzothiazoles have been examined by Chowdhury et al. (1979)

using thin layer chromatography (TLC) in order to develop analytical procedures and to

determine the influence of substituents. No simple relationship was found for the RM values of

the alkyl derivatives but the RM values of the bis-compounds showed linear behaviour,

25

indicative of mixed partition-adsorption and pure partition separation, respectively.

Mercaptobenzothiazole (MBT) was analysed using TLC in 1988 by Larsen et al., also by

Fukuoka and Tanaka in 1987. Thin layer chromatography has been also used by Jung et al.

(1988) and Dhindsa et al. (1990) and High performance thin layer chromatography (HPTLC)

was used by Stumpf et al. (1986).

Table 2.4 summarises some works directed to the identification and quantification of some

benzothiazoles.

Table 2.4: Summary of a few benzothiazoles identification and quantification by some authors

Benzothiazoles Medium Method of identification and /or quantification

Authors Year Observations

MBT and derivatives

Petrotatum and in buffer solutions

HPLC Hansson and Agrup

1993 Good separation

BT River water Solid-phase extraction-GC-MS

Louter et al. 1994 Well suitable for the trace-level of surface water determination

Alkyl, aryl and bisbenzothiazoles

TLC/GC Chowdhury et al.

1979 Satisfactory results

TCMTB Treated wet blue

HPLC Fowler et al. 1987 Reasonably consistent results

MBT Water and Tannery wastewater

HPLD/GC and capllary electrophoresis

Fiehn et al. 1998 BTSA was the main transformation product

2-substituted benzothiazoles

wastewater HPLC/Electrospray ionization tandem mass spectroscopy (HPLC/ESI-MS/MS)

Reemtsma 2000 Suitable method for the direct analysis of wastewaters

BT Water samples from lakes

GC-AED Scott et al. 1996 The AED provided element-specific detection

26

2.2.2 Extraction of benzothiazoles from water

Analysis of organic pollutants in wastewater requires procedures that are both flexible and

capable of detecting specific pollutants. However, the behaviour and performance of the

different extraction materials, even those used successfully for the extraction of common water

samples, may change on account of the nature of the sample matrix, and the pollutants of

interest. Liquid-liquid extraction (LLE) is useful but, as in other fields of environmental

analysis, solid phase extraction (SPE), with all its possible modifications can be used too.

SPE is based on the principle that the components of interest, mainly organic molecules are

retained on a special solvent placed in a disposable extraction mini-column. Goals of the SPE

methods are: less contamination as compared to LLE and easy handling in the field. One of the

main advantages of the use of SPE techniques is the application for storage and transportation

of the analytes.

The most common retention mecanisms in SPE are based on Van der Waals forces (,non polar,

interactions), hydrogen bonding and dipole-dipole forces (,polar, interactions) and cation-anion

interactions (,ionic, interactions). Each sorbent offers an unique mix of these properties which

can be applied to a wide variety of extraction problems. The properties of each sorbent were

described in some literatures. For more details concerning the SPE refer to some

bibliographies.

2.2.2.1 Liquid-liquid extraction, LLC

The use of LLE, was investigated by Warner et al.(1985) who employed dichloromethane, but

limited to TCMTB and MBTS, from spiked distilled water. However, recovery from waste

water was poor (or low). While recoveries of MBTS from distilled water ranged from 75-90 %

of 10-100 µg/L and were greater than 85% in the case of TCMTB at concentrations from 10 to

1000 µg/L, the recoveries in wastewater were low at 23 and 60 %, respectively.

Liquid-liquid extraction for the determination of Methylthiobenzothiazole (MTBT) has been

studied by M. Dietrich, 1988. Fiehn et al. (1994) developed a method for MBT, MTBT and

TCMTB extraction, from industrial wastewater using ethyl acetate and toluene. From distilled

water at different pH-values, they found that MBT was well extractable with ethyl acetate or

toluene and under weakly alkaline condition (8.5) with a recovery >100%, while BT, TCMTB

27

and MTBT appeared to be best extractable from neutral pH (6.5) between 94 and 98 %, or

acidic pH (2) between 89-98 %. Recovery from different wastewaters spiked with the

benzothiazoles cited was shown to exceed 90 % for all compounds and concluded that SPE

extraction used was not suitable for all benzothiazole compounds. LLE with

dichlororomethane, formerly employed in other reports, proved to be insufficient for the more

polar substances MBT and BT.

2.2.2.2 Solid-phase extraction, SPE

In the present work a method based on SPE for the extraction of benzothiazoles was developed

and applied to the water and wastewater samples. However, only few reports on the specific

extraction of benzothiazole compounds from aqueous media are found.

TCMTB was extracted from leather samples with dichloromethane using liquid phase

extraction (Fowler et al., 1987) and from preservative-treated timber by Kennedy (1986).

SPE by on-line trace enrichment for the purpose of surface water pesticide residue analysis was

reported (Liska et al., 1992). The authors employed a PRLP-S polymer. BT and MTBT have

also been determined by SPE on C18 phases within toxicity fractionation procedures applied to

surface waters (Burkhard et al., 1991; Amato et al., 1992).

Fiehn et al. (1994) employed solid-phase extraction (SPE) on different polarities of solid

phases. These authors found that MBT was well recovered on C18 phases with

methanol/methanol-dichloromethane (20:80) and MTBT was most efficiently extracted on the

cyano phase (around 85%), while TCMBT was extractable by C8 and C18 phases with

acetonitrile/acetone/toluene. BT was poorly extracted in all cases. Thus the recovery of BT and

MTBT stayed insufficient and SPE appeared to be of limited value for the parallel extraction of

MBT, BT, MTBT and TCMTB. Thus, SPE has been found not suitable for all benzothiazoles.

2. 2. 3 Microbial transformations of benzothiazoles

Although the ability of microorganisms to degrade various benzothiazoles has been studied, the

microbial degradation of sulfur-containing heterocyclic compounds has not been as extensively

studied as those compounds containing only nitrogen as heteroatom. For this reason, only

limited research into the microbial transformations of benzothiazoles is available.

28

2.2.3.1 Aerobic transformations

Some variety of microorganisms of different genera have been isolated that are capable of

degrading benzothiazoles under aerobic conditions. De Wever et al. (1998) were the first to

isolate a pure culture, identified as a Rhodococcus erythropolis capable of BTSA, OHBT and

BT degradation but not MBT, which was found to inhibit the biodegradation of the cited

benzothiazoles. Chudoba et al. (1977) and Reprika et al. (1983) have studied the behaviour of

benzothiazoles in activated sludge system. In this case, BT and OHBT were considered to be

biodegradable, whereas MBT and BTSA were not. With adapted biomass, Mainprize et al.

(1977) observed a degradation of hydroxybenzothiazol (OHBT) but any degradation of

mercaptobenzothiazole (MBT), benzothiazole (BT) and benzothiazole 2-sulphonic acid

(BTSA) was observed under the same conditions. Whereas, in 1976 the same authors

(Mainprize et al.) considered that BTSA was capable of being completely degraded by micro-

organisms. However, if BTSA-mineral- salts were heavily inoculated with a mixture of the

organisms grow on BTSA-mineral-salts agar plate, growth occurred. BTSA disapeared of the

medium, the single nitrogen and both the sulphur atoms being converted to ammonia and

sulphate respectively.

Under aerobic conditions, OHBT was observed as an intermediate in BT breakdown. Gaja and

Knapp (1997) reported that pure bacterial cultures described as Rhodococcus strain PA, were

capable of growing on benzothiazole (BT) itself as a sole carbon, nitrogen and energy source.

OHBT and BT are considered to be biodegradable (De Vos et al., 1993).

To simulate more realistic environmental conditions, Reddy and Quinn (1997) spiked BT and

OHBT at 15 ppb into a water collected from the highway settling-pond. After 8 days, they

found that both BT and OHBT at reasonable concentration can be microbially degraded in

water in a relatively short time. Additionally, the metabolic fate of benzothiazoles has been

investigated in 1991 by K. Wilson et al..

Reemstma et al. (1995) and De Wever et al. (1997) have shown that BT and MBT were

removed in activated sludge systems. 10% of MBT were methylated by the bacteria to MTBT

within 28d, whereas 87% remained unchanged (Reemstma et al., 1995). MBT was removed

from solution in the presence of both live and dead activated sludge by Gaja and Knapp (1998),

in this case, no evidence has been obtained for the truly degradative removal of MBT and non

29

for ability of microbes to utilise it as a sole carbon source. Additionally they suggested that the

removal of MBT observed from solution in the presence of several sludge was non-enzymatic,

but may have been due to adsorption on the activated sludge or to a chemical reaction

associated with the sludge.

Aerobic degradation of mercaptobenzothiazole (MBT) by sulphur bacteria has been also

studied in aquatic environment (Karelová and Tomasoviková, 1988) who estimated the

degradation rate in the presence and absence of glucose. Concentrations of 5 µg/L and 10 µg/L

were tested. While the degradation was slow by addition on glucose, the degradation rate of

benzothiazole in the absence of glucose was more rapid: after 4 weeks of incubation, 44 % and

74 % of MBT were degraded in samples containing 5 and 10 µg /L MBT, respectively.

However, the situation concerning biodegradation / transformation of mercaptobenzothiazole

(MBT) by Chudoba et al. (1977) and De Wever and Verachtert (1994) is much less clear. Also

the oxidative destruction of MBT was investigated (Fiehn et al., 1998) where MBT has been

ozonated in pure water and in spiked tannery wastewater. In this case BTSA proved to be the

main transformation product.

While the degradation of 2-aminobenzothiazole (ABT) to give high yields of ammonia and

sulphate has been demonstrated by both pure and mixed bacterial cultures by Gaja and Knapp,

Chudoba et al. in 1977 on the other side did not observe any ABT degradation in what appear

to be in that time the only reported study on its biodegradability.

In aerobic degradation, thiocyanomethylthiobenzothiazole (TCMBT) was transformed to

MTBT and BT along unknown pathways (Perraud et al., 1995; Reemtsma et al., 1995) as well

as to mercaptobenzothiazole, all of which proved to be comparatively stable within a study in

aquatic and terrestrial ecosystems (Reemtsma et al., 1995). A provisional metabolic scheme

(Fig. 2.3) was proposed by Reemtsma et al. (1995) from aerobic batch tests with the main

pathway leading from TCMTB via MBT to MTBT, which was not further degradable.

30

S

NS CH2 SCN

anaerobic/aerobic

S

NSH

S

N

S

NSR

S

N

S

NS SCH3

O

CH3

degradation

aerobicaerobic

MTBT

MBTaerobic

MSiBT

aerobic

anaerobic

BT

-?-

-?-

TCMTB

no degradation

-?-

Figure 2.3: Metabolic pathways proposed by Reemtsma et al. (1995)

TCMTB was transformed to MBT, which was than methylated to stable MTBT (Fig. 2.3). BT

was aerobically degraded along unknown pathways. MBT and BT were substantially

eliminated by aerobic treatment, while MTBT was formed.

TCMTB was aerobically degraded for about 95 % within 9-12 days. MBT was the main

product, while about 5% of the TCMTB appeared to be methylated from MBT to MTBT,

which was persistent. These authors mentioned also that only 45 % of employed TCMTB was

recovered as MBT and MTBT in all their experiments. BT was completely degraded after 8

days. The degradation results for each compound are presented in the following table (Table

2.5).

31

Tale 2.5: Degradation results of benzothiazoles under aerobic conditions

Authors BTSA BT OHBT MBT ABT TCMTB

De Wever et al. (1998) + + + -

Chudoba (1977), Reprika (1983) - + + - -

Mainprize (1977) - - + -

De Vos et al. (1993) + +

Mainprize et al. (1976) +

Gaja and Knapp (1997, 1998) + + +

Reddy and Quinn (1997) + +

Reemstma et al. (1995), De Wever (1992) + +

Karelová and Tomasoviková (1988) +

Perraud et al. (1995), Reemstma et al. (1995) +

+ degradation, - no degradation

Based on all these studies, we can observe that the process of benzothiazole biodegradation is

largely unknown. Although some of them are degraded in activated sludge system, most of

data on the pathways concerning the biodegradability of benzothiazoles obtained in the

laboratory are inconclusive.

In spite of this wide occurrence of the thiazole ring system, relatively little is known of its

susceptibility to metabolism. From the few examples reported in the literature it appears that

compounds containing an un-fused thiazole may undergo attack either at one of the

heteroatoms (Ofen et al., 1985), or at one of the ring carbon atoms leading to a ring-scission

product (Grupe and Spiteller, 1982). In contrast, substituted benzothiazoles undergo either the

hetero-ring hydroxylation (Thomas, 1984), or nucleophilic substitution at position 2 if a

suitable leaving group is present.

2.2.3.2 Anaerobic transformations

Understanding and application of anaerobic treatment has made significant gains in the past 30

years. The complexity of anaerobic treatment and less experience with the process compared to

its aerobic counterpart are reasons for variations among mathematics models. To make an

environmental risk assessment for a chemical, it may be important to determine the chemical´s

susceptibility to anaerobic biodegradation. Furthermore, Owin et al. (1979) provided the first

description of such a test method, drawing on previous gas measurement (Nottingham, 1969)

32

and incubation bottles (Miller and Wolin, 1974) methods. They based their method on

measurement of excess gas volume (CH4 + CO2) produced after addition of a test chemical to

an anaerobic seed incubated in sealed bottles.

The gas volume was measured from the displacement of the piston in a glass syringe whose

needle had been inserted into the bottle. Subsequently, Gledhill (1979) improved the method

with goal of defining a simple protocol that could be established by the American Society for

Testing Material (ASTM) as a standard method. He introduced the use of a pressure transducer

to measure the gas pressure, recommended 50 mg of carbon per litre to the test chemical

concentration, and 10% anaerobic sludge as described by Healy and Young (1979).

Daniel et al. (1984) determined the anaerobic degradation by the net increase in gas pressure

in bottles with test chemicals over the pressure in non amended sludge bottles. Gas production

was measured by gas chromatography and by pressure sensor. Net gas production was

calculated by substracting gas produced in non amended bottles from that produced in test

bottles.

Degradation was expressed as percentage of theoretical gas production based on stoichiometry

of mineralisation to CH4 + CO2 and correcting for gas solubility. This gas will be divided

between CO2 and CH4 based on the stoichiometry of the reaction, which can be calculated by

the Buswell equation (Tarvin and Buswell (1934)) (eq. 2.2). Methane production was

quantified by injection of head space gas from serum bottles into a gas chromatography

equipped with a flame ionisation detector.

CxHyOz + (x – y/4 – z/2)H2O → (x/2 –y/8 +z/4)CO2 + (x/2 +y/8 –z/4)CH4 eq.2.2

The anaerobic degradation of benzothiazoles derivatives has not been as extensively studied as

the aerobic degradation. Anaerobic degradation of MBT by sulphur bacteria in aquatic

environment has been studied by Karelová and Tomasovicová (1988). At both MBT-

concentration tested (5 and 10 µg/ml), its degradation was slow by the addition of glucose. In

the absence of glucose, however, the degradation rate of MBT was more rapid. After 5 weeks

of incubation, 44 and 74 % of MBT were degraded in sample containing 10 and 5 µg/ml MBT,

respectively.

De Wever et al. (1998) found under anaerobic conditions, that the concentrations of BT, MBT,

and OHBT remained constant through the test period, whereas BTSA was transformed into

33

OHBT in nearly equimolar amounts. A summary of research on the aerobic and anaerobic

degradation of some benzothiazoles is listed in Table 2.6.

Table 2.6: Aerobic and anaerobic test degradation of some benzothiazoles in aquatic

environment

Authors Year Medium tests Results

Karelová and Tomasovicová

1988 Aquatic environment

Anaerobic degradation

44 % and 74 % of respectively 5 µg/l and 10 µg/l of MBT were degraded after 5 weeks

Gaja and Knapp

1998 Industrial sludge

Microbial degradation

No evidence of MBT degradation

De Wever et al. 1997 Activated sludge

Aerobic degradation

MBT was not bactericidal but rather was bacteriostatic


Toxicity/ degradation

Activated sludge may be intoxicated by MBT (100 mg/l)


degradation MBT→degradation MBT sludge showed a tendency to accumulate BTSO3 in a first degradation product of MBT

Gaja and Knapp

1997 Polluted river water


ABT/ BT/ OHBT→ ammonia + sulphate ABT and BT degradation in polluted river water unsuccessful

Knapp and Callely

1976 Industrial effluents


BTSA→ NH4+ + SO4-- completely biodegradable

Karelová and Tomasovicová

1988 Aquatic environment

Anaerobic degradation

MBT partial degraded

De Wever et al. 1998 wastewater Anaerobic/aerobic Biodegradation

BTSA→ OBT ( anaerobic) BT→ OHBT and dihydrogeno-BT intermediate (aerobic)

De Vos et al. 1993 Activated sludge

Toxicity and biodegradation

MBT → MBTS /BTSO3

Reemtsma et al. 1995 Industrial water

Aerobic/anaerobic Microbial transformation

TCMTB→MBT→MTBT (aerobic) BT: degradation

Brownlee et al. 1992 Natural water

photodegradation MBT→BT + OHBT + unidentified product TCMTB→ MBT→ BT MTBT stable

34

2.2.3.3 Sequencing Bach Reactor, SBR

Despite the few available information on the biochemistry related to benzothiazoles

biodegradation (particularly in batch-type systems employing mixed and /or pure cultures),

information on the feasibility of sequencing batch reactor systems for the treatment of these

compounds is limited or does not exist.

Sequencing batch reactor (SBR) are an attractive alternative to conventional biological

wastewater treatment system, mainly because of their simplicity and flexibility of operation

and cost effectiveness for small-scale treatment facilities (Irvine and Ketchum, 1989).

The SBR has been used since the early 1970´s for the biological treatment of wastewater

(Irvine and Davis, 1971). Owing to the nature of the process, and advances in process control

technology, there is renewed interest in the SBR for the treatment of inhibitory industrial

wastewaters.

Therefore, the purpose of this study was an attempt to investigate the biodegradability of

mercaptobenzothiazole (MBT) in continuously operated SBR units using an activated sludge.

2. 2. 4 Photochemical processes in water

Photochemical processes influence the fate of organic substances in water. Organic compounds

can photolyse directly or indirectly. In direct photolysis, the component of interest absorbs

light and reacts. Indirect photolysis occurs when a chemical species absorbs light and transfers

the energy to the compounds of interest, which then reacts. In fact, a compound may not

undergo photochemical transformation at all in distilled water, but may react at a significant

rate in water containing other constituents. The reason of this behaviour (Schwarzenbach,

1993) are either an energy transfer from another excited species (indirect photolysis) or

chemical reactions of the (non excited) compounds with very reactive species (e.g., hydroxyl

radicals, peroxy-radicals, singlet oxygen) that are formed in the presence of light. The most

important light absorbers that may induce indirect photolytic transformations of organic

pollutants are the chromophores present in dissolved organic material.

Photochemistry of aromatic thiols and disulfides has been extensively studied (Oswald et al.,

1963; Burkey and Griller, 1985; Newcomb and Parks, 1986; Alam et al., 1995; Alam et al.,

1996; Alarm et al., 1998; Aveline et al., 1995; Aveline et al., 1996). The sulphur compounds

are known to be an effective source of sulphur-centred free radicals and ion radicals, which are

35

important intermediates in photochemical reactions and also in kinetic investigations (Ito and

Matsuda, 1979). Flash photolysis methods using laser and Xe lamps as excitation sources have

been employed as attractive techniques to detect reactive intermediates such as free radicals

and ion sulfur compounds (Aveline et al., 1996; Alam et al., 1996).

Literature data concerning the photochemical behaviour of benzothiazole compounds are very

scarce. Since both BT and OHBT absorb sunlight only very weakly at wavelengths > 290 nm

(Fife et al. , 1975), direct photolysis of these compounds will be a minor fate (Schwarzenbach,

1993).

TCMTB is broken down rapidly by direct photolysis in sunlight, producing MBT as the major

product (about 50% yield) and traces of BT (Brownlee and al., 1992). The same authors

studied the sunlight photolysis of MBT in water systems and found BT (30-45%) and OHBT

(4-5%) as metabolites. MTBT was stable toward direct photolysis by sunlight. Laboratory

studies have been reported on the photolysis of MBT in organic solvents using UV lamps and

Pyrex vessels (Párkányi et al., 1985 and Abdou et al., 1987).

MBT was postulated as the first intermediate in the irradiation of MBTS in 96% ethanol in the

presence of oxygen (Párkányi et al., 1985). This was proposed to undergo oxidation to MBTS,

which, upon cleavage and hydrolysis, produced benzothiazole sulphate isolated in 90% yield.

Irradiation of MBT in MeCN in the presence of oxygen gave MBTS, BT, OHBT, and

unchanged MBT (Abdou et al., 1987).

Further experiments are required to distinguish between many possibilities presented by the

chemistry and the various oxidation states of sulfur (Kice, 1980).

On the basis of previous studies, a partial pathway was formulated for TCMTB and MBT in

aquatic systems (Fig. 2.4). Photolysis of TCMTB leads to MBT, which can either photolyse to

BT and OHBT or undergo biomethylation to MTBT. Under water or sewage disinfection with

chlorine, it is possible that BT and MTBT can undergo oxidation by hypochlorite to OHBT

(Lin and Carlson, 1984) and sulfoxide (MsiBT) and sulfone (MsoBT), respectively.

36

S

NSCH2SCN

S

NSH

��

hydrolysis

sunlight

��S

N HS

N

S(O)CH3

S

N

S

N+ OHSCH3

��

BT OHBT

MSiBT

TCMTB MBT

sunlight

��

S(O)2CH3

S

N+

MSoBThypochloriteoxidation

��MTBT biomethylation

hypochloriteoxidation

Figure 2.4: Partial pathway of TCMTB photolysis and its disinfection by chlore

(Brownlee et al., 1992)

2.2.4.1 Some basic principles of photochemistry

If a molecule (A) is exposed to ultraviolet (UV), electrons may get promoted from bonding or

non-bonding orbitals to so-called anti-bonding orbitals (e.i., σ*- or π*-orbitals). The molecule

is then said to be in an excited state (A hγ A*). When a given pollutant absorbs light, it

undergoes transformations as a consequence of that light absorption. This process is commonly

referred to as direct photolysis. Beer-Lambert law relates the light intensity Iλ emerging from

the solution to the incident light intensity I0λ:

Iλ = I0λ . 10 - [(α (λ) + ε (λ) C] l = I0λ .10 - A eq.2. 3

and Absorbance A = log (I 0λ / Iλ) = [(α (λ) + ε (λ) C] l

I aλ = I 0λ - Iλ = (1-10 – A) I0λ

hc 1,196 . 105

E = 6,023 . 10 23 = . 2 J /einstein (1 einstein = 1mol/photon)

λ λ

37

Where,

C: concentration of the compound of interest (mol/L)

l: path length of the light in the solution (cm)

α (λ): attenuation coefficient of the medium(L mol –1 cm -1)

ε (λ): molar extinction coefficient of the compound at wavelength λ(L mol –1 cm -1)

λ exc: λ excitation.

(1- 10 - A): Absorption grade (%, 0-1).

A: absorbance Iaλ = I0λ - Iλ : light absorbed 2.2.4.2 Photolysis rates and sunlight quantum yield

For a given environment, a quantum yield, Φ is defined as the fraction of the excited molecules

of a given compound that react by physical or chemical pathways:

The light quantum yield, Φ is defined as:

total number(i.e., moles) of molecules transformed

Φ (λ) = eq.2.4

total number (i.e., moles) of photons (of wavelength λ) absorbed by the system

Environmental photolysis rates and half-lives can be calculated if the sunlight quantum yields

are known (Zepp and Cline, 1977).

Light quantum yields for benzothiazole compounds were calculated using an adaptation of the

method of Leifer (Leifer, 1988).

Iaλ = I0λ (Ac/V) Fsλ Fcλ eq. 2.5

A/V = cm-1

Where

Iaλ : light absorbed by the chemical at wavelength λ related to volume

I0λs : incident radiation at wavelength λ

38

Ac/V : area/ volume ratio for the reaction cell, cm2/ cm3

Fsλ : fraction of light absorbed by the system

Fsλ = [1-10 - (α (λ) + ε (λ) C) l ]

Fcλ: fraction of light absorbed by the chemical

Fcλ = ε (λ) [C] / (α (λ) + ε (λ) [C]) or since ε (λ) [C] << α (λ), = ε (λ) C / α (λ).

The determination of the light quantity (Kuhn et al., 1989), I0λ was calculated using the

following equation:

I0λ = I0λS . Ac eq. 2.6

Where, I0λS : light amount density (J/s.cm2)

Ac: surface of the cell (cm2)

Unfortunately, there are no simple rules to predict reaction quantum yields from chemical

structures, therefore, Φ (λ) values have to be determined experimentally. In principle, reaction

quantum yields may exceed unity in cases in which the absorption of a photon by a given

compound causes a chain reaction to occur that consumes additional compound molecules.

Such cases are, however, very unlikely to happen with organic pollutants in natural water,

mainly because of the rather low pollutant concentration, and because of the presence of other

water constituents that may inhibit chain reactions (Schwarzenbach, 1993). Consequently, we

always assume a maximum reaction quantum yields of one. We note also the wavelength

dependence of Φ. Although reaction quantum yields differ considerably between different

wavelengths, there are in many cases approximately wavelength-independent for reaction of

organic pollutants in aqueous solutions For more detail, refer to some reviews

(Schwarzenbach, 1993; Hoigné et al., 1985).

39

3. Experimental 3. 1 Chemicals and Materials

TCMTB was supplied by Dr. Eberle chemical plant GmbH (Tübingen, Germany). BT and

OHBT were purchased from Sigma-Aldrich (Steinheim, Germany). MeBT, MBTS and MTBT

were obtained from Ferak (Berlin, Germany). ABT was received from Fluka (Buchs,

Switzerland). MBT was supplied from Merck (Darmstadt, Germany). BTSA was kindly

provided by Heleen de Wever (Leuven, Belgium).

HPLC eluents were prepared from acetonitrile (HPLC grade, Merck) and water that had been

purified by an model ELGA maxima HPLC system (Elga, High Wycombe Bucks, England).

Natrium-dihydrogen-phosphate (NaH2PO4) was HPLC-quality. Other chemicals were of the

highest analytical grade available.

Glass fibre filters were obtained from Schleicher and Schuell (Dassel, Germany) and 0.45 µm

cellulose acetate filters from Sartorius (Göttingen, Germany).

Solid-phase extraction tubes used in this study were from: Lichrolut, Merck (200 mg cartridges

RP 18; 200 mg EN), Supelco (200 mg Envi-18; 0.25 g Envi-Chrom P), and Isolute (200 mg

Env +).

The wastewater used throughout the study was collected from a municipal waste water

treatment plant (Ruhleben, Berlin).

3.2 Analytical Apparatus and procedures

3.2.1 HPLC analysis

Separation, detection and quantification was performed on an Eurospher 100 C18 reversed

phase column (5µm*3mm ID ( Merck); with an L-6200A gradient pump , an AS 2000A

autosampler and a T- 6300 column thermostat (all from Merck-Hitachi, Tokyo (Japan)) and

equipped with a SPD 10 A UV-visible detector (Shimadzu, Japan). Data storage and

processing were performed with ChromStar 3.11 software (Bruker, Bremen). 10 µl samples

were injected.

Separations were performed with an acetonitrile-water gradient. Eluent A was an acetonitrile-

water (5/95) and 3 mmol L-1 NaH2PO4 were added. Acetonitrile-water, (70/30) and 3 mmol L-1

NaH2PO4 were used as eluent B.

40

Separation of the benzothiazole compounds was achieved under the following conditions:

Elution started with 2% B at a flow rate of 0.6 ml min-1 after that was linearly shifted to 100

%B at 15 min with the same flow rate (0.6 ml/min), and stayed constant (100% B) until 18

minutes, then decreased gradually to the beginning of elution (2%B) in 10 min. Column

temperature was maintained at 40°C.

N

S

H

S ��

N

SS H

6.8pH : p 8.2:H

Figure 3.1: UV maximum absorption at pH 6.8 and 8.2

UV detection was performed by time-programmed wavelength changes according to the

absorption maxima (Table 3.1) of the MBT compound, which was detected at 320 nm after 11

min. All the other benzothiazole compounds were detected at 220 nm. In the case of MBT the

retention time and the absorbance maximum were dependent on the pH value of eluent. For

example the third maximum absorption of a MBT solution (in water) was 323 nm until pH 6.8,

presumably due to ionisation above pH 7. But at higher pH (pH >6.8) the absorbance

maximum was shifted from 323 to 313 nm. Also the intensity of the second absorbance

maximum (237 nm) increased with high pH (see Fig.3.1).

41

Table 3.1: Experimental UV- maximum absorption (X.max.= maximum absorption)

substance 1.max. (nm) lg (I0/I) 2.max. (nm) lg (I0/I) 3.max. (nm) lg (I0/I)

BTSA 267 0.945 221 0.765

BT 218 1.786 251 0.755 284 0.224

MBT 237 0.129 228 0.106 323 0.184

MBTS 201 1.305 272 0.301 225 0.75

MeBT 250 0.992 282 0.250 293 0.225

TCMTB 222 1.687 280 0.956 289 0.907

MTBT 223 1.276 278 0.798 288 0.692

OHBT 217 0.561 242 0.327 282 0.158

3.2.2 UV-Absorption

The UV-absorbance was determined with a Lambda 2 spectrophotometer (Perkin Elmer) using

1-cm quartz cells. All samples were filtered through 0.45µm filter before measurement.

3.2.3 Dissolved Organic Carbon, DOC

Dissolved organic carbon (DOC) of waste water samples was determined with an Astro

LiquiTOC 2001-MB analyzer (Foss-Heraeus, Hanau, Germany) by UV-persulfate oxidation.

The organic carbon content was obtained from the difference between the values of total

carbon (TC) and inorganic carbon (IC).

3.2.4 Photolysis

For analytical purpose, we used the instrument shown below (Fig. 3.2) in devices equipped

with a XBO 450 W/1 Xenon lamp whitch was placed on a lamp housing (LAX 1450, Fa.

Müller), an PEM 500 (Müller) recording light meter (to measure the incident radiation), an

MHO 60 IR filter, an optical power meter model 1815-c (optical meter, Newport) with a 818-

UV detector with attenuator (for small light intensity). Glass filters were from type: UG 295

(Scott, Mainz). The solution was constantly stirred by means of a small magnetic stirrer. The

irradiation time was in range of 60 to 200 min.

42

Photolysis was carried out in aqueous solutions. Samples were irradiated in a 381QX 5000 (50

mm) quartz cylindrical reactor or a QS 1000 quartz cell (10 mm).

Filter (313 nm)

Light source IR filter quartz cell

(high pressure Xe-lamp) (water)

hγ

A A*

Figure 3.2: Photolysis instrument

An MHO 60 IR filter was placed to avoid overheating. The high pressure Xenon lamp

stimulated the radiation, this light was reflected in the mirror and focused, thus an

homogeneous and regular bunch light was created (see Fig.3.2 ).

3.3 Identification

Peak identification was based on the retention times of the corresponding standards. External

calibration was used to quantify the benzothiazoles, and MeBT was used as internal standard

in the case of their extraction.

Ratios of the chromatographic peak area for each benzothiazole to that of the I. S. (Internal

Standard) were calculated. Standard curves were constructed by plotting the peak area ratio of

the analyte to the internal standard against the concentration of the analyte.

Accuracy and repeatability of the method were assessed analyzing the blank calculated sample

concentrations and regression parameters from a series of calibration curves of benzothiazoles

in water that were prepared and analyzed on separate days.

43

3.4 Reagents and Sample preparations

3.4.1 HPLC-samples

A stock solution containing a concentration in the range of 2-3 mg/ml of each analyte was

prepared in methanol, except for TCMTB and MBTS, which are poorly soluble in methanol.

These compounds were dissolved in 1-2 ml dichloromethane (CH2Cl2) and then diluted with

methanol.

All standards were freshly diluted in water from the stock solutions which were prepared with

methanol. Mixed standard solutions (calibration standards) were prepared by mixing a stock

solution with adequate water to achieve the desired concentrations of standard solutions

(100µL of the respective stock solutions and diluted to a volume of 1 ml). For the extracted

sample, an internal standard was used with the same concentration as those of the another

analytes.

3.4.2 Aerobic degradation of benzothiazoles

3.4.2.1 Substrate with biomass and glucose

The aerobic degradation tests were carried out according to DIN 29888 (1993). The tests of

BT, OHBT, MBT, and MTBT in water systems were carried out with glucose/glutaminic acid

(750/825 mg/L) as an additional carbon source and nutrients (solutions a, b, c, and d). The

growth medium, which provided the organisms with the required nutrients is shown in Table

3.2. The initial test concentrations were in the range of 2.5 mg/L and 7 mg/L. For 1 L test

medium, 10 ml/L of solution a, 1 ml of each solutions b, c, and d were added. The mixtures

were put and stirred in 1L glass bottles (200U/min). The temperature was maintained at ca.

20°C in a climatic chamber. The bottles were sealed with parafilm to minimise evaporation yet

permit aeration. The duration of test was of 30 days. Samples were taken out every two days,

filtered (0.45 µm) to remove suspended matter and analysed with the developed HPLC-

method.

A blank test (without benzothiazoles but with biomass, glucose, and nutrients) and a reference

test with aniline (for the control of the activity of the biomass) were performed in parallel.

All pipette tips (Pasteur pipettes) and glass bottles were autoclaved at 120 °C for 20 min

before being used. As inoculum, an activated sludge from a municipal waste water treatment

44

plant (Ruhleben, Berlin) was used. Before being used, aeration was adjusted to keep all the

sludge in suspension, then the sludge were washed (and then centrifuged) several times with

tap water to remove any adsorbed organic matters. The volume of suspension sludge added

was kept at around 1 ml/L in all experiments.

Table 3.2: Growth medium Solution (a) Solution (b) Solution (c) Solution (d)

KH2PO4 8.5 g/L

K2HPO4 21.75 g/L

NaH2PO4.2H2O 33.4 g/L

NH4Cl 0.5 g/L

Complete with

distilled Water bis 1L

pH ~ 7.4

MgSO4.7H2O 22.5/L

CaCl2 27.5 g/L

FeCl3.6H2O 0.25g/L

3.4.2.2 Substrate without biomass (abiotic)/ Substrate without glucose (not activated)

In order to ensure that the degradation of benzothiazoles under the conditions from above

(substrate with biomasse and glucose) resulted from biological transformations, the same tests

were carried out under sterile, abiotic (without inoculum addition) and not activated (without

glucose) conditions. The sterilisation was performed in an autoclave by 125 °C at a pressure of

1.5 bar for 20 min. The initial concentrations were 1.76 mg/L for OHBT; 1.64 mg/L for BT; 3

mg/L for MBT. An additional batch with pure biomass was used to determine metabolic

products of the biomass it self.

3.4. 3 Anaerobic degradation

Serum-bottles of 1000 ml capacity were filled with 100 ml RAMM I (reversed mineral

medium), 25 ml RAMM II, 10 ml RAMM III and bidestilled water (Shelton and Tiedje, 1984).

The medium was autoclaved for 10-15 min at 125°C under pressure (1.5 bar) to remove O2.

After cooling to approximately room temperature, 250 µl of ethanol as cosubstract were added

to the medium. As inoculum, 10 ml sludge was added, which originated from several years old

incubation with benzothiazoles, and was adjusted to 500 ml under a N2 gas to remove traces of

45

O2 and then protect the anaerobic bacteria from oxygen. Finally, the Serum bottles were sealed

with 1-cm-thick butyl rubber stoppers and capped with aluminium crimp seals (American

Scientific Products) and were incubated in the dark at 37 °C. Samples were incubated over a

period of 15 weeks.

Samples were taken every week by syringe under N2 after vigorous shaking, The initial

concentrations of benzothiazoles used in this study were in the range of 0.1 to 0.5 mg/L.

Anaerobic tests were followed by DOC and HPLC measurements.

Samples from the aerobic and anaerobic tests were filtered and stored frozen until analysed.

The anaerobic mineral medium (RAMM) consisted of (per liter):

RAAM 1 (phosphate buffer): 0.27 g of KH2PO4 and 0.35 g of K2HPO4 (adjusted to pH 7).

RAMM 2 (mineral salts): 0.53 g of NH4Cl, 75 mg of CaCl2 . 2H2O, 100 mg of MgCl2. 6H2O

and 20 mg of FeCl2.

RAMM 3 (trace metals): 0.5 mg of MnCl2 . 4H2O, 0.05 mg of H3BO3, 0.05 mg of ZnCl2, 0.03

mg of CuCl2, 0.01 mg of NaMo4 .2 H2O, 0.5 mg of CoCl2 . 6 H2O, 0.05 of NiCl2 .6H2O and

0.05 mg of Na2SeO3.

3.4.4 Sequencing Batch Reactor, SBR

Figure 3.3 shows a schematic drawing of the experimental SBR system used in this study,

which consisted of a 2L reactor (reactor 1) and a 1L reactor (reactor 2). The operating liquid

volume was 1.0 L and 500 ml respectively. The influent from a sewage treatment plant

(Ruhleben, Berlin) was filtered trough a glass fibre filter and then stored frozen. This served as

influent 1 to the reactor 1. Both reactors 1 and 2 were operated for 12 week prior to the spiking

of MBT. The addition of MBT began when stable operation of both reactors was obtained, and

detected by the UV-absorbance of the effluents. The initial concentration of MBT was 24.97

mg/L.

In the beginning, the reactor 1 contained 1 L of sewage and 10 ml of bacteria sludge, whereas

the reactor 2 contained only 500 ml tap water and the same amount of bacteria sludge (10 ml).

Reactor 1 was fed to obtain 24.97 mg/L of MBT (200 ml MBT was added as a solution of

MBT in influent 1). Reactor volumes were adjusted every week to 1 l for the reactor 1 and

500 ml for the reactor 2, by adding fresh influent from the same source (described earlier),

46

hence, the active volume of both reactors remained constant throughout the study. Aeration

was provide at the base of the reactor through two air-stones at opposite ends. A magnetic

stirrer-bar ensured complete mixing in the reactors. This was followed with a 30 min settling

quiescence phase, before that influent and effluent samples were taken at regular intervals

(every week). Samples were analysed for DOC, pH, UV-absorption and quantified and

identified with the developed HPLC-method described above. Then the cycle repeats (Fig.3.3).

All samples were filtered prior to analysis.

Figure 3.3: Schematic diagram of the SBR units

During the operation of the SBR system, the temperature was maintained at room temperature.

Collected raw wastewater samples were stored frozen until the SBR use. The end of the

experiment stopped after another 18 weeks.

47

3.4.5 Photolytic transformations

The benzothiazole were determined by irradiation of pure water and wastewater solutions with

initial concentration of 2.97 mg/L and 5.5 mg/L MBT, 11.2 mg/L OHBT, 3.4 mg/L and 8.2

mg/L MTBT, 3.5 mg/L BTSA. On the other hand water solution with the same concentration

of benzothiazole irradiated was prepared but not irradiated for comparison.

The disappearance of studied benzothiazoles and appearance of photolytic products was

followed by HPLC. The solutions were constantly stirred by means of a small magnetic stirrer.

Rates were expressed in terms of total incident light radiation in einsteins per square meter,

measured by the quantum sensor. The experimental conditions are summarised in Table 3.3.

Table 3.3: Photolysis conditions of some benzothiazoles

Test compound

Initial concentration (mg / L)

medium filter E= (mW/cm2)

MBT 2.97 pure water 313 nm 2.70

MBT 5.52 pure water WG 295, 3mm (314.8 nm)

0.62

MBT 5.52 wastewater WG 295, 3mm (314.8 nm)

0.62

MTBT 3.4 pure water 313 nm 2.70

MTBT 8.2 pure water WG 295, 3mm (314.8 nm)

0.72

MTBT 8.2 wastewater WG 295, 3mm (314.8 nm)

0.72

OHBT 11.2 pure water WG 295, 3mm (314.8 nm)

0.65

OHBT 11.2 wastewater WG 295, 3mm (314.8 nm)

0.65

Photolysis runs were begun after it was ascertained that the lamp amplitude and the electronics

had stabilised. The measurement procedure was as follows: the light power was measured with

the power meter, the shutter was then closed. The photolysis cell, containing a measured

volume (as seen above) of benzothiazole, solution was placed in position, and stirring was

started. Photolysis were then begun by opening the shutter. Timing was performed either

manually with a stopwatch or automatically with an electronic timer. Exposure times varied

48

from 60 to 600 s. The light power was immediately measured (to assure its stability) after the

photolysis cell was removed.

3.4.6 Solid-Phase extraction, SPE

The extraction capacity of each SPE material was tested.

-Phases on Silicagel basis: 200 mg cartridges RP 18 (Lichrolut) and 200 mg Envi-18 (Supelco),

-Polymer phases: 250 mg Envi-Chrom P (Supelco), 200 mg EN (Lichrolut, Merck) and 200 mg

Env + (Isolute).

All SPE-cartridges were conditioned by applying 2*2.5 ml methanol, 2*2.5 ml acetone, and

2*2.5 ml of water at 1 ml/min.

A volume of 100 ml of aqueous test solutions of five benzothiazoles (Mixture1):

Benzothiazolsulfonic acid (BTSA), Hydroxybenzothiazol (OHBT), Benzothiazol (BT),

Mercaptobenzothiazol (MBT) and Methylthiobenzothiazol (MTBT) in a range of concentration

2.3 - 2.6 µg/L were extracted on the conditioned solid phases cartridges. A second mixture

(mixture 2) of Aminobenzothiazol (ABT) and Thiocyanomethylthiobenzothiazol (TCMTB) in

the range of concentration as mixture 1 (2.5 - 2.7 µg/L ) was prepared. This volume of 100 ml

of pure water test was separately applied to each cartridge at 1mL/min. The analytes were then

sequentially eluted into the same vials with 2 ml methanol followed by 2 ml acetone.

MBTS was not include in the extraction because of its poor solubility in water and its quick

decomposition.

Extraction was carried out using a Visiprep solid phase extractions-Vacuum-Extraction

Instrument (Supelco). The solid phase eluents were concentrated by approximately 1mbar and

a chamber-temperature of 37 ºC in a Speedvac-instrument (A 160, Savant) to ca. 1ml. Finally,

a volume of 10µl of a solution containing 2.5 µg/L of Methylbenzothiazol (MeBT) as internal

standard was added to the sample. A general scheme of the analytical SPE procedure used in

the present work is shown later.

Prior to extraction, wastewater samples (influent and effluent from wastewater treatment plant)

were filtered in the same day through 0.45 µm membrane filters (Schleicher & Schuell, Dassel,

Germany, cellulose-acetate filter with glass fibre pre-filter).

49

Parallel to the samples, a blank test was performed with each cartridge (without adding

benzothiazoles) and eluted under the same conditions as the samples. Eluents were analyzed

regularly before and after extraction with the developed HPLC-method.

Due to the possible photolytic reactions of benzothiazoles, in particular TCMTB and MBT

(Brownlee et al., 1992) all samples, stock solutions and extracts were stored in the dark during

work-up and, generally, exposed to daylight as shortly as possible.

50

4. RESULTS AND DISCUSSION

4.1 Development of a method for solid phase extraction of benzothiazoles

4.1.1 Search for a suitable phase for the benzothiazole extraction

In order to determine the most efficient means of extracting benzothiazoles, we investigated

the use of SPE for the extraction of these compounds. BT and its derivatives BTSA, OHBT,

MBT, MTBT, ABT and TCMTB were extracted from wastewater.

For these experiments, in all cases, after extraction and reduction to 1 ml, the standard mixture

as well as all extracts were analysed with the developed HPLC-method described as above (as

mentioned in section 3.2.1.) by using acetonitrile /water as mobile phase and which led to a

good chromatographic separation. Figure 4.1 shows the UV-Chromatogram of the enriched

standard mixture in tap water.

O H B T

B

MBT

ABT

BTSA

TCMTBMTBT

B T

M e B T

A

Figure 4.1: UV- chromatograms of a standard mixture (8 benzothiazoles) in aqueous solution

(concentration 1.2-2.8 mg/L). A: mixture1 (M1) : 1 = BTSA, 2 = OHBT, 3 = MBT, 4 = BT,

5 = MeBT, 6 = MTBT. B: mixture 2 (M2): 1=ABT, 2=MeBT, 3=TCMBT. MeBT used as internal

standard

51

Recovery studies were performed in order to evaluate the performance of the developed

methodology for the extraction of benzothiazole analytes from wastewater. First, the

experiment consisted in applying the method of one step extraction. The recoveries of

enzothiazoles are represented in Table 4.1. ABT was not measured in this case.

hat the most

bstances could be not entirely enriched from aqueous solutions (see Table 4.1).

•

recovery rates of

BT could not be increased through elution with acetone (Table 4.1).

•

uent to 100 %. The recovery of BT

could not be increased through the second elution.

•

ith

respectively 58.4 and 78.7 %. BT, OHBT and MBT were not be eluted in this case.

•

on with acetone has increased considerably the recovery rate of

TCMTB (88.5 %).

b

Two solvents were used for elution of the SPE cartridge and it was found t

su

In the case of Envi 18 with methanol as eluent, ca. 98 % OHBT were recovered and

further ca. 2 % were recovered in the second extract giving a total of ca. 100 % (Table

4.1). The recovery rates of MBT, BT and MTBT are on the other hand smaller than 50

%. BTSA and TCMTB could not at all be detected in the first extract. With the second

eluent acetone, BTSA could be eluted in a great amount (50 %). The

Through elution on Env+ cartridge, ca. 80 % OHBT can be recovered. The recovery of

the other benzothiazoles were situated just between 20 and 45 % and for TCMTB by

0%. TCMTB could be but recovered in acetone-el

BTSA and OHBT recoveries rose to > 80 % when Envi-Chrom P-cartridges were used

and with methanol eluted (Table 4.1). Here, MTBT could not be detected. But by

employing the second eluent, acetone, TCMBT and MTBT were well extracted w

By the extraction on RP 18 material and methanol as elution milieu, OHBT, BT and

MTBT were recovered by more than 50 %. BTSA can not be detected by using

methanol. The eluti

52

•

orking. This might be also true for MTBT (23.2 %), the second

volatile compound. This problem would be avoided by on-line trace enrichment (Liska

and Brouwer, 1992).

Recov av age (%) fo PE-extraction (neutral p

BTSA was very good recovered (99.5 %) on EN-cartridge and methanol as elution

medium. OHBT gave a recovery of 68 %. Benzothiazole was poorly extracted by using

acetone. The recovery of BT compared to the other substances may be poor due to

volatilisation during w

Table 4.1: ery rates er r the S H).

Cartridge Eluent BTSA OHBT BT MTBT MBT TCMBT

Envi 18 MeOH 0 98.2 33.8 47.2 32.7 0

Envi 18 .2 acetone 57.5 2.02 0 9.9 29.7 30

Σ(Envi18) .8 57.5 100 33 57.1 62.4 30.2

Env+ MeOH 44.9 83.3 43.4 13.1 20.1 0

Env+ acetone 30.5 6.3 0 37.6 32.2 100

Σ(Envi+) 75.4 .6 .4 .3 89 43 50.7 52 100

Envi-ChromP MeOH 96.6 82.6 26.7 0 21.2 8.3

Envi-ChromP acetone 5.2 0 0 58.4 0 78.7

Σ(Envi chromP) .6 .7 101.8 82 26 58.4 21.2 87

RP 18 MeOH 0 85.7 55.5 71.9 25.6 19.8

RP 18 .5 acetone 31.4 0 0 25.6 36.6 88

Σ(RP18) .4 .2 31 85.7 55.5 97.5 62 108.6

EN MeOH 99.5 68 42.3 23.2 24.3 0

EN acetone 0 33 15.4 0 0 37.7

Σ(EN) 99.5 102 57.7 23.2 24.3 37.7

Table 4.1 leads to Figure 4.2 where the recoveries of benzothiazoles extracted on different

hases and eluted with methanol followed with acetone.

p

53

0

20

40

60

80

100

120re

cove

ry (%

)BTSAOHBTBTMTBTMBTTCMBT

Envi 18 Envi+ Envi-ChromP RP 18 EN

Figure 4.2: SPE on different phases by pH 7 (100 ml extraction volume, tap water).

and EN phases eluted with methanol or acetone. Concentrations ranged from

g/L.

extraction (M1) on RP18 /MeOH

Concentration 234-272 µg/L.

Figure 4.3 displays some chromatogram examples of the extraction using different phases such

as RP 18, Envi+

234-272 µ

a

Figure 4.3 a: : Chromatogram examples of

54

P18 /MeOH;

1 on Envi+ / MeOH and on Envi+ /acetone.

Concentration 234-272 µg/L.

and OHBT are well recovered from the polar polymer phases, namely from the EN-

b

c

1

2

Figure 4.3: Chromatogram examples of extraction on different sorbents. a: M1 on R

b: M1 on EN / acetone; c: (1) M

4.1.2 Sequential solid-phase extraction

From the results shown above, it can be concluded that the polar benzothiazole compounds

BTSA

phase.

The other analytes show good recoveries on the silicagel-cartridges. However, the RP 18 and

EN-phases at pH 7 brought better results as the other solid-phases. Based on their high

55

capacity, RP 18 and EN were chosen for a sequential solid phases-extraction of benzothiazole

ompounds, followed by a sequential elution (methanol followed with acetone; see Figure 4.4).

igure 4.4: Sequential extraction of benzothiazoles

n

actory recoveries. Particularly in the case of acetonitrile. BT and MTBT were very

c

F

4.1.3 Other eluents

Before the sequential extraction, RP 18 and EN-solid phases should be tested with other elutio

milieus as dichloromethane and acetonitrile in order to improve the recoveries of the studied

benzothiazoles from the RP18 phase. Results of this investigation are presented in Table 4.2.

The attempt of elution with use of acetonitrile and dichloromethane (Table 4.2) led to very

unsatisf

badly recovered. By employing dichloromethane, only OHBT and MBT were acceptable (>

50%).

56

Figure 4.5 displays some chromatograms of benzothiazoles extracted on PR18 and eluted with ifferent eluents.

T e iff ti edium

d

able 4.2: R covery of extraction with d erent elu on m .

cartdridge Elution medium BTSA OHBT BT MTBT MBT TCMTB

EN MeOH 99.35 .2 .5 68 41.4 23 22 0

EN Aceton 0 33 14.2 0 0 38.1

RP18 MeOH 0 70.5 .3 50 72.01 24.8 19.8

RP18 Aceton 32.6 0 0 23.3 34.8 85.8

RP 18 CH3CN 80.2 42.45 5.5 15 60.5 44.6

RP 18 CH2Cl2 36.2 58.51 28.45 37.8 55.4 22.3

a

b

c

Figure 4.5: Chromatogram examples (M1) with different elutions on RP18.

(a): MeOH; (b): acetone; (c): CH2Cl2

57

4.1.4 pH influence on the extraction

In order to determine the pH-influence on the extraction (Fig.4.6), further samples were also

xtracted at different pH-values of 4, 6, 7 and 9.

d at any pH-value except

ith neutral-pH where its extraction reaches a recovery of ca. 64 %.

BT and MTBT were recovered at all pH-values to > 50%, BT was

etected at pH 6 < 40%.

benzothiazoles at neutral pH better

covery rates were achieved than at the other pH-values.

e

Figure 4.5 shows that the recovery rates are dependent on the pH-value, particularly and

strongly in case of MBT which was not extractable with any phase an

w

While BTSA, OHBT, A

d

At pH 9, the recovery of TCMTB is < 40% because TCMTB hydrolyses at elevated pH-values.

Furthermore, it can be observed that for all investigated

re

0

20

40

60

80

100

120

BT

reco

very

(%)

pH 4pH 6pH 9pH 7

BTSA ABT OHBT MBT BT MTBT TCM

Figure 4.6: Recovery of sequential extraction and sequential elution of benzothiazole

compounds at different pH- values.

58

4.1.5 Application on wastewaters

Effluent and influent samples of the municipal sewage treatment plant Ruhleben were

extracted with the optimised method, that means sequential extraction and sequential elution at

pH 7. In all cases, the peak areas and retention times were compared with those obtained from

to run blanks due to the widespread presence of disturbing substances in

e studied samples.

ll studied benzothiazoles were

ass-

luent of sewage treatment plant “Ruhleben” are

presented in figures 4.7-4.9 (see below).

rference with other

oextracted contaminants was observed (4.7-4.8).

acceptable and are full

satisfied as shown in Tables 4.3 and 4.4 and displayed in Figure 4.9.

pure standard solutions.

It was also necessary

th

For the sample extraction, 2.3-2.7 µg/L spiked as a mixture of a

used. Recoveries were calculated and are reported in figure 4.9.

Before extraction, the samples were filtered over 0.45 µm cellulose-acetate filters with gl

fibre pre-filters. In this procedure, 50 ml of influent and 100 ml of effluent were extracted.

The extraction results of influent and eff

re

From these experiments, it was found that during chromatographic determination, unknown

compounds, considered as disturbing substances, appeared in great amounts in effluent and

influent at the same retention time as BTSA. Based on its retention time in the

chromatographic system, this compound (or those compounds) must be of a rather high

polarity, like BTSA. For this reason the value of BTSA (Fig. 4.7, 4.8) was not determinable.

The other benzothiazole components are clearly detected and no inte

c

The recovery rates of the SPE-extraction from STP effluent and STP influent are slightly

decreased compared to those obtained with tap water but remained

59

MeBT

MTBTBT

OHBT

BTSA ?MeBT

ABT

a

b

Figure 4.7: UV-Chromatograms example of spiked effluent samples on EN-phase eluted with

methanol. a: M1 (BTSA,OHBT, MBT, BT, MTBT); b: M2 (ABT and TCMTB).

ABTMeBT TCMTB

Figure 4.8: UV-Chromatogram example of spiked influent sample on RP 18-cartridge eluted

with methanol (M2: ABT and TCMTB).

60

Table 4.3: Concentration of benzothiazoles in effluent (10 µl injection)

Benzothiazoles Spiked

(µg/L)

Found (µg/L)

in influent *

Blank

(µg/L) *

Found in blank

(µg/L) *

Concentration

(%) *

BTSA 242 nq. (++) 0 nq. (++) nq. (++)

OHBT 255 248.6 0 << 97.5

ABT 264 259.2 0 << 98.2

BT 263 197 0 - 68.1

MBT 234 54.3 0 87 60.4

MTBT 255 163.6 0 85 97.5

TCMTB 250 131.2 0 85 86.5

* eluted with methanol followed by acetone ( sum of concentrations)

nq. ( ++) non quantified (great amount); << negligeable; - no trace

Table 4.4: Concentration of benzothiazoles in influent (10µl injection)

Benzothiazoles Spiked

(µg/L)

Found

(µg/L) in

influent *

Blank (µg/L)

*

Found in

blank (µg/L)

*

Concentration

(%) *

BTSA 242 nq. (++) 0 nq. (++) nq. (++)

OHBT 255 251.2 0 << 98.5

ABT 272 265.2 0 << 97.5

BT 263 158.3 0 - 60.2

MBT 234 53.3 0 85 59.1

MTBT 255 158.2 0 92 98.1

TCMTB 250 107.2 0 99 82.5

* eluted with methanol followed by acetone ( sum of concentrations)

nq. ( ++) non quantified (great amount); << negligeable; - no trace

61

0

20

40

60

80

100

120

BTSA ABT OHBT MBT BT MTBT TCMBT

Rec

over

y (%

)effluentinfluent

Figure 4.9: Recovery in influent and effluent samples.

To evaluate the results of the extraction efficiency obtained in our study, we compared these

later to those of other workers. But it is mentioned that other sorbents were used in their

studies.

BT and MTBT have also been determined by using SPE but on C18 phases within the toxicity

fractionation procedures to surface waters (Burkhard et al., 1991; Amato et al., 1992). On the

other side, Reemtsma et al. (1994) used a mixture of solvents as eluent and found that 93 % of

MBT was recovered from wastewater on C18 phases (Su 18, SuEN) with a methanolic solvent

system (2 ml methanol, 2ml methanol-dichloromethane (20:80, v/v) while TCMTB was

equally extractable by C8 phases (Su8), cyano (SuCN) and C18 phases (Su18) with the

acetonitrile containing 2 ml acetonitrile, 2 ml acetone and 2 ml toluene. TCMTB was

extractable around 92% on the cyano-phase using the methanolic solvent system. Under those

conditions, however, the recoveries of MTBT and BT from wastewater were considered as

insufficient (Reemtsma et al. 1994). BT was poorly recovered in all phases (87%). The

recoveries of MTBT was 95%. The authors explained that the extremely poor recovery of BT,

the most volatile derivative, might be due to the drying procedure applied before LC analysis.

62

4.2 Aerobic degradation of some benzothiazoles in water samples

In order to elucidate if aerobic mixed cultures are capable of utilising benzothiazoles as a sole

source of carbon, nitrogen and energy, the aerobic biodegradation of BT, OHBT, MBT and

MTBT was investigated in separate batch tests, using activated sludge from Ruhleben (Berlin)

as inoculum.

4.2.1 Substrate with biomass (sludge) and in presence of glucose

Test results of individual substrates in batch cultures are represented in Figures 4.10-4.11 and

Figures 4.13- 4.14. They show that some benzothiazoles are clearly biological degradable. This

is indicated either by the decrease of the substrate concentration over the time or proved by the

formation of metabolites in the case of MBT (Fig. 4.14).

In the beginning of the aerobic essays the batch cultures of all benzothiazoles studied were not

turbid. However, over prolonged period of time the tests of BT and OHBT became turbid.

Bacteria grew aerobically forming pink colonies. Thus the culture present in the sludge, proved

to be able to degrade benzothiazoles in this culture and was able to grow on BT and OHBT and

less on MBT. Whereas no bacteria growth was observed in the case of MTBT, where a clear

solution remained assuming that the culture present is not able to degrade MTBT.

Similar observations were made by De Wever et al.(1997) and Gaya and Knapp (1977) in the

case of BT and OHBT. The mixed culture became turbid and the selected pure culture (strain

PA) grew aerobically forming pink smooth circular colonies.

Figures 4.10 and 4.11 show that both BT and OHBT can be microbially degraded in water in a

relatively short time. BT and OHBT disappeared totally, whereas MBT was hardly

metabolized (Fig. 4.14). Contrary to MBT, no stable metabolites were detected in the case of

BT (Fig. 4.10) and OHBT (Fig. 4.11). BT was not observable after 20 days and its

biodegradation was noted after 2 days and was complete in 17 days. The same remark

concerning OHBT was observed but the OHBT degradation was even faster than that of BT

(Fig. 4.11) and OHBT completely disappeared in 7 days.

63

0

1

2

3

4

5

6

7

0 5 10 15 20 25 30 35time(day)

µg/m

l

0

0,1

0,2

0,3

0,4

0,5

0,6

UVm

ax-A

BS

µg/mlUV-ABS ( 284 nm)

HS

N

Figure 4.10: BT aerobic degradation test

0

0,5

1

1,5

2

2,5

3

0 5 10 15 20 25 30 35timet (day)

µg/m

l

0

0,05

0,1

0,15

0,2

0,25

0,3

0,35

UVm

ax-A

BS

µg/ml

UV-ABS (242 nm)

S

NOH

Figure 4.11: OHBT aerobic degradation test.

64

The results presented of this study confirm previous reports (Mainprize et al., 1976; De Vos et

al., 1993; De Wever et al., 1992) of the biodegradability of BT and OHBT and confirm also

those obtained by Reemtsma et al. (1995), where BT was completely degraded after 8 days,

and Gaya and Knapp (1997) who observed biodegradation in mixed cultures of BT in 10-12

days. BT and OHBT have generally been considered to be easily biodegradable (Chudoba et

al., 1977) and axenic cultures degrading these compounds are available. Pure cultures

(Rhodochrous) were isolated on OHBT (De Wever et al., 1997) and rhodococcus strain A was

isolated on BT (Gaya and Knapp, 1977). Thus it seemed that at least the OHBT and BT

degradation pathways are closely related. Essays of the key enzymes of the ortho- and meta

cleavage pathways of aromatic metabolism strongly suggested that the meta cleavage pathways

is involved in the degradation of BT and OHBT (Gaya and Knapp, 1997). However, it was not

clear whether the ring cleavage compound is catechol or a substituted derivative. Therefore,

interestingly, De Wever (1995) found elevated levels of C23 (key enzymes involved in

cleavage of catechol in the ortho and meta-cleavage pathways) in OHBT degrading pure

culture and came to similar conclusions as Gaja and Knapp.

Biodegradation of organic compounds comprising benzothiazoles under aerobic conditions

most often occurs when bacteria catalyze the breakdown of these molecules and then recover

some of this chemical energy as ATP (adenosine triphosphate) which is absolutely necessary

for maintenance of the bacterial cell (Wiedemeier et al., 1995A). ATP is generated through a

series of oxidation-reduction reactions (the electron transport chain) where electrons are

sequentially transferred from one compound, the electron donor, to an electron acceptor.

Dissolved oxygen is usually the preferred electron acceptor for the degradation of organic

compounds by microbes as it often provides the greatest energy yield (Wiedemeier et al.,

1995A). Dissolved oxygen concentrations of 1 mg/L or greater are considered to define aerobic

conditions. During aerobic respiration, the oxygen present in the environment is converted to

water and thus the dissolved oxygen content can decrease. Thermodynamically, the reduction

of molecular oxygen to water is very favourable for the participating microorganisms.

From literature, it is known that from the heteroatoms of benzothiazoles ammonium and

sulphate are formed (Gaja and Knapp, 1997) (Fig. 4.12).

65

S

NH

S

NOH

BT OHBT

aerobic conditions-?--?-

NH4+ + SO4-- Figure 4.12: Degradation of BT and OHBT under aerobic conditions

forming NH4+ and SO4

-- (Gaja and Knapp, 1997)

In their experiments (Gaja and Knapp, 1997), mineralization (Fig. 4.12) is the final step in

biodegradation of BT and OHBT compounds, and it is probably always preceded by several

intermediary steps, defined as metabolite formation.

The metabolism of heterocyclic compounds very often involves ring hydroxylation, followed

by ring cleavage. When attached to six-member-rings, five-membered-rings are usually

cleaved first, with or without initial hydroxylation (Hill and Wright. 1978; Junker et al., 1994).

While the degradation of BT and OHBT is very fast, MTBT is not degraded under the same

conditions (Fig.4.13). The persistence of the MTBT compound was also observed in

wastewater by Reemtsma et al., 1995), in effluent municipal sewage treatment plant (Ellis et

al., 1982; Clark et al., 1991; Amato et al., 1992) and in surface water by Dietrich et al. (1988)

and Braunstein et al. (1989).

In view of a comprehensive study of benzothiazole-biodegradation, we also wanted to include

a MBT degradation study. Although no degradation of MBT (Fig.4.14 a and b) is observed

under the chosen aerobic conditions, MBT is transformed to MTBT. Here yet after 3 days a

diminution to 56.7 % of employed MBT is observed. After 20 days, the concentration (µg/ml)

of the product (MTBT) is higher than that of the substrate (MBT).

66

3

4

5

6

7

8

9

10

0 5 10 15 20 25 30 35

time (day)

µg/m

l

0

0,5

1

1,5

UVm

ax-A

BS

µg/ml

UV-ABS (278 nm)

SS

NCH3

Fig. 4.13: MTBT aerobic degradation test

0

0,5

1

1,5

2

2,5

3

3,5

4

4,5

0 5 10 15 20 25 30 35time (day)

µg/m

l

0

0,1

0,2

0,3

0,4

0,5

0,6

0,7

0,8

UVm

ax-A

BS

MBTMTBT (metabolite)UV-ABS MBT (313 nm)

SHS

N

Fig. 4.14 a: MBT aerobic degradation test (µg/L)

67

Then no further transformation occurred. In 25 days, the concentration (µmol/L) of MBT is

equal to that of the metabolite (MTBT) (Fig. 4.14 b). After 30 days 75% of MBT are

methylated to MTBT, while 25% (MBT) remained unaltered.

These data probably indicate that specific enzymes involved in OHBT and BT oxidation but

other enzymes involved in MBT methylation.

µmol/ MBT = µmol/L MTBT

0

5

10

15

20

25

0 5 10 15 20 25 30 35time (min)

µmol

/L

MBTMTBT(metabolite)sum µM

SHS

N

Figure 4.14 b: MBT aerobic degradation test (µmol/L)

None of the mixed cultures, could utilise MBT as a growth substrate, possibly due to its well-

known toxicity (see for example, Mainprize et al., 1976 or De Wever et al., 1994). However,

as previously suggested by the work of Mainprize et al. (1976), some culture used for the

degradation of other benzothiazoles can at least partly oxidise MBT. The observed decrease of

MBT concentration is probably due to its oxydation to MBTS.

The above benzothiazole degradations are confirmed by UV-absorbance spectrum and shown

on Figures 4.10-4.11 and 4.13-4.14 a as UVmax versus the time.

Each compound was measured at its specific UV-maximum absorption. When comparing the

two curves (UVmax/time and benzothiazole concentration/time), a similar trend is observed.

The UV-max decreased along the period of incubation assuming so an eventual transformation

of these compounds.

68

De Wever and Verachtert (1994) suggested that MBT could be oxidized by activated sludge to

BTSA or to MBTS, but the production of yellow-coloured products in their study suggested

that some modification to the ring structure of MBT has occurred.

Hence, these different results concerning MBT between our study and that of De Wever and

Verachtert (1994) may due to different populations of bacteria that reside in the test experiment

or other factors (e.g. salts, oxygen amount etc..). However, our results approached those of

Reemtsma et al. (1995) who investigated the aerobic transformation of MBT over 28 d. 10% of

the MBT was methylated to MTBT by the bacteria, while 87% remained unaltered (Reemtsma

et al., 1995).

Methylation of aromatic thiols by bacterial S-methyltransferase is reported and was suggested

as a method of detoxification, because methylthioethers are generally considered less toxic

(Drotar et al., 1985). MBT is additionally suggested to be eliminated by adduct formation with

biogenic compounds; a process that was reported with digestive enzymes of mammals (Larsen

et al., 1988; Fukuoka et al., 1987; Elfarra et al., 1990).

In addition, the formation and measurement of metabolites can definitively show the

transformation of the benzothiazoles of interest. However, results from laboratory can be

greatly influenced by many factors such as the source, collection, and condition of the

incubation sample (e.g. substrate concentration, temperature), and the length of the study

period (and its effect particularly on the initial microbial population during a long study period)

(Wiedemeier et al., 1996).

Some studies demonstrated that microorganisms can remove heterocyclic nitrogen compounds,

without attacking aromatic or aliphatic hydrocarbons (Aislabie and Atlas, 1989). Heterocyclic

nitrogen compounds are not recalcitrant to microbial attack; bacteria have previously been

reported to biotransform carbazoles, pyridines etc.. (Benett et al., 1985; Fedorak et al., 1984;

Finnerty, 1982; Finnerty et al., 1989).

Figure 4.15 represents the degradation of aniline under the aerobic conditions for comparative

purposes, because aniline degrades easily under aerobic conditions.

69

0123456789

0 5 10 15 20 25 30time (day)

µg/m

l

00,20,40,60,811,21,41,6

UVm

ax-A

BS

µg/mlUV-ABS (280 nm)

NH2

Figure 4.15: Aniline aerobic degradationtest (as reference)

4.2.2 /4.2.3 Substrate without biomass (abiotic)/Sustrate in absence of glucose (not active)

To ensure that the degradations observed in 4.2.1 are biotic (biodegradation) and not due to

abiotic or transport processes, appropriate controls (analogous tests) under sterile conditions

(without inoculum addition) and without glucose were carried out, to detect an eventual

chemical transformation. In addition, autoclaving may not be totally suitable, probably due to

incomplete sterilisation. The initial concentrations in this case were 1.76 mg/L for BT, 1.64

mg/L OHBT and 3 mg/L of MBT.

From figures 4.16, 4.17, and 4.18 (see below), it is evident that without inoculum (biomass) no

degradation took place. These controls lacking an inoculum failed to show any decrease in all

studied benzothiazoles, confirming that the degradation in the experimental cultures was

biological rather than of physical or chemical origin. The diminution of concentration at day 1

(without glucose) is probably due to adsorption.

The experiments without glucose-addition show slower degradation of BT and OHBT as those

of aerobic tests (Figures 4.10 and 4.11). BT and OHBT were still degraded without problems.

BT was after 20 days and OHBT after 8 days completely degraded. Also in this case no

metabolites were detected. Glucose was used as cosubstrate. In the absence of glucose, the BT

and OHBT concentration (Fig. 4.16 and Fig. 4.17) decreased less rapidly and their

transformation became very slow after 6 days and 1 day respectively. However, benzothiazoles

70

incubated without glucose were rapidly degraded, confirming that bacteria use benzothiazoles

as source of carbon and energy.

00.20.40.60.8

11.21.41.61.8

2

0 10 20 30 4time(day)

mg/

L

0

under aerobic conditions

w ithout glucose

w ithout biomass (abiotic), sterile andautoclaved

Figure 4.16: BT aerobic degradation test in the absence of glucose and under abiotic

conditions.

0

0.2

0.4

0.6

0.8

1

1.2

1.4

1.6

1.8

0 5 10 15 20 25 30 35 40tim e (day)

mg/

L

under aerobic conditions

w ithout glucose

w ithout biomass (abiotic), sterile andautoclaved

Figure 4.17: OHBT aerobic degradation test in the absence of glucose and under abiotic

conditions

71

Because the structures of BT and OHBT do not have hydrolysable or reactive moities (like

ester bonds), abiotic chemical transformations should not occur.

For MBT, without glucose addition, (Fig. 4.17) the same results are obtained, as MBT was

transformed to MTBT as seen in the previous experiment (4.5).

For the formation of “stable metabolites” from MBT (Fig. 4.18), two explanations are possible:

either the bacteria cannot produce important enzymes to degradation, or the formed metabolites

are toxic and obstruct or inhibit the bacteria. No degradation in sterile control was visible.

0

0,5

1

1,5

2

2,5

3

3,5

4

4,5

0 5 10 15 20 25 30 35time (day)

mg/

L

under aerobic conditionswithout glucosewithout biomass, sterile and autoclavedMTBT as product (under aerobic conditions)MTBT as product (without glucose)

Figure 4.18: MBT aerobic degradation test in the absence of glucose and under abiotic

conditions

An aerobic degradation essay was performed by using a substance which can catalyse the

transfer of electrons from an enzyme to a potential substrate.

Figure 4.19 and Figure 4.20 reveal the presence of dithiobisbenzothiazole (MBTS) as oxidation

product of MBT in both cases (abiotic and biotic conditions). After 35 days 10% of the

instance MBT were found as MBTS. Traces of MTBT appeared in the abiotic essay (Fig.

4.20) but this compound (MTBT) was practically absent in the case of the biotic test. The

decrease of the MBT concentration is probably due to the adsorption of MBT on the biomass.

72

Methylation of aromatic thiols by bacterial S-methyltransferase is reported and was suggested

as a method of detoxification, because methylthioethers are generally considered less toxic

(Drotar, 1985). From this experiment it was shown that the anthraquinone sulfonate (AQS)

addition was not helpful.

0

20

40

60

80

100

120

0 10 20 30 40tim e(day)

mm

ol/L

02468101214161820

µg/m

l

MBT substrate (µmol/ml)MBTS product µmol/Lsum (µmol/L)MBT substrate (µg/mL)MBTS product (µg/ml)

Figure 4.19: MBT aerobic degradation test in presence of Antraquinone-sulphonate and

biomass

0

20

40

60

80

100

120

0 5 10 15 20 25 30 35time (day)

µmol

/L

02468101214161820

µg/m

l

MBT substrate (µmol/L) MBTS product (µmol/L)MTBT product (µmol/L) Sumµg/ml MBT µg/ml MBTSµg/ml MTBT

Figure 4.20: MBT aerobic degradation test with antraquinone sulfonate (abiotic essay).

73

4.2.4 Biodegradation of mercaptobenzothiazole (MBT) in the Sequencing-Batch Reactor

(SBR)

MBT is a toxic and inhibitory substrate, but also a carbon source for an acclimated biomass.

One of the consequences, of toxic compounds in biological wastewater treatment is process

instability which can lead to a wash-out (or dissolution) of the effective organisms, sometimes

providing little warning of the impending failure (Gaudy and Rozich, 1982; Rozich et al.,

1985).

In order to minimise any potential toxic effects due to the presence of MBT, reactors 1 and 2

were initially fed as described in 3.4.4.

The influent from a sewage treatment plant served as influent 1 to reactor 1. The initial

concentration of MBT was 24.94 mg/L. The reactor 1 contained 1 L of sewage water and 10 ml

of bacteria sludge, whereas the reactor 2 contained 500 ml tap water and the same amount of

bacteria sludge. The stability of the systems was judged by the constant UV-absorption along

with a stable biomass concentration.

Reactor 1 had a high DOC (110 mg/L) and reactor 2 had a low DOC (20 mg/L). At this point

(day 1 of MBT addition), in feed solution, 25 mg/L MBT (12.6 mg/L DOC) was introduced in

the reactor 1. The pH, DOC, UV and HPLC-results are represented in following curves.

The reactor pH was within the range 7.4 to 8.6 (pH of influent, effluent 1 and 2 ), throughout

the study as shown in Figure 4.21. The DOC-contents of the SBR experiment from the start to

the end of MBT spiking ranged from ca. 90 to110 mg/L in the case of influent, from 10 to 30

mg/L in effluent 1 and from 5 to 10 mg/L in effluent 2 (Fig. 4.22). Carbon balance prior and

after MBT addition is presented in Table 4.5.

After MBT addition, the UV-absorption value at 254 nm in influent 1 increased from 1.1 to 2.5

cm-1 compared with those of effluent 1 and effluent 2 where a constant UV254-value was noted.

Knowing that MBT does not absorb at this wavelength but only BT posses a second maximum

at 254 nm, this increase might be explained by the presence of substances absorbing at 254 nm

in influent 1. On the other hand, between 0 and 15 days the UV-absorption in influent1 and

effluent 2 remained slightly constant in the range of 0.5- 0.4 cm-1 and 0.4 -0.5 cm-1 respectively

(Fig.4.23). While the effluent 1 revealed an augmentation of 0.5 cm-1 (double) to rich 1 cm-1.

74

6.8

7

7.2

7.4

7.6

7.88

8.2

8.4

8.6

8.8

0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18time (week)

pH-v

alue

inf luent 1effluent 1effluent 2

Figure 4. 21: pH-value of SBR experiment

before spiking start of MBT spiking end of MBT spiking

0

20

40

60

80

100

120

-5 0 5 10 15 20time (week)

mg/

L

influent 1effluent 1effluent 2

Figure 4. 22: DOC-SBR experiment.

75

before spiking MBT spiking

0

0,5

1

1,5

2

2,5

3

-6 -4 -2 0 2 4 6 8 10 12 14 16 18 20time (week)

UV-

abso

rptio

n (2

54 n

m)

inf luent 1

eff luent 1

eff luent 2

Figure 4.23: UV-absorption of SBR

Table 4.5: Carbon balance prior and after MBT addition reactor Prior MBT addition after MBT addition DOC

(mg/L) UV254 (cm-1)

SAK254/DOC (mg-1 L m-1)

DOC (mg/L) UV254 (cm-1)

SAK254/DOC (mg-1 L m-1)

Influent 1 99.4 1.1 1.1 103 2.5 2.4 Effluent 1 11.3 0.5 4.4 22 0.9 4.1 Effluent 2 9.4 0.4 4.3 13.8 0.5 3.6

4.2.4.1 Benzothiazoles in influent 1

The HPLC-results of influent 1 (Fig. 4.24) show that the initial MBT concentration of about

150 µmol/L followed a gradually decreasing trend as the operation of the system progressed.

However, no remarkable difference in the MBT degradation pattern was observed in the

influent 1 between 2 and 16 weeks and the MBT content eventually stabilized around the 70

µmol/L level. After 16 weeks, end of MBT spiking (Fig. 4.24), no trace of MBT was detected (

no MBT addition after 16 weeks).

The strong diminution observed between 0 and 2 weeks is probably due to the sorption (40 %

sorption). That means only 60 % MBT were recovered.

76

Explanation is perhaps that one part of MBT in contact with the biomass and in presence of

oxygen oxidises easily to give BTSA and traces of MBTS as shown in Figure 4.24. After one

week the concentration of BTSA (14 µmol/L) remained unchanged to the end of the

experiment. The concentration of MBTS was very low (traces).

All benzothiazoles found either BTSA, MBTS and also the substrate MBT remained constant

between 3 and 16 weeks.

From Figure 4.24, the disappearance of MBT present in influent 1 between the start and the

end of MBT addition was of 50 %, giving 9.4 % BTSA and traces of MBTS. That means a

40.5 % elimination. Table 4.6 summarises the behaviour of MBT in influent 1 between 3 and

16 weeks. No degradation was noted but MBT was washed out into the influent 1.

0

20

40

60

80

100

120

140

160

0 5 10 15 20time(week)

µmol

/L

MBTMBTSBTSASum

start of MBT spiking

end of MBT spiking

Figure 4.24: HPLC-results of influent 1.

77

Table 4.6: Influent 1 between 3 and 16 weeks

Week MBT removed

(%)

MBTS found

(%)

BTSA found

(%)

Σ benzothiazoles

found (%)

MBT sorption (%)

3 52 0.6 8.8 58 42

16 50 0.7 9.4 60 40

4.2.4.2 Benzothioazoles in reactor 1 (effluent 1 or influent 2)

During MBT addition in the reactor 1 (100 % of MBT addition (24.94 mg/L)) (Fig. 4.25), a

surprising result was observed because no signs of MBT appeared in the effluent 1 (effluent of

reactor 1). Instead, 60 µM of BTSA was detected in this time.

0

20

40

60

80

100

120

0 5 10 15 20week

µmol

/L

MTBTOHBTBTBTSASum

start of MBT spiking

end of MBT spiking

Figure 4.25: Concentrations in effluent 1 (reactor 1)

78

An important period is to noted at the end of spiking (after 16 weeks). BTSA and MTBT were

not degraded after this time, whereas BT and OHBT were degraded.

Comparison with total dissolved concentration in influent 1

We note also that the loss of MBT concentration (90 µmol/L) (total disappearance) (Fig. 4.25)

comparing to the total concentration of benzothiazoles found (sum = 60 µmol/L) was very

large and corresponded to 60%. Additionally, some benzothiazoles as OHBT, BT, and MTBT

were produced.

The first evidence of BTSA degradation appeared after the first week. However, the

transformation of BTSA was very rapid as the study progressed (1 to 3 weeks). This

degradation still occurred but less rapidly (3 to 7 weeks), lasting for about 7 weeks. It can be

assuming, that this decrease in BTSA produced BT and OHBT.

OHBT and BT are formed continuously and degraded continuously. During the next 16 weeks

of the operation, no measurable OHBT or BT occurred. This corresponds to the behaviour in

the batch tests.

We assume that a part of MBT was methylated to MTBT, which was present after the first

week (Fig. 4.25) and another part was oxidised to BTSA. This confirms hereby what has been

reported in the literature (De Wever and Verachtert (1994), who proposed an eventual MBT

oxidation by activated sludge. However, MTBT proved also to be stable and not degradable in

this SBR experiment.

From Figure 4.25 which represents the concentration of benzothiazole in the effluent 1 as a

function of time Table 4.7 is derived on the mass balance of benzothiazoles found in this

effluent 1.

Table 4.7: Mass balance of benzothiazoles in reactor 1(period 7-16 weeks) week MBT

added

(µmol/L)

MBT

found in

influent1

(µmol/L)

BTSA

effluent1

(µmol/L)

MTBT

effluent1

(µmol/L)

OHBT

effluent1

(µmol/L)

BT

effluent1

(µmol/L)

Σ

benzothia

zoles

found in

effluent 1

(µmol/L)

Differenc1

max (%)

difference2

min (%)

7 149.5 88.5 30.22 17.98 6.87 11.54 66.61 55 25

16 149.5 88.5 30.22 17.95 6.63 11.34 66.14 56 25

79

4.2.4.3 Benzothiazoles in effluent 2 (reactor 2)

After a complete MBT disappearance (degradation?) was established in reactors 1, the second

reactor 2 (Fig. 4.26) indicated the presence of BTSA and MTBT only. BT and OHBT were not

detectable in this case. This is due probably to the low concentrations of these products

measured in effluent 1. BTSA and MTBT contents remained constant during the period of

study.

0

10

20

30

40

50

60

70

0 5 10 15 20time (week)

µmol

/L

MTBTBTSASum

s tart of MBT spiking end of MBT spiking

Figure 4.26: Concentration in effluent 2 (reactor 2)

From the above transformations, we can summarise the SBR transformations as follows

(Fig.4.27):

80

MBT

(149.5 µmol/L)

MBTS (sorption) ?

Influent 1

Sorption (0.7 µmol/L) BTSA MBT

to sluge (13 µmol/L) (75 µmol/L)

?

Reactor 1

sorption

BTSA MTBT BT OHBT

(30 µmol/L) (18 µmol/L) (12 µmol/L) (7µmol/L)

Reactor 2

BTSA MTBT mineralisation

(52 µmol/L) (11 µmol/L) (0 mmol/L)

Partial mineralisation

Figure 4.27: SBR transformations in reactor 1 and reactor 2

An increase in BTSA concentration in effluent 2 (~50 µmol/L) compared to effluent 1 (~ 30

µmol/L) (Fig. 4.26; Fig. 4.27) was noted. This increase of BTSA in reactor 2 might originate

from the sludge. Additionally, this increase remained constant between two weeks after the

start of the SBR operation and the end of the operation.

Concerning MTBT, the concentration of this product increased in 14 weeks between the

second week (~ 6 µmol/L) and the 16th week (~11 µmol/L) and reached 8.9 % of

augmentation. But compared to effluent 1 (~18 µmol/L MTBT), a diminution of 7 µmol/L of

this compound was noted. This decrease on MTBT would be explained by a partial

mineralization.

The sum of BT and OHBT (18 µmol/L) detected in effluent 1 were probably mineralized.

81

Results from the SBR experiment indicated that losses of MBT were considerable. It should be

noted that the main mechanism of MBT disappearance is probably the biological

transformation. Volatilization appears to be insignificant, but adsorption on biomass appeared

to be important.

The detection of metabolites or intermediate products, MTBT, OHBT, BT, and BTSA,

throughout the study suggests that MBT biotransformation may have occurred.

82

4.3 Anaerobic degradation of some benzothiazoles

Transformations of benzothiazoles were also investigated under anaerobic conditions in serum

bottles with an organic carbon source (ethanol) at initial concentrations ranging from 0.1 to 0.5

mg/L. Samples were taken every week under N2.

4.3.1 Mixture of benzothiazoles

Figure 4.28 shows the behaviour of a mixture of 396.6 µg/L of BT, 131 µg/L of MBT and 46.6

µg/L of OHBT. No alteration of OHBT was found, while a slight increase of BT-content and a

decrease of MBT was observed (Fig. 4.28). MBT was not detectable after 4 weeks.

The elevation of the BT concentration (3.4 µmol/L) during the period of anaerobic incubation

was not clear but such phenomenon might be related to the decrease of MBT concentration.

Additionally, no other types of benzothiazoles as metabolites were detected in this experiment.

0

50

100

150

200

250

300

350

400

450

500

0 5 10 15 20trime (Week)

µg/L

BT OHBT MBT

Figure 4.28: Anaerobic behaviour of a benzothiazole mixture.

83

4.3.2 Individual substances

The experiment results of individual substances in anaerobic batch cultures are depicted in the

Figures 4.29-4.34. They show that some of studied benzothiazoles have been clearly altered in

anaerobic tests.

OHBT (Figure 4.29) was not transformed, as previously observed in case of the benzothiazole

mixture (Figure 4.28). MTBT was partly degraded (Figure 4.30) under the anaerobic

conditions. After 6 weeks only 311µg/L (67.9 % ) of MTBT were found, but no metabolite was

detected by HPLC-UV analyses.

Through the anaerobic treatment, MBT was partly metabolised to BT. This confirms the results

of the mixture degradation test (Figure 4.28). After 4 weeks, MBT is not detectable (Figure

4.31). It is also important to note that in the fifth week, not only no MBT was detected but also

only a very small quantity of BT was present. That means a loss of MBT estimated at ca. 96%

(1.86 µmol/L) in 5 weeks. Whereas at end of the anaerobic test this loss was only 46 % (1.37

µmol/L). This could be interpreted either by slow production of BT, or the accumulation of

MBT compound on the sludge. And certainly by the formation of unknown products, not

identified by HPLC-UV.

496

498

500

0 5 10 15 20time (Week)

OH

BT

((µg/

L)

Figure 4.29: Anaerobic biotransformation of OHBT.

S

NOH

84

050

100150200250300350400450500

0 2 4 6 8 10 12 14 16 1time (Week)

µg/L

MTB

T)

8

Figure 4.30: Anaerobic biotransformation of MTBT.

SS

NCH3

0

0,5

1

1,5

2

2,5

3

3,5

0 2 4 6 8 10 12 14 16time (week)

µmol

/L

MBTBTsum

SHS

N

sorption orunknown intermediate

HS

N

Figure 4.31: Anaerobic biotransformation of MBT

85

The results obtained below (Fig. 4.32) under the same anaerobic conditions show that BTSA is

partially transformed to OHBT, which was not observable earlier than 5 weeks together with

other unknown substances that were not identified.

After 12 weeks only 29.7 % (0.83 µmol/L) of BTSA remained in the reactor and the

concentration of this compound was about the same as that of the product OHBT (0.88

µmol/L). BTSA continued to decrease after this time but slower as in the beginning of the

anaerobic incubation. At the end of the experiment the OHBT concentration in terms of

percentage reached 32.3% (0.9 µmol/L) while 26 % of BTSA remained (0.72 µmol/L). That

means a loss of BTSA of 58 % (Table 4.8) so 42 % were transformed to unknown products or

were mineralized.

0

0,5

1

1,5

2

2,5

3

0 2 4 6 8 10 12 14 16 1time (week)

µmol

/L

8

BTSAOHBTsum

N

SSO3H

S

NOH

Figure 4.32: Anaerobic biotransformation of BTSA

86

TCMTB was transformed to BT and MBT (Figure 4.33). The product MBT reaches its

maximum of 69 % (1.75 µmol/L) after 10-11 weeks. Then a diminution of this compound was

noted together with a continuous increase of BT. MBT disappeared completely after 15 weeks.

In the seventh week no TCMTB remained and 42 % (1.06 µmol/L) occurred as MBT and 6.7

% (0.17 µmol/L) as BT (Fig. 4.33). At the end of anaerobic treatment a loss of 47 % was

observed (Table 4.8). In the case of TCMBT, BT is, besides MBT, therefore an important

metabolite. It is recorded through the present anaerobic study that BT and OHBT are resistant.

0

0,5

1

1,5

2

2,5

3

0 2 4 6 8 10 12 14 16time (week)

µmol

/L

18

TCMBTMBTBTsum

N

SSCH2SCN

SHS

N

HS

N

Figure 4.33: Anaerobic biotransformation of TCMTB.

87

Table 4.8: Transformation of benzothiazoles under anaerobic treatment.

benzothiazole Spiked (0 week) µmol/L

Anaer.treatment (16 weeks) µmol/L

Metabolites µmol/L

loss (%)

OHBT 3.3 3.3 - 0

MTBT 2.53 1.73 - 32

MBT 2.99 0 1.62 (BT) 46

BTSA 2.79 0.72 0.9 (OHBT) +

unknown products

58

TCMTB 2.52 0 1.75 max (MBT) → 0

1.34 (BT)

47

The following scheme can be drawn to summarise the results of anaerobic experiments (Figure

4.34).

3

N

SSO3H

N

SSCH2SCNS

SCH

N

MTBT BTSA TCMTB

SHS

N

S

NOH H

S

N

partial degradation unknown products

OHBT MBT

no degradation no degradation

Figure 4.34: Observed anaerobic transformations of benzothiazoles.

88

Because the anaerobic treatment of benzothiazoles is poorly known, we have chosen to

compare our results to those of literature. There is a striking parallel in compositional trends

observed between our experiments and some of other authors. For instance, Reemstma et al.

(1995) showed that during anaerobic treatment of tannery wastewater, the total benzothiazole

concentration increased, primarily due to a significantly increasing concentration of BT. This

result is similar with that found in our study. The product BT has been formed from unknown

precursors in the anaerobic treatment according to Reemstma et al. (1995), whereas, the origin

of BT in our case comes from MBT transformation. Reemtsma et al. found also that MTBT

concentration was significantly diminished and MBT appeared to be refractory against

anaerobic treatment.

The same authors (Reemtsma et al.) reported also from Fiehn (unpublished data) that

analogously, during incubation under methanogenic conditions, the formation of BT and MBT

from TCMTB was observed. They considered also, that the adsorption onto biomass appeared

not to be a major elimination process for MTBT, BT, or MBT.

De Wever et al. (1998) revealed in anaerobic assays that BT, and OHBT concentrations

remained constant throughout the test period, whereas BTSA was transformed into OHBT in

nearly equimolar amounts. OHBT therefore was an intermediate in BTSA degradation as well,

and in this case the hydroxyl group probably originates from water (reported from the same

authors). Moreover, it seems that the introduction of the hydroxyl function occurs

simultaneously with the elimination of the sulfonate group. Otherwise, BT, rather than OHBT,

would have accumulated during anaerobic incubation on BTSA.

In this system, the mechanisms that are responsible for the transformation of benzothiazoles

are still not clear. Anaerobic digestion of complex wastes is an interrelated multistage

microbial process of serial and parallel reactions (Gujer and Zehnder, 1983; Zeikus, 1979). The

most probable reaction scheme for anaerobic digestion of organic waste (Pavlostathis and

Giraaldo-Gomez, 1991; Pavlostathis and Gosset, 1986) goes through three stages: hydrolysis,

fermentation and methane production.

In contrast to the strategies discussed above that are used by aerobic bacteria to degrade

benzothiazoles, anaerobic bacteria initiate degradation of benzothiazoles and heterocyclic

89

compound by reductively removing sulphonate or thio groups from the heteroatom part. In this

transformation, sulphonate or thio-atoms are replaced with hydrogen atoms.

It is also reported that heterocyclic and aromatic compounds are attacked by bacteria that

obtained energy by anaerobic respiration in which nitrate, sulphates, or ferric iron (Evans,

1977; Lovley et al., 1989; Lovley et al., 1990) serve as electron sinks. This mode of

metabolism results in conversion of a large part of substrate carbon into CO2.

90

4.4 Photolytic degradation of some benzothiazoles

Because very little information is available in the open literature on the photodegradation of

benzothiazoles in aquatic environment, we wished to report our studies not for all

benzothiazoles discussed above but only some of them like MBT, MTBT and OHBT. In this

part, we describe our research on the photolysis of some benzothiazoles. All irradiated samples

were followed by spectrophotometry and HPLC-analysis.

4.4.1 MTBT photolysis in pure water

The absorption spectra observed after Xe-lamp photolysis of MTBT in water with 313 nm (λ

exc) and a photometric energy (E = 2.70 mW/cm2) exhibit two absorption bands at 223 nm and

280 nm, as shown in Figure 4.35. Figure 4.35 shows the time dependent transient absorption

spectra recorded in water after phototolysis (313 nm excitation light, 2.70 mW/cm2) of MTBT.

The broad transient absorption band in the region of 250-310 nm decays continuously (Fig

4.35).

At an initial concentration of 18.8 µmol/L (3.4 mg/L), MTBT proved to be not completely

stable but is slowly transformed to the two photolysis products BT and OHBT under UV-

radiation (Fig 4.36). The concentration of MTBT had declined by less than 60 % (10.3 µmol/L)

in 60 min. At this time of photolysis (60 min), 2.52 µmol/L (13.4 %) of BT and 1.85 µmol/L

(9.8 %) OHBT were detected by HPLC analysis (Fig. 4.36). BT is predominant. This

concentration is lower than that observed in the case of MBT after 1 min as shown after (see

below).

It is also remarkable that the sum of all constituent present in the aqueous solution was nearly

constant. This could be explained by the resistance of the photolytic products (BT and OHBT)

and by the absence of other photolytic products. Whereas, between the beginning and the end

of photolysis a loss of 3.3 µmol/L of MTBT (17.5 %) was observed. This loss could be due to

the fact that MTBT is considered as the second volatile compound. No photolysis product was

observed before 10 min.

With another photonic energy (E = 0.72 mW/cm2), and by using the same concentration (3.4

mg/L) or another higher concentrations (for example 8.2 mg/L), the photolysis degradation (λ

exc = 314.8 nm) was very slow or practically no transformation of MTBT was observed

compared to the results obtained with 3.4 mg/L. Figures 4.37 shows no decrease of UV-

91

absorption at the two maximum absorption of MTBT (223 and 280 nm) under photolytic

conditions.

0

0,05

0,1

0,15

0,2

0,25

0,3

200 220 240 260 280 300 320 340wavelength (nm)

UV-

abso

rptio

n0 min3 min10 min30 min60 min

Figure 4.35: UV-transformations of MTBT under photolysis ( from 0 to 60 min, downward)

(3.4 mg/L, λ exc = 313 nm, and E = 2.70 mW/cm2).

02468

101214161820

0 10 20 30 40 50 60 7t im e (m in)

µmol

/L

0

MTBTBTOHBTsum

Figure 4.36: HPLC-results of MTBT-photolysis (3.4 mg/L, λ exc = 313 nm, and E = 2.70

mW/cm2).

92

0

0,1

0,2

0,3

0,4

0,5

0,6

0,7

0 20 40 60 80 100 120 140 160 180 200

time (min)

Abs

orpt

ion

280 nm

223 nm

Figure 4.37: UV-absorption of MTBT at 223 and 280 nm during photolysis (8.2 mg/L, E =

0.72 mW/cm2, λ exc = 314.8).

4.4.2 MBT photolysis in pure water

In the case of MBT, spectra quite different from those of MTBT were observed. The

absorption spectra observed during Xe-lamp photolysis of MBT in water with 313 nm and E

(2.70 mW/cm2) exhibits three absorption bands at 206, 228 and 316 nm, as shown in Figure

4.38.

The absorption peak at 316 nm in water was attributed to the thiol form (MBT), (Burkey and

Griller, 1985; Bordwell and Hughess, 1982; Mascijewski and Steer,1993), which shifted to a

longer wavelength with the appearance of another broad absorption band in the region of 400-

500 nm in polar solvent and in aqueous solution (not shown). The broad absorption band in the

longer wavelength region is characteristic of a C=S chromophore (Bordwell and Hughess,

1982; Steer and Ramamurthy, 1988; Maciejewski and Steer, 1993; Alam et al., 1998)

indicating that the MBT resonance structure makes a larger contribution than does MBTS

(Bordwell and Hughes, 1982). In the absorption spectra of the disulfide (MBTS), such an

appearance of new absorption in the longer wavelength was not observed (not shown).

93

0

0,1

0,2

0,3

0,4

0,5

200 220 240 260 280 300 320 340

wavelength (nm)

UV-

abso

rptio

n (1

/cm

)0: 0 min 3: 3 min 1: 1min 2: 2 min4: 4 min 5: 5 min 6: 6 min 7: 7 min8: 8 min 10: 10 min 11: 29 min

1

3

2

4

6

78

10

11

5

Figure 4.38: UV-transformations of MBT under photolysis (2.97 mg/L, λ exc = 313 nm, and E =

2.70 mW/cm2; from 0 to 29 min, downward).

The importance of thione photochemistry from chemical and biological viewpoints has been

recognized for various thiol-thione tautomers (Steer and Ramamurthy, 1988; Maciejewski and

Steer, 1993; Alam et al. 1998).

At an initial concentration of 17.8 µmol/L (2.97 mg/L, λ exc = 313 nm, and E =2.70 mW/cm2),

MBT proved to be easier to photolyse than MTBT and was completely transformed. The decay

of the absorption bands of MBT increases with irradiation time. To confirm and to quantify

the formation of metabolites, HPLC was used.

Within 1 min, MBT decayed by 54 % (9.6 µmol/L) and was transformed to 0.67 mg/L (24 %)

OHBT and 0.67 mg/L (28 %) of BT (Fig. 4.39). After 29 min, only 2.6 µmol/L (14.6 %) of

MBT remained in the cell (a loss of 15.2 µmol/L between the start and the and of photolysis).

It is also noted that no other photoproducts were found in this case with exception of OHBT

and BT. At 6 min the concentration of MBT and the concentration of BT as photoproduct were

about the same (6.9 µmol/L and 6.7 µmol/L respectively) as well as for the concentration of

MBT and OHBT at 8 min (5 and 4.8 µmol/L respectively). The total of benzothiazoles (sum)

94

in the cell between 8 and 29 min was constant indicating that all the photolysis products of

MBT were recorded.

0

5

10

15

20

25

0 5 10 15 20 25 30 35time(min)

µmol

/L

MBTOHBTBTsum

Figure 4.39: HPLC-results of MBT-photolysis (2.97 mg/L, λ exc = 313 nm, and E =2.70

mW/cm2; from 0 to 29 min, downward).

Another photonic energy (E = 0.62 mW/cm2) and another concentration of MBT (~ double of

2.97 mg/L) were used in order to know the effect of energy on the photolysis. While the

photolytic degradation was very fast with 2.70 mW/cm2, the degradation was slower in the

case of a smaller energy (0.62 mW/cm2) compared to the first photolytic test (Figure 4.40).

This is clearly shown in this Figure (Fig. 4.40) which represents the starting and the ending of

photolysis (λ exc = 314.8 nm). At 180 min MBT was still not transformed compared to the first

photolysis where MBT was almost completely photolysed in 29 min.

95

200,0 220 240 260 280 300 320 340 360 380 400,0-0,01

0,1

0,2

0,3

0,4

0,5

0,6

0,69

NM

A

1

4

2

3

6

5

5

Figure 4.40: MBT photolysis in pure water (5.52 mg/L; λ exc = 314.8 nm, 0.62 mW/cm2)

1: 0 min, 2: 30 min , 3: 60 min , 4: 90 min, 5: 120 min, 6: 180min.

In this case, the transformation began in fact after 30 min (Fig. 4.41). It is remembered that at

this time (30-60 min), MBT was completely photolysed in the first case. From an initial

concentration of 33 µmol/L MBT (5.52 mg/L), 8.7 µmol/L remained in the cell at the end of

the photolysis, corresponding to 40 % of the initial concentration. The same metabolites

(OHBT and BT) were formed in this second experiment. At the end of photolysis, 18.5 µmol/L

BT and 5.2 µmol/L OHBT were formed (Figure 4.41). Traces of MBTS were indicated by

HPLC. No loss of MBT was observed during the photolysis test regarding the constant sum

(µmol/L) of benzothiazoles found (Figure 4.41).

96

0

5

10

15

20

25

30

35

0 50 100 150 200

time (min)

µmol

/L

MBT BT OHBT sum

Figure 4.41: HPLC-results of MBT photolysis (5.52 mg/L, λ exc = 314.8 nm, 0.62 mW/cm2 ).

A comparison between a solution of MBT photolysed and another one, which was not

subjected to photolysis (same concentration and volume) has been done.

Figure 4.42 show the behavior of a solution of 5.52 mg/L MBT not irradiated. A slight

diminution of UV-absorption was observed at 5 hours (Fig. 4.42) (a loss of 0.2 µmol/L). In 12

hours, a diminution of 0.92 µmol/L (17 %) was mentioned whereas, transformation of 74 % of

MBT was observed under photolytic conditions in three hours (Fig. 4.41). As consequence,

transformation observed in Fig 4.41 are due to photolysis and do not occur in the absence of

light. Figures 4.43 and 4.44 show a linear decrease in UV-absorption and in concentration

between 40 and 180 minutes.

97

200,0 220 240 260 280 300 320 340 360 380 400,00,00

0,2

0,4

0,6

0,8

1,00

NM

A

2 13

4

5

5

Figure 4.42: MBT solution in absence of photolysis (5.52 mg/L) 1: 0 min, 2: 180 min, 3: 5 hours, 4: 7 hours, 5: 12 hours.

0

1

2

3

4

5

6

0 50 100 150 200 250 300 350time (min)

mg/

L

photolysis (320 nm)no photolysis (320 nm)

Figure 4.43: Comparison of MBT behaviour at 320 nm under and in absence of photolytic

conditions (C: 5.52 mg/L, λ exc = 314.8 nm, 0.62 mW/cm2).

98

0

0,1

0,2

0,3

0,4

0,5

0,6

0,7

0,8

0 50 100 150 200 250 300 350time (min)

UV-

abso

rptio

nphotolysis (320 nm)no photolysis (320 nm)

Figure 4.44: UV-Comparison of MBT at 320 nm under and in absence photolysis (λ exc = 314.8 nm, λ exc = 314.8 nm, 0.62 mW/cm2).

Tests solution of BT and OHBT were also irradiated, but no degradation of these compounds

was observed in aqueous solution. All bands exhibit no absorption changes over several hours,

indicating the stability of these compounds.

Although a photochemical investigation of benzothiazoles has been reported in non polar

solvent, a detailed study in polar solvents and in aqueous solutions has not been reported.

Some laboratory studies have been reported on the photolysis of MBT in organic solvents

using UV lamps (Párkányi and Abdelhamid, 1985; Abdou et al., 1987).

Párkányi and Abdelhamid (1985) found that UV irradiation of 2-mercaptobenzothiazole

(MBT) in ethanol, methanol or acetonitrile resulted in the isolation of the benzothiazoles

analogs c and e as shown in Figure 4.45. The disulfides (b) and disulfone (c) are intermediates

and must be precursors of the major products (e) (Fig. 4.45). Both oxygen and water have been

found to be essential for completion of the photoreactions.

99

N

SSH

��hv

EtOH

N

S2

N

S SO3H.+

�� hv

H2O

N

SH

MBT

a b

2

N

S SO2

hv ��

S

��hv H2O

c

de

O2

BTSA

H2SO4

Figure 4.45: UV-irradiation of MBT in ethanol (Párkányi and Abdelhamid,1985)

Párkányi and Abdelhamid, considered in their communication (1985) that sulfate is the main

product beside the disulfone. It seems to us and based on our experimental results unlikely for

the following reasons:

• it is known that BTSA (e) (Fig. 4.45) is easy to decompose (Cocker, 1936) in the

presence of water.

• Párkányi and Abdelhamid (1985) attributed the sum formula C7H5NS2O3 to BTSA (e)

are based upon an elemental formula, however, the correct elemental composition has

been states in 1936 to be C7H5NS.H2SO4, since it has been reported that attempts to

dehydrate this hydrate led to its decomposition (Cocker, 1936).

• the disulfone (c) has been reported to be extremely insoluble (Párkányi and

Abdelhamid, 1985); upon our photolysis experiment, we were unable to detect any

significant amount of this compound (no BTSA was found).

It is also interesting to note that the results of Crank and Mursyidi (1982) which found that the

irradiation of MBT in the presence of air led to the thion form, sulfur and hydrogen sulfide,

were contradictory compared to Párkányi and Abdelhamid (1985). Whereas, Abdou at al

(1987) found that the irradiation of MBT in acetonitrile, ethanol or benzene (λ exc =313 nm)

produced a pattern of products displayed in Figure 4.46. The main products obtained in their

100

study (Abdou et al.,1987) were 5.6% benzothiazole (i), 9.7 % elemental sulfur, 20.8 %

2(3)benzothiazolethione (a thione form), 32 %bis-(-benzothiazolyl)-disulfide (b), and 20.5 %

2-3(3H)-benzothiazolone (OHBT) (h). A mechanism that accounts for the formation of the

obtained photoproducts by Abdou et al. (1987) is presented in Fig. 4.47. Thus, upon UV

irradiation, MBT cleaves at the sulfur-hydrogen bond. Subsequent recombination of the thiyl

radicals formed leads to the disulfide (b) (Dürr, 1975; Knight, 1974). These intermediates have

been proposed and identified not only by Abdou et al. (1987) but also by Párkányi and

Abdelhamid (1985).

a

MBT

N

SSH

�� N

SS

�� N

SS S

S

N

b

N

SS.

f

+

S

N

SS.

+

N

S.��

N

SO

��hv

��

hv

O2 hv

g 2 g 1

2

��

��

N

SSH

MBT

a��

S (H)2-

N

SH

(H)

(O)

hv

CH3CN

i

BT

H

disulfide

H

N

SOH

��

OHBT

MBTS

Figure 4. 46: UV-irradiation of MBT in ethanol (Abdou at al., 1987).

101

Upon UV irradiations of the disulfide (b) (Fig 4.46), besides the retro S-S-cleavage into to 2

thiyl radicals (f) (Dürr, 1975; Knight, 1974 in Patai (ed.); Kice, 1973 in Kochi (ed.); Haines et

al., 1956), another homolysis can be discussed leading to 2-benzothizolyl radical (g1) and the

corresponding disulfan radical (g2). This type of competitive carbon-sulfur cleavage has been

discussed earlier (Coyle, 1986) for the photoreactions of disulfide. Radical g1 transforms to

benzothiazole BT (i) via hydrogen abstraction, while, by oxygen attack on g1, 2(3H)-

benzothiazolone (h) is formed (OHBT) (Crank, 1982). On the other hand, the disulfane radical

(g2) can break down to elemental sulfur and MBT after H-transfer.

The rapid formation of disulfide (b) might act as quenchers for singulett oxygen (Foote, 1979;

Foote and Peters, 1971). A direct transformation of MBT, with residual traces of singulett

oxygen cannot be excluded, since there are many C=S, C=O exchange reactions reported for

thiones after treatment with singulett oxygen (Wamhoff and Ertas, 1882; Ishibe et al. 1971;

Suzuki et al., 1981).

The transformation of MBT to OHBT can also occur as a direct photochemically induced

reactions with singulett oxygen.

In our experiment, the irradiation of MBT in water (pH 7) led to BT and OHBT. This result

correlates with those obtained by Brownlee et al. (1992). The production of BT and OHBT is

consistent with the mechanism of Abdou et al. (see above Figure 4.46). As a consequence, we

assume a different reaction course leading primarily to disulfide. Regarding the work of

Párkányi and Abdelhamid (1985), we can perhaps assume that upon photolysis of MBT, the

disulfide is in fact an intermediate as well as main photoproduct but, on the other hand, we

were unable to confirm in our study the presence of the sulfone (c) or the hydrosulfate (e)

found by Párkányi and Abdelhamid (1985) (Fig. 4.45).

As thiols are very reactive and provide many oxydation states (Kice, 1980) other pathways

may be possible for the formation of OHBT and BT from MBT.

The environmentally important conclusion to be drawn from our results is that BT and OHBT

are the anticipated stable products of MBT photolysis in aquatic environments.

In addition to information about photodegradation products, some estimations of the rate of

photodegradation in natural systems is needed to predict the photochemical fate of

contaminants.

102

The Quantum yield (Φ) were calculated from the equations given in chapter 2.2.4.1 and were

summarised in Table 4.9. The estimated quantum yield for MBT in wastewater was

significantly higher than that in water. Quantum yield of intersystem crossing (Φ) was

evaluated to be in the range 0.023-0.112 in water and 0.839-0.987 in wastewater.

Table 4. 9: Quantum yield (Φ) of benzothiazoles in water and wastewater.

Test compound

Initial concentration (mg / L)

medium E= (mW/cm2) Quantum yield (Φ) /λ

MBT 2.97 pure water 2.70 0.112 (/320 nm)

MBT 5.52 pure water 0.62 0.023 (/320 nm)

MBT 5.52 wastewater 0.62 0.987 (/320 nm)

MTBT 3.4 pure water 2.70 0.063 (/280 nm)

MTBT 8.2 pure water 0.72 -

MTBT 8.2 wastewater 0.72 0.839 (/280 nm)

OHBT 11.2 pure water 0.65 -

OHBT 11.2 wastewater 0.65 -

4.4.3. Application in wastewater

Similar conditions were used as in abiotic photolysis. The (wastewater) effluent was filtered

and neutralized. The main objective of this experiment in wastewater with was to show the

influence of DOM on the photodegradability of benzothiazoles.

In fact, a compound may not undergo photochemical transformation at all in distilled water,

but may react at a significant rate in water containing other constituents. The reason for this

behavior are either an energy transfer from another excited species or (chemical) reactions of

the non-excited compounds with very reactive species (e.g., hydroxyl radical, peroxy-radicals,

singlet oxygen) that formed in the presence of light (Schwarzenbach et al., 1993) in indirect

photolysis

4.4.3.1 MTBT photolysis in wastewater (biotic)

At an initial concentration of 3.4 mg/L (18.8 µmol/L) in effluent, MTBT had a similar behavior

as that tested in pure water, but the concentration decreased slightly compared to the reference.

103

At the end of photolysis only 13.2 µmol/L was detected. That means ca. 30% of MTBT were

transformed, giving 5.2 µmol/L BT as photoproduct. In this case no OHBT was detected. It is

also interesting to remark that the concentration of BT in case of photolysis in wastewater

(effluent), the production of BT was 2 times higher as that produced in pure water.

The sum of benzothiazoles (MTBT and BT) present in the cell was constant during the

photolytic irradiation (Figure 4.47). without irradiation, MTBT was stable also in the effluent

matrix (Fig. 4.47).

0

2

4

6

8

10

12

14

16

18

20

0 20 40 60 80 100 120 140 160 180 200

time (min)

µmol

/L

MTBT (photolysis)MTBT(no photolysis)BT (photoproduct)

Figure 4. 47: MTBT behaviour in wastewater (effluent) under and in the absence of

photolysis (8.2 mg/L, E =0.72 mW/cm2, λ exc = 314.8).

4.3.2 MBT photolysis in wastewater

In spite of filtration and neutralization of wastewater, the spectrum of MBT was not very clear

and the two important absorption maxima (320 and 220 nm) were masked (Figure 4.48) by the

wastewater matrix.

Without photolysis, the spectrum of MBT indicated no reaction took place. This was confirmed

by HPLC-analysis (Figure 4.48).

104

Photolysis of MBT occurred very rapidly. Also in this matrix, the absorbance characteristic

band of MBT at 320 nm was significantly reduced within 20 min (Figure 4.49). The second

absorption band (220 nm) was also reduced and removed to lower wavelengths.

0 min

180 min

Figure 4. 48: MBT behaviour in wastewater without photolysis (C 5.2 mg/L).

200,0 250 300 350 400,0

-0,45

-0,2

0,0

0,2

0,4

0,61

NM

A

1

2

3

Figure 4.49: MBT photolysis in wastewater (5.52 mg/L; λ exc = 314.8 nm, 0.62 mW/cm2).

1: 0 min, 2: 10 min , 3: 20 min.

105

0

5

10

15

20

25

30

35

0 20 40 60 80 100 120 140time (min)

µmol

/LMBT BTOHBTsum

Figure 4. 50: MBT in wastewater (effluent) under photolysis (5.52 mg/L; λ exc = 314.8 nm,

0.62 mW/cm2).

Figure 4.50 which displays the behavior of MBT in wastewater (effluent) reveals the

predominance of BT as photoproduct. This correlates with the results obtained in pure water.

Not only BT and OHBT were detected as photoproducts but also traces of MBTS were found.

Yet after 1 min, only a concentration of 26.7 µmol/L of MBT was present in the photolytic

cell. In 10 minutes, a decrease of 79 % was observed (only 7.1 µmol/L was present; Figure

4.50). After 20 min, no MBT was detectable. The concentration of BT and OHBT at the end

of irradiation were 16.6 and 4.4 µmol/L respectively. These concentrations are comparable

with those obtained by the irradiation in pure water using the same concentration conditions.

It is also noteworthy (Figure 4.50) that the total concentration of benzothiazoles found in 20

min was only 5.9 µmol/L. At the end of the experiment (120 min), 36 % (a total of 21µmol/L)

were found). The loss, particularly at 10 min, cannot be considered as loss, but is probably due

to the mechanism of tautomery of MBT. The MBT could be transformed in part to unknown

intermediates, for example MBTS (poorly soluble). Figure 4.51 presents the comparison of

MBT under and in absence of photolysis.

106

0

0,5

1

1,5

2

2,5

3

3,5

4

4,5

5

0 20 40 60 80 100 120 140 160 180time (min)

mg/

Lphotolysis

no photolysis

Figure 4. 51: Comparison of MBT with and without photolysis.

4.3.3.3 OHBT photolysis in effluent

Contrary to MBT and MTBT, OHBT proved to be stable during photolysis in wastewater

(Figure 4. 53).

0,44

0,45

0 50 100 150 200 250 300 350

time (min)

UV-

abso

rptio

n

photolysis 220 nm)no photolysis (220 nm)

Figure 4. 52: OHBT behaviour in wastewater under photolytic conditions (11.2 mg/L; 0.65

mW/cm2).

107

It is concluded from this part of study of MBT by photolysis that in both cases, water and

wastewater, two stable photoproducts were formed: BT and OHBT, supporting previous work

done by Brownlee et al. (1992) which was the only author which studied MBT and MTBT in

aqueous solutions.

Concerning MTBT, Brownlee et al. found that this compound was stable and no change was

observed under solar radiation whereas we observed a reduction of its UV-absorption and the

production of the main photoproduct BT.

The probable reaction mechanism for MBT observed in our study could be the same as

proposed by Abdou et al. (1987) in ethanol.

108

5. CONCLUSIONS 5.1 Extraction

The recovery rates of benzothiazoles (2.3-2.7 µg/L) depended on the pH-value, particularly

and strongly in the case of MBT, which was not extractable with any phase and at any pH-

value except with neutral-pH, where its extraction reaches a recovery of ca. 64 %. While

BTSA, OHBT, ABT and MTBT were recovered at all pH-values to >50%, BT was detected at

pH 6 < 40%.

The polar benzothiazole compounds such as BTSA and OHBT are well recovered from the

polar polymer phases, namely from the EN-phase. The other analytes showed good recoveries

on the silicagel-cartridges. Based on their high capacity, RP 18 and EN were chosen, for a

sequential solid phases-extraction of benzothiazole compounds, with a sequential elution,

methanol followed with acetone.

The recovery rates of the SPE-extraction from STP effluent and STP influent are slightly

decreased compared to those obtained with tap water.

5.2 Aerobic degradation

Aerobic mixed cultures are capable of utilising benzothiazoles as a sole source of carbon,

nitrogen and energy. BT, OHBT, MBT and MTBT were investigated in separate batch tests.

BT and OHBT can be microbially degraded in water in a relatively short time, whereas MTBT

proved to be stable under aerobic conditions (Fig. 5.1). MBT was methylated to MTBT. After

3 days a diminution to 57 % of employed MBT was observed. The controls lacking an

inoculum failed to show any decrease in all studied benzothiazoles, confirming that the

degradation in the experimental cultures was of biological but not of chemical origin.

The different results concerning MBT between our study and that of De Wever and Verachtert

(1994) may be due to different populations of bacteria that reside in the experiment or other

factors (e.g. salts, oxygen amount etc..). However, our results approached those of Reemtsma

et al. (1995).

Results from the Sequencing Batch Reactor (SBR) experiment indicated that losses of MBT

were considerable. It should be noted that the main mechanism of MBT disappearance is

probably the biological transformation. Volatilisation appears to be insignificant, but

adsorption on biomass appeared to be important. The detection of metabolites as intermediate

109

products (MTBT, OHBT, BT, and BTSA (Fig. 5.2)), throughout the SBR study suggested that

MBT biotransformation may have occurred.

N

SSH

N

SOH

N

SSCH3

N

SH

��

��

��

��

complt. degradation

MBT OHBT BT

aer.aer. aer.

MTBT

no degradation

(NH4+ SO4

--)+ ?

aer.

Figure 5.1: Observed aerobic transformations

SHS

NS

S

NS

S

N��SO3H

S

N

OHS

N

SCH3S

NH

S

N

mineralisation

��

��

partal mineralisation

BTSA MBT MBTS

BTOHBT

sorption

��

��

�� MTBT

��

��

��

��

��

sorption ?

Figure 5.2: Observed SBR transformations

110

5.3 Anaerobic degradation

Anaerobic treatment with initial concentrations ranging from 0.1 to 0.5 mg/L revealed no

transformation of OHBT, whereas MTBT was partly degraded. After 6 weeks only 68 % of

MTBT were found. MBT was partly metabolised to BT. BTSA was partially transformed to

OHBT, which was observable only after 5 weeks together with other unknown substances that

were not identified.

Concerning TCMTB, BT is beside MBT an important metabolite (Fig. 5.3).

N

SSCH3

N

SSO3H

N

SOH

N

SSCH2SCN

N

SH

N

SSH

��

��

��

��

MTBT BTSA TCMTB

BTMBTOHBT

unknownproducts

��

no degradation��

partialdegradation

Figure 5.3: Observed anaerobic transformations

5.4 Photolytic transformations

Under photolytic conditions in water and wastewater, both MBT and MTBT revealed the

formation of two stable photoproducts: BT and OHBT (Fig. 5.4).

MTBT proved to be not stable but is slowly transformed to BT and OHBT under UV-radiation

(λ exc = 314 nm). The concentration of MTBT had declined by less than 60 % in 60 min. MBT

is easier photolysed than MTBT and was completely transformed (Fig. 5.4). This result does

not correlate with that obtained by Brownlee et al. (1992), who found that MTBT was stable

and no change was observed under solar radiations.

111

Under UV-radiation (λ exc = 313 nm), MBT at a concentration of 2.97 mg/L showed a decay of

ca. 54 % after 1 min and was transformed to 24 % OHBT and 28 % BT. This behaviour of

MBT under photolytic conditions supports previous work done by Brownlee et al. (1992).

The estimated quantum yield for MBT and MTBT in wastewater was higher than that in water.

The photolysis of MBT in wastewater effluent occurred very readily and revealed a correlation

with pure water by the predominance of BT as photoproduct. Traces of MBTS have been also

detected.

The production of BT and OHBT correlates with the mechanism of Abdou et al (1987).

consequently, we assume a different reaction course leading primarly to dissulfide.

Regarding the work cited by Párkányi and Abdelhamid (1985) we can perhaps assume that

upon photolysis of MBT, the dissulfite is in fact an intermediate as well as a main

photoproduct.

MTBTN

SSCH3

N

SSH

MBT

N

SH

OHBT BT

+N

SOH

��

N

SS

S

NS

MBTS

(traces)

��

��

no degradation

hv hv

hv

��

Figure 5.4: Observed photolytic transformations

112

7. REFERENCES Abdou W.M., Sidky M.M. and Wamhoff H., Photochemistry of pesticides, (12) on the photoconversion of 1,3-Dihydro-2H-benzimidazole-2-thione, 2(3)-Benzothiazolethione, and 2-chlorobenzothiazole. Z. Naturforsch. 42b, pp. 1153-1158, 1987. Alan R. K., Wojciech Kuzmierkiewicz and Aurrecoechea Jose M. Reactions of (benzothiazol-2-ylthio)(trimethylsilyl)methane. A general method for alpha-mercaptoalkylation by alkylation and alkylative desilylation. J. Org. Chem. Vol. 52, Iss. 5, pp. 844-849, 1987. Alam M. Maksudul, Watanabe A., Ito O. Laser Flash photolysis of dithio-2,2´-dipyridine; structure and reactivity of pyridyl-2-thio radical, J. Org. Chem. Vol. 60, Iss. 11, pp. 3440-3444, 1995. Alam M. M., Watanabe A., Ito O. Laser flash photolysis of 2,2 dithiobis(pyrine noxide): Reactivity of V-oxypyridyl-2-thio radical. Photochem. Photobiol. Vol. 53, Iss.1, January, 1996. Alam M. M., Kanomi H., Watanabe A., Ito O. Laser flash photolysis study of dithiobis(tetrazol); reactivities of tetrazole-thioradical. Int. J. Chem. Kinet., Vol. 28, Iss. 6, pp. 405-411, 1996. Alam M. M., Fujitsuka M., Watanabe A., Ito O. Photochemical properties of excited triplet state of 6-H-purine-6-thione investigated by laser flash photolysis. J. Phys. Chem. A, Vol. 102, Iss. 8, pp. 1338-1344, 1998. Amato J. R., Mount D. I., Durhan E.J., Ankley G.T., Robert E. D. and Lukasewycz M.T. An example of the identification of dizinon as a primary toxicant in an effluent. Environ. Toxicol. Chem. Vol. 11, pp. 209-216, 1992. Ambro M., Otepka L. Antimicrobial treatment of protective clothing. Czech CS 210,066. Cited in Chemical Abstracts CA 98: 73795w, 1982. Anonym. Mercaptobenzothiazole; proposed test rule register 50, 46121-46133, 1985. Aislabie J. and R.M. Atlas. Microbial removal of heterocyclic nitrogen compounds from shale oil. In Proceedings on Bioprocessing of Fossil Fuels Workshop. Virginia, USA, 1989. Aveline B.M., Kochevar I.E., Redmond R.M. Photochemistry of N-hydroxypyridine-2-thione-derivatives: Involvement of 2-pyridylthiyl radical in the radical chain reaction mechanism. J. Am. Chem. Soc. Vol. 117, Iss. 38, pp. 9699-9708, 1995. Aveline B. M., Kochevar I.E. Redmond R.W. Photochemistry of the non specific hydroxylradical generator, N-hydroxypyridine-2(H)-thione. J. Am. Chem. Soc. Vol. 118, Iss. 42, pp. 10113-10123, 1996. Barker J. F., Barbash J.E., and Labonte M.J. Groundwater contamination at landfill sited on fractured carbonate and shale. Journal of Contam. Hydrology 3, pp. 1-25, 1988. Beiko O. A., Gerasimova N. N. and Sagachenko T. A.. Nitrogen bases of industrial west Siberian crude oil. Petrol. Chem. USSR 27(3), pp. 200-210, 1987. Bennett J.L., Updegraff D.M., Pereira W.E. and Rostad C. E. Isolation of four species of quinoline-degrading Pseudomonads from cresol-contaninated site at Pensacola, Florida. Microbios Lett. 29, pp. 147-154, 1985. Bester K, Hühnerfuss H. triazines in the baltic and North Sea. Mar. pollt. Bull. 26, pp. 423-427, 1993a. Bester K, Hühnerfuss H. triazines in the baltic and North Sea. Mar. pollt. Bull. 26, pp. 657-658, 1993b. Bester K., Hühnerfuss H., Brockmann U., Rick H. J. Biological effects of triazine herbicide contamination on marine phytoplankton. Arch. Environ. Contam. Toxicol. 29, pp. 277-283, 1995.

113

Bohne C., Boch R. and Scaiano J.C. Exploratory studies of chemistry of N-hydroxypyridine-2-thione esters. Generation of excited radicals by laser photolysis and in a conventional fluorescence spectrometer.J. Org. Chem. Vol. 55, Iss. 19, pp. 5414-5418, 1990. Bordwell F.G., Hughess D.L. Thiol acidities and thilate ion reactivities toward butyl chloride in dimethyl sulfoxide solution. The question of curvative in Broensted plots. J. Org. Chem. Vol. 47, Iss. 17, pp. 3224-3232, 1982. Braunstein L., Hochmüller K., Sengler K. Analyse Gaschromatographierbarer Wasserinhaltsstoffe durch Kopplung von Gaschromatographie und Massenselektivem Detektor. Vom Wasser 73, pp. 167-189, 1989. Brownlee B., Carey J.H., Fox M.E. Areview of benzothiazoles in the aquatic environment. National Water Research Institute Scientific Series No. 126, Ontario: Environmental Canada, I. W.D., 1981. Brownlee B.G., Carey J.H., MacInnis G.A. and Pellizzari I. T. Aquatic environmental chemistry of 2-(Thiocyanomethyl)benzothiazole and related benzothiazoles. Environ. Toxicol. Chem. 11, pp. 1153-1168, 1992. Bugby A., Parbery C., Sykes R.L. and Taylor C.D., J. Soc. Leather Technol. Chem. 74, pp. 134, 1990. Bujdáková H., Kuchta T., Sidóová E. and Gvozdjaková. Ati-candia activity of four antifungal benzothiazoles. Microbiology Letters 112, pp. 329-334, 1993. Burkey T. J., Griller D. J. Micelar systems as devices for enhancing the lifetime anf concentrations of free radicals. Am. Chem, Soc. Vol. 107, Iss. 1, pp. 246-249, 1985. Burkhard L.P., Durhan E.J. and Lukasewycz M.T. Identification of non-polar toxicants in effluents using toxicity-based fractionation with gas chromatography/massspectrometry. Anal. Chem. 63, pp. 277-283, 1991. Carey J.H., Fox M. E., Brownlee B. G., Metcalfe J. L. Mason P. D. Yrex W.H.. The fate and effects of contaminants in Canagagigues Creek. I. Stream Ecology and Identification of major contaminants; Scientific Serie 135 I.W.D.Ontario, Canada, 1983. Catallo W.J., Cleland D.R. and Bender M.E. Toxicity of nitrogen-containing aromatic compounds (NCACs). Quinoline and azafluorene behavior in an Escherichia coli test system: evidence of membrane effects. In Landis W.G. and van der Schalie W. H. (eds.). Aquatic Toxicology and Risk Assessment ASTM Special Technical Publication 1096, American Society of Testing Material, Philadelphia, PA, pp. 199-221, 1990. Chaudhry G.R. Biological degradation and bioremediation of toxic chemicals, 1994 Ching-Erh Lin, Chia-Chong Chen. Migration behavior and separation of aromatic triazole and thiazole compounds by cappilary zone electrophoresis. Journal of chromatographya 731, pp. 299-303, 1996. Chowdhury F.A., Cole E.R., and Crank G. Chromatography of benzothiazoles. Journal of Chromatography, 170, pp. 73-80, 1979. Christopher M. Reddy and Quinn James G. Environmental chemistry of benzothiazoles derived from rubber. Environ. Sci. Technol. 31, pp. 2847-2853,1997 Chudoba J.F. Tucek, and Zeis K., Biochemischer Abbau von Benzothiazolderivaten. Acta Hydrochim. Hydrobiol. 5, pp. 495-498, 1977. Clark L. B., Rosen R.T., Hatman T.G., Alaimo L.H., Louis G.B., Herts C., Ho C.-T. and Rosen J.D. Determination of non-regulated pollutants in three New Jersey publicy owned treatment works (POTWs). Res. J. wat. Poll. Control Fed. 63, pp. 104, 1991. Cocker W. J. Soc. Chem. Ind. 55, pp.2285, C.A. 30, 8213, 1936. Coleman W.E., Melton R.G., Kopfler F.C., Barone K.A., Aurand T.A. and Jellison M. Identification of organic compounds in mutagenic extract of a surface drinking water by a

114

computerizied gas chromatography/mass spectrometry system (GC/MS/COM). J. Environ. Sci. Technol. Vol. 14, Iss. 5, pp. 576-588, 1980. Coyle J.D. Introduction of organic photochemistry. pp. 159 Wiley, Chichester, 1986 Crank G. and Mursyidi A., Aust. J. Chem. 35, pp. 775, 1982. Cserjesi A.J., Johnson E.L. Mould and sapstain control: laboratory and field tests of 44 fungicidal formulations. For Prod. J. 32(10), pp. 59-68, 1982. Czekowski M., and Rossmoor HW. The effect of selected industrial biocides on lactate metabolism in Desulfovibrio desulfuricans. Dev. Ind. Microbiol. 22, pp. 794-804, 1981. Dacre Jack C. A preminary toxicalogical evaluation of eight chemicals used as wood preservatives. Technical report 8405, March –June, 1981. Dacre J.C. Apreliminary toxicological evaluation of eight chemicals used as wood preservatives. Technical Repport 8405, NTIS, AD, A144526, 1984. Daniel R. Shelton, and Tiedje J.M. General method for determining anaerobic biodgradation potential+, Applied and Environmental Microbiology. pp. 850-857, Apr. 1984. Daniels C.R. and Swan E.P. HPLC essay of the anti-stain chemical TCMTB applied to lumber surface. J. Chromatogr. Sci. 25, 43-45, 1987. Days K.E., Holze K.E., Metcalfe-Smith J. L., Bishop C.T. and Dutka B.J. Toxicity of leachate from automobile tires to aquatic biota. Chemosphere, Vol. 27, No. 4, pp 665-675, 1993. De Vos D.H., De Wever and Verachtert H. Parameters affecting the degradation of benzothiazoles and benzimidazoles in activated sludge systems. Appl. Microbiol. Biotechnol. Vol. 39, pp. 622-626, 1993. De Wever Helene, De Moor K. and Verachtert H. Toxicity of mercaptobenzothiazole towards bacterial growth and respiration, Appl. Microbil. Biotechnol. Vol. 42, pp. 631-635, 1994. De Wever Helene, De Man I. and Verachtert H. Microbial degradation of 2-hydroxybenzothiazole. Proceding. 2nd International Symposium on Environmental Biotechnology. Bringhton, UK, July 4-6, pp. 128-130, 1994. De Wever Helene, Saskia Van Den Neste and Hubert Verachtert. Inhibitory effects of 2-Mercaptobenzothiazole on microbial growth in a variety of trophic conditions. Environmental Toxicology and Chemistry. Vol. 16. No. 5. pp. 843-848, 1997. De Wever Helene, De Cort H.S., Noot I. and Verachtert H. Isolation and characterization of Rhodococcus Rhodochrous for the degradation of the wastewater component 2-hydroxybenzothiazole. Appl. Microbiol. Biotechnol. Vol. 47, pp. 458-461, 1997. De Wever Helene, Vereecken K. Stolz A. and Verachtert H. Initial transformation in the biodegradation of benzothiazoles by Rhodococcus isolates. Applied and Environmental Microbiology, pp. 3270-3274, Sept. 1998. Dietrich A.M., Millington D.S. and Seo Y.-H. J. Specific identification of synthetic organic chemicals in river water using liquid-liquid extraction and resin adsorption and accurate mass measurement gas chromatography-mass spectroscopy analyses. J. Chromatogr. Vol. 436, pp. 229-241, 1988. DIN EN 29888. Bestimmung der aeroben biologischen Abbaubarkeit organischer Stoffe im wässrigen Medium. Statischer Test (Zahn-Wellens-Verfahren), deutsche Norm ISO, 9888, 1993. Dhindsa G.S., Sawhney S.N. Thin-layer chromatographic separation of some benzothiazole derivatives. Chim. Acta., 18(2), pp. 277-284, 1990. Drotar A.-M. and Fall R. Microbial methylation of benzothiazoles and release of methylthiobenze. Experimenta 41, pp. 762-764, 1985. Dürr Cf. H.. in Houbon-Weyl-Müller, Methoden der organischen Chemie, vol. 4, part 5b, Thieme verlag, , pp. 1008ff, Stuttgart,1975.

115

Elfarra A. A., Hwang I.Y. In vivo metabolites of S-(2-Benzothiazolyl)-L-Cysteine as markers of in vivo cysteine conjugate ß-lyase and thiol glucuronosyl transferase activities. The American Society f or Pharmacology and Experimental Terapeutics. Vol.18, No. 6, pp. 917-922, 1990. Ellis D.E., Jone M., Larson R.A. and Schaeffer D.J. Organic constituents of mutagenic secondary effluents from wastewater treatment plants. Arch. Environ. Contam. Toxicl. Vol. 11, pp. 373-382, 1982. Encyclopedia of Chemical Technology, 1982. Engels W.H., Weidenhaupt H.J.Abele M., Pieroth M., Hofmann W. Ullmann´s Encyclopedia of Industrial Chemistry, 5th ed., Elvers B., Hawkins S., Russey W., Shullz G. Eds. VCH: Weinheim, Vol. A23, pp. 365, 1993. Eslyn W.E. and Cassens D.L. Laboratory evaluation of selected fungicides for control of sapstain and mould on southern pine lumber. For. Prod. J. 33(4), pp. 65-67, 1983. Evans W.C. Biochemistry of the bacterial catabolism of aromatic compounds in anaerobic environments. Nature 270, pp. 17-22, 1977. Evans D. N. & Levisohn I. Fungitoxic properties of volatile corrosion inhibitors. Chemy Ind. pp. 1386, 1964. Fedorak P.M. and Westlake D.W.S. Microbial degradation of alkyl carbazoles in Norman Wells Crude Oil. Appl. Environ. Microbiol. Vol. 47, pp. 858-862, 1984. Fiehn O., Reemtsma T. and Jekel M. Extraction and analysis of various benzothiazoles from industrial wastwwater. Anal. Chim. Acta 295, pp. 297-305, 1994. Fiehn O., Wegener G., and Jekel M. Capilary electrophoretic analysis of organic anions in tannery wastewaterincluding high contents of chloride and sulfate. Int. J. Environ. Anal. Chem., 1997. Fiehn O., Wegener G., Jochimsen J. and Jekel M., Analysis of the ozonation of 2-mercaptobenzothiazole in water and tannery wastewater using sum parameters, liquid-and gas chromatography and capillary electrophoresis. Wat. Res. Vol. 32,No. 4, pp. 1075-1084, 1998 Fife T.H., Hutchins J.E.C.,Wang M.S. J. Am. Chem. Soc. 97, pp. 5878, 1975. Finnerty W.R. Microbial desulfurization and denitrogenation of fossil fuels. In energy Technology Processings of the Ninth Energy Technology conference. Washington, D.C. Government Institutes Inc., 1982. Finnerty W.R., Shocley K. and Attawy H. Microbial desulfurisation and denitrogenation of hydrocarbons. In microbial Enhanced Oil Recovery. Ed. J.E. Zeijic, D.C. Cooper, T.R. Jack, N. Kosaric, Oklahoma: Renn Well Books Tulsa, 1982. Foltinova P., Sutoris V., Blockinger G., Ebringer L. Antimicrobial effects of some benzothiazol derivatives. Folia Microbiol. 23, pp. 225-228, 1978. Foote C.S. and Peters G.W. Chemistry of singlet oxygen. XIV. Reactive intermediate in sulfide photooxidation. J. Am. Chem. Soc.Vol. 93, Iss. 15, pp. 3795-3796, 1971. Foote C.S., in H. H. Wasserman and R. W. Murray (eds): Singlet Oxygen, Academic Press, New York, pp. 139-161, 1979. Fowler W.M., Russell A.E., Kruger I.H., and Pinchuck S.C. The extraction and high performance liquid chromatography (HPLC) determination of TCMTB samples of treated wet blue. J. Soc. Leather Technol. Chem., 71, pp. 100-107, 1987. Franke S., Hildebrandt S., Schwarzbauer J., Link M., Francke W. Organic compounds as contaminants of the Elbe river and ist tributaries part II, GC-MS screening of contaminants in the Elbe water. Fresenius`Z Anal. Chim. 353, pp. 39-49, 1995.

116

Frensch Klaus, Jens-Uwe Hahn, Kasten Levsen, Josef Nießen, Heinz Friedrich Schöler und Dirk Schoenen. Lösemittel aus dem Anstrichmaterial eines Hochbehälters als Ursache für eine koloniezahlerhöhung in Trinkwasser. Vom Wasser, 68, pp. 101-109, 1987. Fukuoka M., Tanaka A. The metabolic origin of the sulfur atom in the sulfhydryl group of the 2-thiobenzothiazole metabolites derived from benzothiazolyl sulfenamide in rat.Arch. Toxicol. 61, pp. 158-160, 1987. Galloway and Cooper. Fungicides for treating wet-blueHides; LIRI Research Bulletin No. 649, 1973. Gaudy Jr.A., and Rozich A.F. Design and operational model for activated sludge treating inhibitory carbon sources. Civil Eng. Pract. Design Eng. 2, pp. 55-70, 1982 Gaya M.A. and Knapp J.S. The microbial degradation of benzothiazoles. J. Appl. Microbiol. 83: pp. 327-334,1997 Gaya M.A. and Knapp J.S. Removal of 2-mercaptobenzothiazole by activated sludge a cautionary note. Elsevier Science, Vol. 32, pp. 3786-3789, 1998. Gledhill W.E. Proposed standard practice for the determination of the anaerobic biodegradability of organic chemicals. Working document. Draft 2, no. 35.24. American Society for Testing Materials, Philadelphia, 1979. Gold L.S., Slone T.H., Stern B.R. and Bernstein L. Comparison of target organs of carcinogenity for mutagenic and non-mutagenic chemicals. Mutat. Res. 296, 75-100, 1993. Gujer W. and Zehnder A. J.B. Conversion processes in anaerobic digestion. Water Sci. Tech. 15(8/9), pp. 127-167, 1983 Hahn J. Kühlabwasser, der Entwurf des Anhang 31 zu § 7a WHG. UTA 1993. Haines E., Cook G.L., and Ball J.S. Gaseous decomposition products formed fy the action of light on organic sulfur compounds. J. AM. Chem. Soc. Vol. 78, Iss. 20, pp. 5213-5215, 1956. Halassa AF, Smith GEP. Substituted ammonium salts of benzothiazoline-2-thione. Nuclear magnetic resonance studies of ion pairs in polar and non polar media. J. Org. Chem. 38, pp. 1353-1355, 1973. Hanssen H.W., Henderson N.D. and Wared J.E.H. A Review of the Environmental Impact and Toxic Effects of TCMTB. British Columbia Ministry of Environment, Victoria, British Columbia, 1991. Hansson and Agrup. Stability of mercapto benzothiazole compounds. Contact Dermatitis 28, pp. 29-34, 1993. Headley J.V. GC-MS identification of organosulphur compounds in environmental samples. Biomedical and Environmental Mass Specrometry, Vol. 14, pp. 275-280, 1987 Healy J.B. and Young L.Y. Anaerobic biodegradation of eleven aromatic compounds to methane. Appl. Environ.. microbiol. 38, pp. 84-89, 1979. Hendriks A.J., Maas-Diepevieen J.L., Noordsu A. and van der Gaag M.A. Monitoring response of XAD-concentrated water in the rhine delta: a major part of toxic compounds remain unidentified. Water Res. 28, pp. 581-598, 1994. Hill I. R. and Wright S. L. Pesticide microbiology: Microbial aspects of pesticide behaviour in the environment. Academic Press, London, United Kingdom, 1978. Hinderer R.K., Myhr B., Jagannath D. R., Galloway S. M., Mann S.W., Riddle J.C. and Brusick D.J. Mutagenic evaluations of four rubber accelerators in a battery of in vitro mutagenic assays. Environ. Mutagen. 5, pp. 193-215, 1983. Hoigné J., Bader H., Haag W.R. and Staehelin J. Rate constants of reactions of ozone with organic and inorganic compoundsin water III Wat. Res. 19, pp. 993-1004, 1985.

117

Hu Jia-Z., Skrabal P., Zollinger H. Acomparison of absorption spectra of series of blue disperse dyes with the chlorimetric evaluation of their dyeings. Dyes and Pigments Vol. 8. Iss. 3, pp.189-209, 1987. International Sorbent Technology, IST, Chromatographie Produkte, Ishibe N., Odani M., Sunami M., I. Chem. Soc. Chem. Commun. 118, 1971. Irvine R.L. and Davis W.B. Use of sequencing batch reactor for waste treatment-CPC. International Campus Christ. Texas. In Proceeding of 26th Industrial Waste Conference, Purdue University. Lafayette Ind, 1971. Irvine R.L and Ketchum L.H. Jr. Sequencing Batch Reactors for biological wastewater treatment. CRC Crit. Rev. Environ. Control 18, pp. 255-294, 1989. Ito O. and Matsuda M. Evaluation of addition rates of p-chlorobenzenthiyl radicals to vinylmonomers by means of flash photolysis.J. Am. Chem. Soc. Vol. 101, Iss.7, pp. 1815-1819, 1979. Jeans C.L. and Li G.C. Screening of mutagenic pesticides using microbial systems. K´o Hsueh K´an. 6(8), pp. 780-788. CA, 1978. Jones M. H. and J.T. Woodcook. Determination of mercaptobenzothiazole (MBT) in flotation liquors by solvent extraction and ultraviolet spectrometry. Anal. Chem. 47, Iss. 1, pp.11-16, 1975. Jung J. H., Mclaughlin J. L., Stannard J. and Guin J. D. Isolation, via activity-directed fractionation, of mercaptobenzothiazole and dibenzothiazyl disulfide as 2 allergens responsible for tennis shoe dermatitis. Contact Dermatitis 19, pp. 254-259, 1988. Jungclaus Gregory A., Larry M. Games and Ronald A. Hites. Identification of trace organic compouds in tire manufactures plant wastewaters. Analytical chemistry, vol. 48, No. 13, November 1976. Jungclaus Gregory A., Lopez-Avilla V. and Hites R.A. Organic compounds in an industrial wastewater: Acase study of their environmental impact. Environmental Science & Technology. Vol. 12, pp. 88-96, January 1978 Junker F., Leisinger T. and Cook A. M. 3-sulfocatechol 2,3-dioxygenase and other dioxygenases (EC 1.13.11.2 and EC 1.14.12-) in the degradative pathways of 2-aminobenzenesulfonic, benzenesulfonic and 4-toluenesulfonic acids in alcaligenes sp. Strain O-1. Microbiology 140, pp. 1713-1722, 1994. Karelová E., Tomasovicová D. Aerobic degradation of 2-mercaptobenzothiazole by sulphur dacteria. Biológia Bratislava; 43 (7), pp. 617-622, 1988. Kennedy M.J. High-Performance Liquid Chromatographic Analysis of preservative-timber for 2-(thiocyanomethylthio)benzothiazole and methylene bisthiocyanate. Analyst, Vol. 111, June 1986. Kice J.L., in J.K. Kochi (ed.): Free Radicals, Vol. 2, Wiley, New York, pp. 711ff, 1973. Kice J.L. Mechanisms and reactivity in reactions of organic oxyacids of sulphur and their anhydrides. In V. Gold and D. Bethell, eds., Advances in physical Organic Chemistry, Academic Press, New York, NY, Vol. 17, pp. 65-181, 1980. Knight A.R., in S. Patai (ed.): Chemistry of the thiol Group, vol. 1, Wiley, New York, pp. 455ff, 1974. Krstulovic A. M. and Brown P. R. Reversed Phase High Performance Liquid Chromatography. John Wiley and Sohns, New York, 1982. Kuhn H.J., Braslawsky S.E., Schmidt R. Pure and App. Chem. 61,187-201, 1989. Kumata H., Takada H., Ogura N. Determination of 2-(4-morpholinyl) benzothiazole in environmental samples by gas chromatograph equipped with a flame photometric detector. Analy. Chem. 68. No. 11, pp. 1976-1981, 1996.

118

Larsen G.L., Bakke J.E., Feil V.J., Huwe J.K. In vitro metabolism of the methylthio group of 2-methylthiobenzothiazole by rat in liver. Xenobiotica 18, pp. 313, 1988. Leifer A. The kinetics of environmental aquatic photochemistry. American Chemical Society, Washington, DC, 1988. Leisinger T., Microorganisms and xenobiotic compounds. Experientia 39, pp. 1183-1191,1140 (5), 1983. Lin S. and Carlson R.M. Subsceptibility of environmentally important heterocycles to chemical disinfection: Reactions with aqueous chlorine, chlorine dioxine, and chloramine. Environ. Sci. Technol. 18, pp. 743-748, 1984. Liska I., Brouwer E. R., Ostheimer A.G.L., Lingeman H., Brinkman U.A.T., Geerdink R.B. and Mulder W.H. Rapid screening of a large group of polar pesticides in river water by on-line trace enrichment and column liquid chromatography. Inter. J. Environ. Anal. Chem. Vol. 47, pp. 267-291, 1992. Louter A.J.H., Rinkema F.D., Ghijsen R.T. and Brinkman A.T.H. Rapid identification of benzothiazole in river water with on-line Solid-Phase Extraction-Gas Chromatographiy- Mass selective Detection. Intern. J. Environ. anal. Chem.. Vol. 56, pp. 49-56, 1994 Lovley D. R., Baedeker M. J., Lonergan D. R., Cozzarelli I. M., Phillips E.J.P. and Siegel D. I. Oxidation of aromatic contaminants coupled to microbial ion oxidation. Nature 399, pp. 297-300, 1989. Lovley D. R. and Lonergan D. J. Anaerobic oxidation of toluene, phenol and p-cresol by the dissimilatory ion-reducting organism, GS-15. Appl. Environ. Microbial. 56, 1858-1864, 1990. Maciejewski A., Steer R.P. The photophysics, physical photochemistry, and related spectroscopy of thiocarbonyls. Chem. Rev. Vol. 93, pp. 67-98, 1993. Mainprize J., Knapp J.S. and Callely A.G. The fate of benzothiazole-2-sulfonic acid in biologically treated industrial effluents. J. Appl. Bacteriol. 40, pp. 285-291,1976. Makoto Ema, Jun Sakamoto, Thoshimi Murai and Hironoshin Kawasaki. Evaluation of the teratogenic potential of the rubber accelerator didenzthiazyl disulfide in rats. Journal of applied toxicology, Vol. 9(6), pp. 413-417, 1989. Meijers A.P. and Van der leer R.Chr. The occurrence of organic micropollutants in the river Rhine and the river MAAS. Water Res. 10, pp. 597-604, 1976. Metcalfe J., Fox M., Carey J.H. Environ. Monit. Assess., 11, pp. 147, 1988. Miller T.L. and Wolin M.J. A serum bottle modification of the Hungate technique for cultivating obligate anaerobes. Appl. Microbiol. 27, pp. 985-987, 1974. Mylari Branavar L., Larson E.R., Beyer Thomas A., Zembrowski W.J., Aldinger C.E., Dee M.F., Siegel T.W. and Singleton D.H. Novel, potent aldose reductase inhibitors: 3,4-dihydro-4-oxo-3((5-8trifluoromethyl)-2-benzothiazolyl)methyl)-1-phtaladizineacetic acid (zopolrestat) and cogeners. J. Med. Chem. Vol. 34, Iss. 1, pp. 108-122,1991. Newcomb M., Parks Seung U. N-hydroxypyridine-2-thione esters as radicals precursors in kinetic studies, measurements of rate constants for hydrogen-atom-abstraction reactions. J. Am. Chem. Soc. Vol. 108, Iss. 14, pp. 4132-4134, 1986. Nikl Donna l. and Anthony P. Farrell. Reduced swimming performance and gill structural changes in juvenile salmonids exposed to 2-(thiocyanomethylthio)benzothiazole. Aquatic Toxicol. 27, pp. 245-264, 1993. Nippon Kayaku, Japanese Patent 06 136 283, Chem. Abs. 121: 257927r, 1994. Nishiuchi Y. Toxicity of agrochemicals to some freshwater organisms. XXXVI. Suisan Zoshoku 24 (1), pp. 25-26, 1976.

119

Ostwald A. A., Griesbaum K., Hudson B. E. Organic sulfur compounds, XII. Factors determining a- 1,2-vs. 1,4-mechanism of radical reactions og conjugated diolefins. Co-oxidation with thiols by oxygen. J. Org. Chem. Vol. 28, Iss. 9, pp. 2355-2361, 1963. Owen W.F., Stuckey D.C., Healy J.B., Young Jr. L.Y. and McCarty P.L. Bioessay for monitoring biochemical methane potential and anaerobic toxicity. Water Res. Vol. 13, pp. 485-492, 1979. Parbery C. and Taylor C.D. Determination of methylene bis(thiocyanate) and-(Thiocyanomethylthio)benzo(d)thiazole in leather process liquors by high-performance liquid chromatography. Analyst Vol. 114, pp. 361-263, 1989. Párkányi C. and Abdelhamid A.O. Heterocycles 23, pp. 2917,1985. Pavlostathis S.G. and Gosset J.M. A kinetics model for anaerobic digestion of biological sludge. Biotechnology and Bioengineering Vol. 38, pp. 1519-1530, 1986. Pavlostathis S.G. and Giraldo-Gomes E. Kinetics of anaerobic treatment. Water Sci. 24(8), pp. 35-39, 1991. Paxeus N. Organic pollutants in the effluent of large wastewater treatment plants in Sweden. Wat. Res. Vol. 30, pp. 1115-1122, 1996. Perraud R., Papazian M., Krahe E. Zersetzung biocider Verbindungen in wässerigem Milieu. GIT Fachz Lab. 4, pp. 312-316,1995. Pesticide Chemistry. Studies in Environmental Science 32, New York, 1998. Platford R.F. The octanol-water partitioning of some hydrophobic and hydrophilic compounds. Chemosphere 12, pp. 1107-1111, 1983. Puig A., Ormad P., Roche P., Sarasa J., Gimeno E., Ovelleiro J.L. Wastewater from the manufacture of rubber vulcanization accelerators: Characterization, downtream monitoring and chemical treatment. Journal of Chromatography A, vol. 733, pp.511-522, 1996. Reddy C.M. and Quinn J.G. Environmental chemistry of benzothiazoles derived from rubber. Environ. Sci. Technol. Vol. 31, pp. 2847-2853, 1997. Reichert J.K., Lochtman J., Müller R.J., Könen I. Jahres Tagung der FG Wasserchemie in der GdCh. Badenweiler, May 1993. Reemtsma T., Fiehn O. and Jekel M. Extraction and analysis of various benzothiazoles from industrial wastewater. Analytica Chemica Acta 295, pp. 297-305, 1994. Reemtsma T., Fiehn O., Kalnowski G. and Jekel M. Microbial transformations and biological effects of fungicide-derived Benzothiazoles determined in industrial wastewater. Environmental Science & Technology, Vol. 29, No. 2, 1995. Reemtsma T. Determination of 2-substituted benzothiazoles of industrial use from water by liquid chromatography/electrospray ionization tandem mass spectrometry. Rapid Commun. Mass Spectrom. 14, pp. 16112-1618, 2000. Rennie P.J. Determination of 2-mercaptobenzothiazole in river water by high performance liquid chromatography. Chromatographia Vol. 26, pp. 297-299, 1988. Reprika V.I., Dokudovs S.A., Umrikhina R.A. and Samokhina V.A., Maximum permissible concentrations of benzothiazole and 2-mercaptobenzothiazole during biochemical treatment of wastewaters. Khim. Prom-st. Ukr. Vol. 10, pp. 598-599, 1983. Rivera J., Ventura F., Caixach J., De Torres M. and Figueras A.. GC / MS, HPLC and FAB mass spectroscopic analysis of organic micropollutants in Barcelona´s water supply. Inter. J. Environ. Anal. Chem. Vol. 29, pp. 15-35, 1987. Rogge W.F., Hildemann L.M., Mazurek M.A., Cass G.R., Silmoneit BRT. Source of fine organig aerosol. 3. Road dust, tire debris, and organometallic brake lining dust: roads as sources and sinks. Environ. Sc. Technol. Vol. 27, Iss. 9, pp. 1892-1904, 1993.

120

Rozich A.F., Gaudy Jr A.F. and D`Adamo P.C. Selection of growth rate model for activated sludge treating phenol. Wat. Res. 19, pp. 481-490, 1985. Runge G., Steinhart H. Determination of volatile sulphur compounds in the edible part of carp. Agribiol. Res. Vol. 43, pp. 155-163, 1990. Russel A.E., Pinchuck S.C. and Cooper D.R., J. Soc. Leather Technol. Chem., 69, 135, 1985. Russel A.E., Pinchuck S.C. Review of laboratory Fungicide Surveys for wet-blue Protective Storage; LIRI Research Bulletin No. 899, 1985. Santodonato J., Davis LN, Howard PH, Saxena J. Investigations of selected potential environmental contaminants: mercaptobenzothiazoles. Final report (June 1976), contact no 68-01-3128. prepared for: office of toxic substances U.S. Environmental Protection Agency, USA, 1976. Sax N. I., Danderous Properties of Industrial Materials, Van Nostrand Reinhold, New York, 1984. Saxton Wilbour L. Emergence temperature indices and relative retention times of pesticides and industrial chemicals determined by linear programmed temperature gas chromatography. Journal of Chromatography vol. 393, pp. 175-194, 1987. Schwarzenbach R.P.,Gschwend P.G., Imboden. D. Environmental Organic Chemistry; wiley: New York, 1993. Scott B. F., Struger J. and Tse H. Examination of surface water samples using GC/Atomic emissin defection. Inter. J. Environ. Anal. Chem. Vol. #, pp. 1-14, 1995. Scott Brian F., Edward Sverko and James R. Maguire. Determination of benzothiazole and alkylphosphates in water samples from the great lakes drainage basin by gas chromatography/atomic emission detection. Water Qual. Res. J. Canada, 1996. Shelton D.R., Boyd S.A. and Tiedje J. M.. Anaerobic biodegradation of Phthalic acid esters in sludge. Environ. Sci. Technol. Vol. 18, pp. 93-97, 1984. Shuto A., Ohgai M., Eto M. Screening of tryptophan synthase inhibitors as leads of herbicide candidates. J. Pestic. Sci. Vol. 14, pp. 69-74, 1989. Shuttleworth L., Weaver MA. In. Warning DR, Hallas G., editors. The chemistry and application of dyes. New York: Plenium, (chapter 4), 1990. Simpson C.F., ed. Techniques in Liquid Chromatography. John Wiley and Sons. New York, 1982. Sitting S. (Editor), Handbook of Toxic and Hazardous Chemicals, Noyes Publications, Park Ridge, nj, 1981 Snyder L. R. and Kirkland JJ. J. Introduction to Modern Liquid Chromatography. John Wiley and Sons. New York, 1979. Spies Robert B., Andresen, Brian D., Rice, David W., Jr. Lawrence. Benzothiazoles in estuarine sediments as indicators of street runoff. Nature (London), 327 (6124), 697-9 (Eng), 1987. Steer R.P., Ramamurthy V. Photolytics and intramolecular photochemisty of thione in solution. Acc. Chem. Res. Vol. 21, Iss. 10, pp. 380-386, 1988. Stierle Andrea A. and John H. Cardellina. Benzothiazoles from putative bacterial Symbiont of the marine sponge Tedania ignis. Tetrahedron Letters, Vol. 32, No. 37, pp. 4847-4848, 1991. Stolková Magdaléna, kaszonyi A., Liptaj T., Hronec M. Determination of the products of the reaction of N-alkyl-2-benzothiazolesulfinamide by high-performance liquid chromatography with acetone. J. of Chromatography A, Vol. 847, pp. 351-358, 1999. Storrs F J. Dermatitis from clothing and shoes. In: Fisher AA (ed): Contact dermatitis. Philadelphia: Lea and Febiger, pp. 283-337, 1986. Standard methods for the determination of water and wastewater, 16th edition, 1985.

121

Stumpf B., Abraham W.R. and Kieslich K. Microbial transformations of N-(4-chlorphenyl)-benzothiazolone. Appl. Microbial Biotechnol. Vol. 23, pp. 215-217, 1986. Suzuki N., Sano K., Tani N. and Izawa Y. Heterocycles 16, pp. 1133, 1981. Tarvin D. and Buswell A.M.. The methane fermentation of organic acids and carbohydrates. J. am. Chem. Soc. Vol. 56, pp. 1751-1755, 1934. Taylor J.S. Rubber. In Fisher A. (ed): Contact Dermatitis. Philadelphia: Lea & Febiger, pp. 603-643, 1986. Theobald N., Lange W., Gählert W., Renner F. Mass spectrometric investigations of water extracts from the river Elbe for the determination of potential inputs pollutants into the North Sea. Fresenius Z. Anal. Chem. Vol. 353, pp. 50-56, 1995. Towns A.D. Developments in azo disperse dyes derived from heterocyclic diazo components. Dyes Pigmments Vol. 42, pp. 3-28, 1999. Ullmann´s Encyclopedia of indusrial chemistry, Vol. A 23, 1993. United States International Trade Commission synthetic organic chemicals. United States Production and sales 1981 USITC Publication 1292. US Government Printing Office. Washington DC, 1982. Usov G.P., Tryakhova T.I., Svetlakova G.N. and Rozanova L.I. Biotesting of industrial wastewaters from foam rubber manufacture. Khim. Technol. Vody. Vol. 9, pp. 182-184, 1987. Wallnöfer P., Tillmanns G., Thomas R., Wünsche C., Kurz J. und Jarczyk H.J. Mikrobieller Abbau des Herbizids Methabenzthiazuron und Identifizierung der Metaboliten. Chemosphere No. 5, pp. 377-382, 1976. Wamhoff H. and Ertas M., Angrew. Chem. Vol. 94, pp. 800, 1982. Warner J. S., T. M. Engel and P. J. Mondron. Determination of MBTS and TCMTB in industrial and municipal wastewaters. NTIS, PB 85-189025, 1985. Warshawsky D. Environmental sources, carcinogenicity, mutagenicity, metabolism and DNA binding nitrogen and sulphur heterocyclic aromatics. Environ. Carcino. & Ecotox. Rev. C(10), pp. 1-71, 1992. Welch D.I. and Watts C.D. Collection and identification of trace organic compounds in atmospheric deposition from a semi rural site in the UK. Intern. J. Environ. Anal. Chem., Vol. 38, pp.185-198, 1990. Wiedemeier T. et al. Technical Protocol for Implementing Intrinsic Remediation with long-term monitoring of natural attenuation of fuel contamination dissolved in groundwater. Vol. 1. Technology Transfer Division Brooks, San Antonio, TX, 1995A Wiedemeier T. et al. Approximation of biodegradation rate constants for monoaromatic hydrocarbons (BTEX) in groundwater. Ground Water Monit. Remediat. 16, pp. 186-194, 1996. Wilson K., Chissick H., Fowler A.M., frearson F.J., Gittins M. and Swinbourne F.J. Metabolism of benzothiazole. I identification of ring-cleavage products. Xenobiotica Vol. 21, No. 9, pp. 1179-1183, 1991. Woltersdorf O.W., Jr., H. Schwam, J.B. Bicking, S.L.Brown, S.G. deSolms, D.R. Fishman, S.L: Graham, P.D. Gautheron, J.M. Hoffman, R.D. Larson, W.S. Lee, S.R. Michelson, C.M. Robb, N.L. Share, K.K. Shepard, A.M. Sondey, K.M. Strohmaier, M.F. Sugrue and M.P. Viader, J. Med. Chem. Vol. 32, pp. 2486, 1989. Wright R., Lee W., Quinn J., Vshisth P., Reddy C. Assessment of Water Pollutants from Asphalt Pavements Containing Recycled Rubber in Rhode Island; final report to Rhode Island Department of Transportation, Providence, RI, on Grant ME-534, 1996. Yoshioka Y., Ose Y. Environ. Toxicol. & Water Qual. Vol. 8, pp. 87, 1993. Zehnder A. J. B. and Wuhrmann K. Physiology of methanobacterium strain AZ. Arch. Microbiol. 111, pp. 1999-205, 1977.

122

Zeikus J.G. Microbial populations in digesters. In anaerobic digestion, Proceeding of the international Sdymposium on anaerobic Digestion, University College, Cardiff, Wales, 1979. Zepp R. G. and Cline D. M. Rates of direct photolysis in aquatic environment. Enviro. Sci. Technol. 11, pp. 259-366, 1977.

Microbial and photolytic degradation of benzothiazoles in ... · Microbial and photolytic...

Documents

Transcript of Microbial and photolytic degradation of benzothiazoles in ... · Microbial and photolytic...