degradation of priority compounds by uv and uv- oxidation · “advanced oxidation techniques”...

Degradation of priority compounds by UV and UV-oxidation

December 2010

© 2010 TECHNEAU TECHNEAU is an Integrated Project Funded by the European Commission under the Sustainable Development, Global Change and Ecosystems Thematic Priority Area.All rights reserved. No part of this book may be reproduced, stored in a database or retrieval system, or published, in any form or in any way, electronically, mechanically, by print, photoprint, microfilm or any other means without prior written permission from the publisher

Degradation of priority compounds by UV and UV-oxidation

December 2010

Colofon

Title

Advanced oxidation processes Authors C.H.M. Hofman-Caris, D.J.H. Harmsen, E.F. Beerendonk Quality Assurance M.M. Nederlof Deliverable number D 2.4.1.2b

Advanced oxidation processes KWR Watercycle Research Institute

© TECHNEAU - 1 - December 2010

Preface

Techneau is a European research project on drinking water purification, funded by the European Commission. Part of the research is focused on “Advanced Oxidation Processes” (AOPs). These processes mainly rely on the formation of reactive and short-lived oxygen containing intermediates such as hydroxyl radicals (•OH). They are formed under the influence of UV irradiance, and they are powerful, non-selective oxidants, which can be used to convert organic micro pollutants to (better biodegradable) small compounds. Within the framework of the Techneau project experiments were carried out to investigate whether a representative set of emerging compounds can be converted by means of the UV/H2O2 process, and how much energy is required for each compound. Furthermore, a study of recent literature (past 5 years) is included in this report, describing recent developments in the field of Advanced Oxidation Processes.



Summary

Techneau is an international research project on drinking water purification, funded by the European Commission. Part of the research is focused on “Advanced Oxidation Processes” (AOPs), which may be applied to convert organic micro pollutants to (better biodegradable) small compounds. This report consists of two parts. In the first part, a literature overview is given on the present situation and recent developments with regard to AOPs. In the second part, pilot scale experiments to convert a number of known micro pollutants in water by UV/H2O2 oxidation are described. For each compound the conversion yield and the process efficiency (“electrical energy per order”) in the pilot reactor are measured. It was found, that the UV/H2O2 process is very effective for the conversion of many emerging substances, like alachlor, atrazine, desisopropylatrazine, carbamazepine, sulfamethoxazole, para-chlorobenzoic acid (pCBA). However, for some substances, like atrazine-desethyl and MTBE the conversion did not exceed 50%. These data are in good accordance with literature data. We also determined the Electrical Energy per Order (EEO; the amount of electrical energy required to obtain a conversion of 90% for a certain compound). In our pilot reactor, which has not been optimized as full scale reactors, an average EEO of about 1kWh/m3 is found. This too is in accordance with literature data. It means that the UV/H2O2 process requires a relatively large amount of energy. However, the energy demand strongly depends on lamp type, water quality and reactor efficiency. By optimizing the reactor design, it will be possible to lower the EEO. In order to obtain enough hydroxyl radicals, an excess of H2O2 has to be added to the reactor. This excess has to be removed after the reaction, which in general is achieved by filtration over activated carbon. This step has the additional advantage, that possibly harmful compounds, that have not been converted or even may have been formed during the AOP, also will be removed from the water.



Contents

Preface 1

Summary 2

Contents 3

1 Introduction 4

2 Literature overview recent developments in Advanced Oxidation Processes 6

2.1 Introduction to AOP’s 6

2.2 AOPs in general 7

2.3 Ozone combined with UV 8

2.4 Processes with H2O2 11 2.4.1 UV/ H2O2 for disinfection 11 2.4.2 Removal of various compounds by means of a UV/ H2O2 process 12 2.4.3 Modelling 15 2.4.4 Byproducts of UV processes 16 2.4.5 Ozone in combination with Hydrogen Peroxide 16

2.5 Heterogeneous catalysis 17

2.6 Ultrasonication 21

2.7 The Fenton process 22

2.8 Summary of literature review. 23

3 Pilot experiments with emerging substances 25

3.1 Introduction 25

3.2 Emerging substances 25

3.3 Experimental section 25

3.4 Results and discussion 28

4 Conclusions 32

5 Literature 33

I Appendix 39

II Appendix 40



1 Introduction

Water quality may be threatened by the presence of new contaminants, increasing concentrations of contaminants and seasonal or diurnal variations in water sources. Solid barriers against (organic) micro pollutants in drinking water are becoming more and more important, as there is an increasing variation in pollutants in ground and surface water, and a continuous improvement in the analytical techniques. Priority compounds are compounds which, because of their properties, emissions and/or concentrations in the environment represent a more than negligible risk for man or environment. In general they can be characterized as persistent, bio accumulative, ecotoxic, harmful for human health, carcinogenic, repro-toxic, mutagenic or hormone disturbant. There is a lot of international interest in the removal of these emerging substances. The use of UV light for disinfection of water has been known for several decades. Since the mid nineties the combination of UV light and hydrogen peroxide (H2O2) has been studied as a barrier for organic micro pollutants. This combination is known as an “advanced oxidation process.” “Advanced Oxidation Techniques” (AOPs) are promising techniques to efficiently and effectively convert these compounds into better biodegradable or less harmful substances. Several types of AOPs are being developed. Often these techniques are based on ozone, hydrogen peroxide, UV irradiation and combinations of these techniques. Nowadays, AOPs are becoming important barriers against organic micro pollutants in drinking water treatment. Application of UV combined with H2O2 oxidation, followed by filtration over active carbon, has been proven to be a solid and flexible barrier against a broad range of organic micro pollutants By means of UV irradiation the H2O2 is converted into hydroxyl radicals, which aselectively react with all kinds of organic compounds. Apart from this oxidation process, some compounds will be converted by direct fotolysis as a result of the UV irradiation. The state-of-the-art UV/H2O2 technology is equipped with Medium Pressure (MP) UV-lamps, which have a higher power and a broader UV spectrum than Lower Pressure (LP) lamps. On the other hand, LP lamps have a higher energy efficiency, and a longer expected life span. As they only emit UV irradiance with a wavelength of 253.7 nm, they are less effective for photolysis, but they also result in the formation of less by-products. The work described in this report is part of “work package 2.4” (2.4.1a) of the Techneau project.”Water treatment” is the subject of work area 2. Within this framework, Workpackage 2.4 focuses on oxidation processes, and more specific, 2.4.1 on UV-based processes. The aim of 2.4.1a is to contribute to a report with additional test results performed on the gaps in knowledge to ensure a safe application of UV disinfection and UV/H2O2 oxidation. In this report, first an overview is given on the recent developments in the field of AOPs. Which types of AOPs are there? What are their characteristics,



advantages and disadvantages, and which topics are getting most attention at present? Furthermore, pilot scale experiments were carried out with a UV/H2O2 process to study the conversion of a series of selected emerging substances.



2 Literature overview recent developments in Advanced Oxidation Processes

In this chapter an overview is given on the open literature published during the past five years on application of AOPs. It describes the various types of AOPs, and their characteristics. It mainly focuses on the topics which receive most attention in this field at present.

2.1 Introduction to AOP’s

Many organic compounds are able to absorb light of a certain wavelength, as a result of which reactions will occur. This photolysis can be applied to determine the concentration of certain compounds (like e.g. unsaturated carbons or aromatic compounds, pharmaceuticals) in water (Lee 2008, Rosario-Ortiz 2010). It can also be applied to convert organic micro pollutants in water. However, not all compounds are sensitive to photolysis. In sunlit surface waters or in the earth’s atmosphere all kinds of photo-initiated processes take place. Irradiation may result in the formation of reactive species that may trigger a series of reactions in which compounds, insensitive towards direct photolysis, may be converted after all. Advanced Oxidation Processes are based on these natural processes. They mainly rely on the formation of reactive and short-lived oxygen containing intermediates such as hydroxyl radicals (•OH). These radicals are formed under the influence of UV irradiance, and they are powerful, non-selective oxidants, which can be used to convert also those compounds that are not sensitive to photolysis (Oppenländer 2002). These AOPs are becoming more and widely utilized in water treatment (Cooper 2008). In literature several AOPs have been described: UV combined with H2O2, UV combined with ozone (O3), O3 in combination with H2O2, UV combined with O3 and H2O2, photo- and electron Fenton, heterogeneous photo oxidation using titanium dioxide (TiO2/hυ), γ-radiologis and sonolysis (Ikehata 2006). The Fenton process is based on decomposition of H2O2 in the presence of Fe(II) as a catalyst. Radiologis means the organic micropollutants are decomposed by means of γ irradiation, whereas with sonolysis the molecular bonds of the pollutants are broken by means of ultrasound. These topics will be explained in more detail in the following sections. Although advanced treatment technologies have been shown to be effective, nearly all of these technologies are too energy and/or material intensive to be applied to wastewater treatment, especially membrane processes and activated carbon adsorption. Introduction of ozonation or AOPs before or after a biological treatment process may be feasible because of the chemical/photochemical oxidation renders recalcitrant xenobiotics more biodegradable and less toxic, and improves their degradation in the following treatment process or in the environment (Ikehata 2006). However, in general AOPs are not very cheap techniques.



This is an overview of the recent developments in the area of Advanced Oxidation Processes.

2.2 AOPs in general

Advanced oxidation processes (AOPs) have emerged as powerful technologies and are widely used in drinking water treatment for the degradation of all kinds of compounds, including endocrine disrupting chemicals (Badriyha 2007). An overview of several AOPs and their effectiveness in the degradation of all kinds of organic micro pollutants was described by e.g. Wang et al. (degradation of chloro acetic acid and its derivatives) and Ikehata et al. (pharmaceuticals) (Wang 2009; Ikehata 2006). The performance of AOPs is affected by the presence of other water and waste water constituents, such as natural organic matter, dissolved and suspended solids, and alkalinity as well as by water pH and temperature (Ikehata 2006). For example, suspended solids and color can hinder photochemical reactions by light scattering and absorption and maybe impair the performance of photochemical AOPs. Carbonate, bicarbonate and chloride ions, as well as some natural organic compounds are known to act as radical scavengers. These compounds compete with target pollutants for hydroxyl radicals; therefore their presence increases oxidant demands and lowers the treatment efficiency. In addition, the costs of materials and equipment, as well as energy requirements and efficiency must be taken into account when assessing the overall performance of AOPs. In principle advanced oxidation techniques can result in the mineralization of organic contaminants. However, in most cases this is not the goal. By means of oxidation processes, most micro pollutants will be converted into smaller molecules, which, in general, are better biodegradable. Ikehate et al. (Ikehata 2006, 2008) give an overview of AOP applications for conversion of pesticides and pharmaceuticals. Most of the modern pharmaceuticals are small organic compounds with a molecular weight below 500 Daltons. These chemicals are moderately water soluble as well as lipophilic to be bioavailable and biologically active. These compounds undergo a series of oxidation and spontaneous transformation reactions. Primary degradation products often are degraded further. Complete mineralization of e.g. pharmaceuticals in general is not an absolute requirement in both water and wastewater treatment. However, the disappearance of parent compounds does not always indicate the successful treatment because the degraded products may be more biologically active (or toxic) as the parent compounds. Other examples are the formation of possibly carcinogenic bromate, when bromide containing water is treated with ozone, the formation of disinfection byproducts during chlorination, and their possible reactions during advanced oxidation processes. Furthermore, the natural organic material, present in water, may also be converted, which in some cases may result in the formation of genotoxic compounds. Therefore, it is desirable to assess the residual toxicity or estrogenicity after the treatment to ensure the safety of treated waste water or water. The



degradation by-products after ozonation or advanced oxidation may be characterized in terms of their chemical structure, toxicity and biodegradability.

2.3 Ozone combined with UV

Ozone has been used for disinfection and purification purposes for a long time now. Molecular ozone reacts selectively with organic molecules having nucleophilic moieties such as carbon-carbon double bonds, aromatic rings, and functional groups bearing sulfur, phosphorous, nitrogen and oxygen atoms. In water treatment plants, the combination of ozone and UV as separate treatment steps is well known. The UV irradiation in such plants is used for disinfection purposes, after ozonation of the micro pollutants (Ried 2006). Ozone based AOPS, including UV/ O3, O3/ H2O2 and O3/ H2O2/UV possess a high potential form removing pollutants from water sources and from wastewater (Song 2009). Apart from a direct reaction of ozone with the organic pollutants, aqueous solutions of ozone can give reactions, resulting in the formation of •OH radicals (Peyton 1988). Therefore, ozonation at high pH (>8) for example, is also considered as an AOP, because of the enhanced generation of hydroxyl radicals under such conditions (Ikehata 2006). The combination of ozone and UV light has proven effective with many types of pollutants. There is a marked difference between ozone and •OH radicals with respect to their reactions with the water matrix. The rate of ozone consumption by the water matrix decreases rapidly with increasing ozonation, as ozone-reactive components are consumed and only the more or less ozone-refractive components remain (Sonntag 2007). In contrast, the rate of •OH radicals with the water matrix remains practically the same as long as mineralization is not reached, and mineralization never is the goal in micro pollutant degradation. In general, it will be sufficient when these compounds are degraded to become less toxic or better biodegradable. Even bicarbonate/carbonate react with an appreciable rate. Hydroxyl radicals are non-selective towards various organic and inorganic compounds, and they may react through hydrogen abstraction, radical-radical reactions, electrophillic addition and electron transfer reactions. This may eventually lead to complete mineralization of organic compounds (Ikehata 2006). The independence of the rate of the •OH radical reaction as the •OH radical reaction proceeds has the advantage that it is much easier to model •OH radical reactions than ozone reactions. Furthermore, not all pollutants can be well oxidized and destroyed by ozone, whereas the •OH radical can degrade micro pollutants when other oxidants such as O3 fail. Ozone in aqueous solution decomposes through a complex mechanism initiated by reaction with a hydroxide ion and followed by formation of several radical oxidizing species, such as HO, HO2 and H O3. There are a wide variety of compounds able to promote or inhibit the chain reaction processes. Promoters of the free-radical reaction are substances capable of regenerating the superoxide anion (O22-) from the hydroxyl radical. Common organic promoters include formic and glycoxylic acids, primary



alcohols and humic acids. The inhibitors of the free-radical reaction are compounds capable of consuming hydroxyl radicals without regenerating the superoxide anion. These include bicarbonate and carbonate ions, tertiary alcohols like tert-butanol and some humic substances, and sulfate (Rodriguez 2008, Peyton 1988). The homogeneous rate of production of hydroxyl radicals from ozone strongly depends on pH, since the active species in the initiation of the ozone decomposition mechanism are HO- and HO2-, the concentrations of which are directly related to the concentration of hydroxide. The most important problem in using ozone for water purification is the formation of bromate in bromide containing water. A recommended strategy to reduce the bromate formation is to use pH<7 because bromate formation also strongly depends on pH. Under acidic conditions the formation of hydroxyl radicals and the rate of mineralization are much lower than in conventional ozonation (neutral pH). According to Kornmueller (Kornmueller 2007) in marine water this is due to the masking of HOBr for further reactions with O3 as well as by side reactions of ozone with ammonia. Another stable byproduct in these types of water is bromoform (CHBr3). Ikehata et al. (Ikehata 2006, 2008) give an overview of AOP applications for conversion of pesticides and pharmaceuticals. Very often during water treatment these compounds are converted by means of ozone based processes. Wang et al. (Wang 2009) studied several AOPs for the conversion of halo acetic acids (HAAs) like dichloroacetic acid and trichloroacetic acid in water. They found that the combination of ozone and UV is the most suitable for the decomposition of these compounds. Single O3 or UV treatment did not result in perceptible decompositions of HAAs within the applied reaction time. The combination of O3 and UV appeared to be suitable, although continuous adding of O3 to the system appeared not to be an efficient method to improve the decomposition. The reaction can be described using a pseudo-first-order reaction model under constant initial O3 concentration and fixed UV irradiation. The presence of humid acids was shown to cause a decrease in rate constants in the O3/UV process. According to Song et al. (Song 2009) the cost and synergy of treatment with UV/ O3 depends on the type and concentration of the target substance, and on the degree of removal required. Several variables affect the process: pH, ozone dose and bulk temperature. The authors showed that mineralization of 4-chloro-3, 5-dimethylphenol (PCMX) with ozone was more effective at pH 11.0 than 4.0, indicating that more hydroxyl free radicals were formed from ozone decomposition at high pH, while the molecular ozone remained as the main oxidant at low pH. Generally, the rate of contaminant degradation increased due to the elevated pH. However, the mineralization of PCMX did not follow this trend when UV irradiation was combined with the ozonation system. Increasing pH in the range 4.0-11.0 did not show an enhancement of the mineralization efficiency. In contrast, the lower the pH of the solution, the higher the reaction rate. The rate of the reaction of •OH with the various phosphate species added to the solution to buffer the pH in the experiments would be expected to be slowest in the pH range 2.8-4.5. Besides, the self-combination of •OH radicals at high pH may be responsible for the slow



mineralization of PCMX under alkaline conditions. Furthermore, high pH in the solution leads to the creation of more free radical scavengers (i.e. C O32-, HC O3-, etc.) derived from mineralization of organic material, resulting in a decrease in the concentration of •OH. However, at pH 2.0 molecular ozone is present, due to the slower generation of hydroxyl radicals from ozone decomposition, resulting in a poor mineralization rate. Additionally, the degree of ionization is a key factor contributing to the UV/ O3 oxidation of PCMX with a pKa of 9.7. Since ozone prefers to react with target pollutants in the ionized form rather than the undissociated forms, the mineralization rate is expected to be slower at more acidic pH. Increasing the ozone dose promotes light absorption and utilization and therefore increases the degradation rate. Nonetheless, the trend of the effect should not be extrapolated to extremely high doses of zone, since the excess ozone dose cannot efficiently enhance the mineralization of PCMX. The solubility of ozone is limited at a certain temperature, and applying a higher than optimum dose means that the concentration of hydroxyl free radicals and ozone in solution are nearly invariant. Under these conditions ozone may even become a •OH scavenger itself. Besides, since the use of UV/ O3 is a costly oxidation process, the ozone consumed per TOC (total organic carbon) elimination is a fundamental parameter to take into account. The effect of the temperature on the UV/ O3 mineralization rate appeared to be complicated. Both positive and negative effects on TOC removal rate were encountered with respect to temperature dependence: a high temperature promoted ozone decay and the reactions of the hydroxyl radicals and other oxidative species with organic compounds in the bulk solution. However, meanwhile a high temperature decreases the solubility of ozone in water, thus limiting the enhancement of the reaction rate obtained by the temperature increase. The authors also showed that mineralization of PCMX was accompanied by the formation of aromatic intermediates, carboxylic acids and chlorine-containing ions. Rui et al. (Rui 2006) describe the oxidation of dimethyl phthalate (DMP), which is dominated by ozone oxidation, both in the ozonation process and in a UV/ O3 process. The importance of •OH in the oxidation of DMP was enhanced as the concentration of DMP decreased in the UV/ O3 process. Baus et al. (Baus 2007) studied the conversion of fuel additives. These compounds (MTBE, TAME, ETBE en DIPE) hardly react with ozone, but •OH radicals do lead to a significant decrease in concentration of these compounds, although pH and the presence of (bi)carbonate radical scavengers play an important role in the efficiency of this process. The UV/ O3 process appeared to be more efficient than the UV/ H2O2 process for this purpose. The main disadvantage of applying UV/ O3 for fuel additive conversion are the costs involved for investment and maintenance for the process. Kornmueller (Kornmueller 2007) studied AOPs in marine waters, which appears to be considerably different from drinking, process and wastewater applications. The major secondary oxidants in marine waters are bromine species, which might form the DBPs of concern bromate and bromoform. This is especially important in case UV irradiation is combined with ozonation in



these waters. Marine waters are characterized by their complexity and variability, caused by geographic and seasonal alterations such as in water temperature, dissolved gases especially oxygen and by algae blooms. Furthermore, some variations are due to local conditions, which are influenced by tide, wind and ship passages and dredging activities in harbors as well as by the location and design of water intakes. Beside algae blooms, pollutants in harbors and waste water discharge in coastal waters are decisive for the oxygen demand of the water. Only a few parameters such as salinity, pH, temperature and dissolved oxygen content are widely available for different locations. While bromide concentration ranges vary from 10 to 1,000 µg/L in drinking water sources, seawater of 35 psu (practical salinity units) has a bromide concentration of 0.067 g/L(67000 µg/L), and concentrations of 19.35 g/L chloride, 2.71 g/L sulfate and 0.1142 g/L hydrogen carbonate. The higher salt content in marine waters has an impact on every mass-transfer dependent oxidation process, like ozonation. The effects depend on the kind, affinity and concentration of salts. The presence of chloride causes a lower mass transfer as well as a lower dissolved oxidant concentration compared to drinking water. The water temperature affects the mass transfer gas/liquid and the oxidation reactions. In general, the gas solubility decreases with increasing temperature, while the rate of chemical reactions rises according to the activation energies. In marine water, the radical reaction of ozone is suppressed in the initial stage due to the fast reaction of ozone with bromide. Byproduct formation (most importantly the formation of bromate and bromoform) is an important topic in water treatment. While in drinking water treatment chloroform may be the dominating DBP and the formation of bromoform normally is negligible, it is exactly the opposite in marine water chlorination. The formation of byproducts in marine water predominantly is influenced by pH, bromide, alkalinity, dissolved organic carbon, ammonia and the oxidant concentration applied. The complex interactions of the known initiators like iron, promoters (phosphate, carbonate etc.) and scavengers (e.g. carbonate, phosphate). Apart from the problems mentioned above, application of UV irradiation in coastal and harbor water may be hindered by the low transmission, which depends on the quality of the source water, particle size distribution, and previous mechanical separation.

2.4 Processes with H2O2

2.4.1 UV/ H2O2 for disinfection

Disinfection by absorption of UV irradiation by thymine or uracil bases in DNA or RNA is well known, however, the effect of the oxidation process on inactivation of pathogens has not been studied very intensively. Mamane et al (Mamane 2007) describe the inactivation of microorganisms by means of oxidation processes. In general, two primary mechanisms control the oxidant disinfection efficiency by hydroxyl radicals: (1) oxidation and disruption of the cell wall and membrane with resulting disintegration of the cell (oxidation



ability is due in part to its standard reduction potential of 2,7 V for hydroxyl radicals, and (2) diffusion of the disinfectant into the cell or particle where it may inactivate enzymes, damage intracellular components, interfere with protein synthesis etc. Diffusion of the disinfecting species into the cells is a function of the charge, molecular weight, and half-life of the disinfectant. Hydroxyl radicals react with most biological molecules at diffusion controlled rates. Therefore, disinfection by hydroxyl radicals may be limited by mass transfer through the cell wall or cell membrane. The authors found that disinfection, due to the presence of •OH radicals is very small compared to the damage from UV irradiation, although for viruses there may be some oxidative enhancements that can assist disinfection efficacy. According to Buchanan et al. (Buchanan 2006) H2O2 can also interfere with disinfection processes by reaction with chlorine.

2.4.2 Removal of various compounds by means of a UV/ H2O2 process

According to Kwon et al. (Kwon 2009) the hydroxyl radical is the most important strong oxidant in advanced oxidation processes, which has the strongest effect on the degradation of recalcitrant anthropogenic pollutants in water. Linden et al. (Linden 2007) studied the effectiveness of UV/ H2O2 treatment for endocrine-disrupting compounds (EDCs). Direct UV treatment of contaminants is effective only if the UV light emitted by a UV lamp is absorbed by the contaminant. The emission spectrum of a medium pressure mercury UV (MP) lamp overlaps much of the major absorbance features of the contaminants, and therefore it was expected that an MP lamp may be more effective at degrading EDCs than a low pressure mercury (LP) UV lamp. The authors found, that both LP and MP lamps can effectively be used in UV/ H2O2 processes to reduce the estrogenic activity in the water. Rosario et al (Rosario-Ortiz 2010) found that the efficacy of UV/ H2O2 treatment for the removal of pharmaceuticals from waste water is a function of not only the concentration of NOM but also of the pharmaceutical’s inherent reactivity towards hydroxyl radicals. The removal of pharmaceuticals also correlates with reductions in ultraviolet absorbance. Larger UV fluences are required to promote greater •OH exposure required to overcome the scavenging capacity of the wastewater and produce greater oxidation of pharmaceuticals. Six compounds were investigated (meprobamate, carbamazepine, dilantin, atenolol, primidone and trimethoprim). Of these compounds, meprobamate appeared to be the most recalcitrant of the compounds evaluated, and its removal appeared to be directly related to scavenging of wastewater. Tawabini et al (Tawabini 2009) found that for the removal of methyl tert-butyl ether (MTBE) a 150 Watt MP UV lamp showed better removal efficiency than an LP 15 Watt lamp. However, in order to be able to make a fair comparison, it would be better to compare the Electrical Energy per Order (EEO). This is the amount of electrical energy, required to convert a certain compound by one order of magnitude in a certain reactor. Li et al. (Li 2008) also studied the efficiency of a UV/ H2O2 process to remove MTBE and tertiary butyl alcohol (tBA) from a drinking water source. They found that several factors affect the performance of UV/ H2O2:



1. Reactivity of the target compound toward the •OH radical 2. UV lamp technology (low pressure, high output versus medium

pressure, high output) 3. Operational and water quality factors, such as H2O2 dosage, presence

of carbonate species, NOM, pH and reduced metal ions (iron and manganese).

The species present in the raw water (like NOM, nitrate and ferrous ions) can decrease the performance of a UV/ H2O2 process by either absorbing UV light or by scavenging the hydroxyl radicals. The pH of a solution determines the acid dissociation equilibrium of carbonate species, H2O2 and organic compounds in the process. The impact of the species distribution needs to be carefully balanced in the process design. For (bi)carbonate ions a lower pH is preferred because carbonate species have a higher reaction rate with •OH radicals than bicarbonate. For H2O2 (with pKa = 11.6) a higher pH is preferred because HO2- absorbs UV light more effectively than H2O2. Organic compounds, like organic acids and sulfa drugs, exist either in their protonated or unprotonated form (or even as a mixture of both), depending on pH. In general, the unprotonated forms react with the •OH radical at rate constants one or two orders of magnitude larger than their protonated forms. In some cases it may be beneficial to pre-treat the water to remove TOC, carbonate species and iron and manganese in the raw waters. According to Li et al. (Li 2008), effective pretreatment alternatives can be an ion exchange softening system with seawater regeneration (NaIX), softening using pellet reactors (pellet softening), weak acid ion exchange processes (WAIX), and high pH lime softening followed by reverse osmosis (RO). Dealkalization can be considered with all the alternatives in order to minimize the presence of alkalinity in the AOP. Kinetic models can be used to simulate the complex interaction of those factors and removal efficacy. Thus, useful information on the benefits of different pretreatment options can be obtained. Baus et al (Baus 2007) wrote a paper on the treatment of several fuel additives with AOPs. Only applying UV irradiation by means of a low pressure UV lamp did not yield any decline of MTBE in tap water. With the addition of H2O2 at 5 mg/L, no significant improvement of the removal efficiency could be observed. The H2O2 concentration has to be high enough to ensure that the absorption of UV irradiation yields an efficient conversion of H2O2 to •OH radicals. However, with the H2O2 concentration being too high, a hindering effect is achieved by the recombination of •OH radicals (Baus 2007, Li 2008). The H2O2 concentration applied by Baus and co-workers is rather high (50 mg/L, whereas in general such a concentration is about 5-10 mg/L). Nienow (Nienow 2009) successfully applied a UV/ H2O2 process to convert 3-MCPD from water. For the conversion of NDMA addition of H2O2 does not seem to improve the process, as this compound is very sensitive towards photolysis by UV irradiation (Plumlee 2008). Jo et al (Jo 2009) describe the removal of odorous aldehydes by means of UV/ H2O2. They found that direct UV photolysis is the main mechanism involved. Felis c.s. (Felis 2007) studied UV/ H2O2 versus only UV for the conversion of naproxen in drinking water and waste water. In urban waste water many organic and inorganic substances are present, which can “compete” with



naproxen for UV radiation. Therefore, the naproxen oxidation process in wastewater (especially at the beginning of the treatment) proceeded not as efficiently as oxidation of naproxen in drinking water. Min Rui et al (Rui 2006) also describe how the oxidation of dimethyl phthalate (DMP) by UV/ H2O2 is strongly affected by the presence of humic acids in water. Ikehata et al (Ikehata 2006) mentions that UV/ H2O2 treatment in order to convert pharmaceuticals may be less effective in case other organic matter is predominantly present, as the oxidant requirement can be exceedingly high in order to achieve effective degradation of trace target pollutants. NOM is a complex mixture of humic and fulvic acids, proteins, lipids, hydrocarbons and amino acids (Murray 2006), which is ubiquitously present in surface waters and poses several challenges to drinking water treatment options. The range of organic components found in NOM varies from water to water and seasonally as does its reactivity with chemical disinfectants such as chlorine. With respect to water quality, it can act as a precursor for disinfection byproducts (DPBs) and has potential to increase the biological regrowth potential of water in distribution systems (Sarathy 2007). During UV/ H2O2 AOP, NOM screens light needed for the photolysis of H2O2 and scavenges hydroxyl radicals, required for contaminant removal (Badriyha 2007). Badriyha c.s. tried to model the process kinetics of AOPs in the presence of NOM. Under strong advanced oxidation conditions (i.e. long irradiation time and/or high H2O2 concentration), studies have found that NOM becomes mineralized, indicated by a decrease in total organic carbon. From a water quality point of view, the removal of NOM is beneficial since this leads to reductions in the formation of (possibly harmful) DPBs and biological regrowth potential (Buchanan 2006, Sarathy 2007). According to Buchanan c.s. (Buchanan 2006) low molecular weight compounds, produced by the initial rapid breakdown of larger NOM compounds can be halogenated, leading to an increase in compounds like chloroform, dichlorobromomethane, dibromochloromethane and bromoform. Often, a decrease in pH is observed, due to the generation of CO2 in the solution. Furthermore, degradation products of NOM often are responsible for unpleasant odors and tastes, discoloration. However, the strong AOP conditions required for mineralization are not economically feasible. Commercial UV/ H2O2 systems, applied for the removal of trace organic contaminant in drinking water, operate at conditions sufficient for degradation of the target pollutants. Under these conditions NOM is not mineralized but rather partially oxidized. This may lead to increases in the formation potential of DBPs and biodegradability of NOM. On the other hand, UV irradiation of NOM in surface water may also result in the formation of •OH radicals without addition of H2O2 (Pereira 2007). The reaction of UV/ H2O2 with natural organic matter (NOM) has intensively been studied by Sarathy et al. (Sarathy 2009). They used a system equipped with LP-UV lamps, and found that especially the hydrophobic fractions of NOM are sensitive for reaction in this system, resulting in the formation of hydrophilic products, which in general are better biodegradable. Often a reduction in aromaticity is observed, and a reduction in molecular size. It was found that UV/ H2O2 treatment leads to preferential oxidation of aromatics and conjugated double bonds, and of higher molecular size NOM accompanied by the formation of lower molecular size NOM. It was observed



that an increase in either initial H2O2 concentration or UV fluence was accompanied by the formation of more aldehydes. This probably was due to the increased •OH concentration, as this concentration increases with either increasing H2O2 concentration or with increasing UV fluence. Chu et al. (Chu 2009) studied the conversion of halogenated pesticide 2,4-dichlorophenoxyacetic acid. They found that the UV/ H2O2 process is less expensive for removing herbicide than UV irradiation alone, at the same degradation reaction conditions. By means of micro-aeration the efficiency can be further improved. Micro-aeration not only plays an important role in mixing but also increases the production rate of •OH radicals. According to Toor et al (Toor 2007), implementing a UV- H2O2 AOP as a stand alone treatment may not be practical because of the high energy costs or insufficient reductions in the formation of DBPs. Integrating UV- H2O2 with a downstream biological process may offer a promising alternative because it allows for the removal of the biodegradable intermediates formed as a result of partial oxidation of NOM. The intermediates of UV- H2O2 oxidation act as a major food source for microbial communities and are utilized as substrates in bioreactors. The combination of AOP and BAC seemed to be effective at significantly reducing DBPs at a moderate UV fluence.

2.4.3 Modelling

In the PhD thesis of Bas Wols, “Computational Fluid Dynamics” (CFD) has been applied to both ozone- and UV/H2O2 processes. Although modeling can be applied to all kinds of processes, recently much attention has been paid to modeling UV and UV/H2O2 processes. According to Cooper et al. (Cooper 2008), kinetic models for AOPs can be divided into three discrete sections, that describe the formation of the radicals (such as •OH radicals), the radical induced destruction chemistry of the contaminant of interest and fluid dynamics or the reactor type to be used. Badriyha et al. (Badriyha 2007) developed a free-radical reaction kinetic model, to predict the concentrations of all kinds of principal species (alachlor, NOM, H2O2, carbonate and intermediate radicals). The model incorporates e.g. the gradual pH decrease during the oxidation period (attributed to NOM mineralization). Pereira c.s. (Pereira 2007) studied the direct and indirect photolysis of pharmaceutically active compounds (PhACs) using a low pressure UV batch reactor. They developed UV and UV/ H2O2 photolysis models that were compared with experimental results obtained in surface water. The model predicted the experimental UV photolysis removals well but underestimated the UV/ H2O2 results, possibly caused by physical processes not accounted for, such as photosensitization induced by NOM. The two main parameters that influence the direct photolysis of a compound are the decadic molar



absorption coefficient and the quantum yield. The quantum yield represents the ratio between the total number of molecules of the compound degraded to the total number of photons absorbed by the solution due to the compound’s presence. Kinetic models, describing the photolysis and oxidation, were extensively described in literature (model (Crittenden 1999, Schwarzenbach 1993, Stefan 2004, Bielski 1985, De Laat 1999, Liao 2001). These models have to be combined with mathematical models describing the water flow through reactors (Glaze 1995, Wols 2010).

2.4.4 Byproducts of UV processes

Irradiation of water with high nitrate concentration may pose a significant health risk due to nitrite formation, according to Buchanan et al. (Buchanan 2006). The NOM in the water sample studied was not a significant contributor to the formation of nitrite during UV irradiation. According to Goldstein et al (Goldstein 2008), the level of NO2- strongly depends on pH and concentration of inorganic and organic contaminations. However, in general a UV/H2O2 process is followed by filtration over active carbon. Partly, this is done to remove the excess H2O2, but it also removes any not fully converted micro pollutants, and possible harmful compounds, like nitrite. Recently, it was also shown that application of medium pressure UV lamps may result in the formation of genotoxic compounds (Heringa, 2011).

2.4.5 Ozone in combination with Hydrogen Peroxide

During ozone decomposition significant amounts of hydrogen peroxide can be formed, according to Vandersmissen et al. (Vandersmissen 2008). The concentration of formed hydrogen peroxide is independent of the initial ozone concentration or pH, but only depends on the temperature: more H2O2 is formed at higher temperature. This H2O2 formation results either directly from O3 decomposition through O3 + •OH → O2 + HO2- or from the hydrolysis of organic ozonation products. The hydrogen peroxide formed in this way appeared to enhance the O3 decomposition rate. Indirect oxidation can be enhanced by realizing higher radical concentrations for the same ozone concentration. One possibility to increase the total radical concentration relative to ozone is the peroxone process, where H2O2 is added to an ozone containing solution. Hydrogen peroxide initiates the ozone decomposition and accelerates the decay and thus generates extra radicals. The AOP O3/ H2O2 is a common used purification treatment for ground water and surface water contaminated with refractory compounds. When ozone and peroxone processes are compared one sees an advantage for ozone, due to its electrophilic character, regarding oxidizing compounds with double bounds, iron and manganese and for its disinfection ability. The rate of direct oxidation with molecular ozone is relatively slow, except for olefins and aromatic compounds, compared to hydroxyl radical oxidation, but the concentration of ozone is of course always much larger than the hydroxyl radical concentration. As with ozone alone, pH and bicarbonate alkalinity



play a major role in the peroxone effectiveness. This role is primarily related to bicarbonate and carbonate competition. By adding hydrogen peroxide to the water, the ozone transfer from gas phase to liquid phase can be improved due to an increase in ozone reaction rates. By adding concentrations of hydrogen peroxide below 10-5 M to pure water systems, it is already possible to realize AOPs in the acidic and neutral pH range. Ozone decay rates can be increased, relative to the pure systems, by at least a factor 10 through the addition of as little as 6.0 * 10-6 M H2O2. Due to the already fast ozone decays at alkaline pH, much higher H2O2 concentrations will be needed to see an AOP effect at high pH values (Vandersmissen 2008).

2.5 Heterogeneous catalysis

Heterogeneous catalysis can be applied to improve the efficiency of AOPs. However, in drinking water treatment it is not always easy to use particles in the process. From the point of view of water treatment application, immobilized particles have the advantages of easy operation and energy saving, but show a reduced rate of reaction due to mass transfer effects. Furthermore, immobilization very often results in a reduced area available for reaction, whereas a slurry of particles has the disadvantage that the particles will have to be removed again (Coleman 2007). The formation of hydroxyl radicals from ozone can be enhanced by the presence of solid catalysts (Vandersmissen 2008). Heterogeneous catalytic ozonation is a complex process, whose underlying chemistry is not well known. Several mechanisms have been proposed for describing it that can be classified according to the kind of surface interaction proposed. Other mechanisms exclude the adsorption equilibrium and lead to models in which the rate of catalytic process does not depend on the concentration of the oxidant. Adsorption-limited kinetics seem to be more realistic considering the difficulty of adsorption encountered by organics in aqueous solutions, especially on the surface of oxides. In the case of metal oxides, heterogeneous ozone decomposition is determined by the presence of surface hydroxyl groups acting as Brönstedt acid sites. These sites also determine the charge of the surface as a function of pH, and therefore the ion-exchange behavior of the catalyst. Depending on the pH of the solution, the surface of an oxide may be charged or not. In addition to this, metal oxides have Lewis acid sites that, in an aqueous solution, allow water molecules to coordinate on their surface. The adsorption of ozone requires the displacement of coordinated water and is strongly dependent on the presence of other bases. In the case that a Lewis site is accessible to ozone, the mechanisms for its adsorption/decomposition on a catalytic surface would follow a mechanism similar to that used for explaining gas-phase decomposition: O3 → (O3)ads (O3)ads → (O)ads + O2



The interaction of the ozone molecule with an oxidized site may yield adsorbed or non-adsorbed oxygen: O3 + (O)ads → 2O2 O3 + (O)ads → O2 + (O2)ads → 2O2 In aqueous solution, the hydroxide ion is expected to act as a strong inhibitor of the adsorption ability of the catalyst by blocking Lewis acid sites. Therefore, the catalytic activity at high pH should proceed by a redox mechanism involving surface hydroxyl groups. Ozone would react with them to yield an ozone anion radical or another active species able to oxidize organic compounds either in solution or on the surface. On oxides, such as titanium dioxide, the reaction probably can be described by an interaction between Lewis acid sites and organic molecules, with an optimum mineralization rate obtained in slightly acidic conditions. Activated carbon was shown to be particularly efficient as an initiator in the decomposition reaction of ozone in the liquid phase. TiO2 seems to be an ideal photo catalyst in several ways (Hussain 2010). It is relatively cheap, highly stable from a chemical viewpoint, and easily available. Moreover, its photogenerated holes are highly oxidizing, and the photogenerated electrons are reducing enough to produce superoxide from dioxygen. According to Hussain et al. (Hussain 2010) TiO2 nanoparticles having a high specific surface area can successfully be synthesized in a vortex reactor by means of a sol-gel process with optimized operating parameters. The high surface area-to-volume ratios lead to improved effectiveness for surface-limited reactions. Substances which are readily adsorbed are degraded at faster rate, indicating that the reaction is a surface phenomenon. Another important parameter is the type and amount of surface OH groups and or physisorbed H2O. The holes can react with water to produce the highly reactive •OH radicals. These holes and the hydroxyl radicals are very powerful oxidants, which oxidize the organic materials. Hussain et al. found that TiO2 promotes ambient temperature oxidation of the major classes of indoor air pollutants, and does not need any chemical additives. According to Hussain et al., it has been demonstrated that organics can be oxidized to CO2, water and simple mineral acids at a low temperature on TiO2 catalysts in the presence of UV or near-UV illumination. The photo catalytic activity of TiO2 is greatly influenced by the crystalline form (anatase or rutile), and the size (actually the specific surface area) of the particles. This of course also is important for water purification processes, based on interactions with TiO2, which have been described lately. According to Yu et al. (Yu 2010), TiO2 has been applied as a promising environmentally friendly photo catalyst in many fields such as environmental remediation, hydrogen production and solar energy utilization. It is valued for its chemical stability, lack of toxicity and low cost.



One benefit of TiO2 photo catalysis (Murray 2006) is that the catalyst can be regenerated using UV light, and is therefore reusable. The regeneration process is similar to the process applied to zeolites, used in adsorption processes (Koryabkina 2007). In order to effectively photo catalytically degrade an azo (R-N=N-R’)dye under solar irradiation, Yu et al. (Yu 2010) used anatase that was co-doped with cerium and nitrogen (Ti1-xCexO1-yNy) nanoparticles. These were synthesized using a one-step technique with a modified sol-gel process. The enhanced photo catalytic degradation was attributed to the increased number of photogenerated •OH radicals. Yu et al. applied TiO2 nanoparticles in the field of organic pollutant removal from wastewater. These applications have been limited by the large energy band gap (3.2 eV), which can capture only less than 3% of the available solar energy (λ< 387 nm) as well as by the fast recombination of photogenerated electron-hole (e- - h+) pairs, both on the surface and in the core of TiO2 NPS. Photo catalysis that functions in the visible wavelengths (400nm < λ < 800 nm) is desirable from the viewpoint of solar energy utilization. In order to improve the properties of TiO2, TiO2 particles are doped with lanthanide, Cerium or nitrogen. It was found that cerium and nitrogen co-doped anatase can be prepared, resulting in a synergistic effect, effectively inhibiting the recombination of photogenerated electrons and holes. A process that can be applied to obtain effective TiO2 particles is the “sol-gel process” (Murray 2006). Here, a thin film of nano crystalline TiO2 is deposited onto a support substrate. The process typically involves the transformation of a titanium salt (usually an alkoxide) in an adequate solvent and/or under acid/base conditions. The combined processes of hydrolysis, polycondensation and drying of a colloidal suspension result in the formation of a ceramic coating. Hydrolysis of the metal salt alkoxide is carried out to convert the titanium into the anatase form of TiO2.

Murray et al. (Murray 2007) studied the adsorption of NOM onto pelletized TiO2 and the oxidation of the pellet surface by UV light, in order to reduce the dissolved organic carbon concentration of source water. The authors describe how an excess of NOM, e.g. during heavy rainfall, can be dealt with by adding extra coagulant, containing trivalent metal ions. Coagulants, such as ferric sulfate, can achieve a significant reduction in both ultraviolet absorbance at 254 nm and dissolved organic carbon, but often at high doses, and generate solid residuals needing further treatment and disposal. Titanium dioxide can be applied as a powder or immobilized as a thin film for photo catalytic purposes. The problem with drinking water applications would be that an energy intensive filtration stage would be required to separate and recover the TiO2 powder while thin films may have mass transfer limitations. Murray et al. used pellets of TiO2 for the adsorption of NOM and side-stream UV irradiation to bring about NOM oxidation and regeneration of the pellets. They found that the pellets preferentially adsorb molecules with high molecular weights. These pellets probably will not offer



a single stage process for water purification, but may be incorporated into the flow-sheet as a pre-treatment to coagulation. Ikehata et al. (Ikehata 2006) describe several AOPs for the degradation of pharmaceuticals, including the complete transformation of 15 mg/L buspirone by TiO2/hν treatment, using a 1500 W xenon lamp with a 340 nm cut-off filter at 50 ºC. Coleman et al. (Coleman 2007) studied the effect of several types of TiO2 on the 1,4-dioxane removal. They found that TiO2 photo catalysis with both UVA and solar light is effective in degrading 1,4-dioxane, and is much more efficient than the H2O2/UVC process, UVA or UVC radiation alone. Very often, studies were carried out using a TiO2 slurry reactor system. The advantages of a slurry type reactor are the large surface area of catalyst and the intimate contact between the target compounds and the suspended particles, reducing mass transfer effects inherent in an immobilized system. The disadvantages are inhibition of light transmittance by the catalyst and the difficulties encountered when attempting to recover the particles from the treated effluent, due to the need for a solid-liquid separation process which is both time and energy consuming. The authors tried to improve this separation, by applying insulated magnetic core particles coated with a layer of photoactive TiO2 (Magnetic Photocatalyst, MPC). The magnetic core allows for increased ease of separation of the particles from the treated effluent whereby the particles can easily be removed by the application of a magnetic field. The magnetic particles are coated with an initial layer of SiO2, in order to prevent the interaction between the magnetic cores and the TiO2, thus preventing the photo dissolution of the cores during dissolution. Subsequently, a layer of TiO2 was applied. This process, however, results in a larger particle size and density, which gives a lower available particle surface per unit mass for reaction. The MPC particles, which are considerably larger than the more commonly used P25 TiO2 particles (P25 is a product of Degussa, consisting of TiO2 particles with a diameter of about 30 nm (Caris 1990), give different light scattering properties, thus resulting in a different optimum catalyst loading. Photo catalysis using the sol-gel reactor was slower compared to that obtained in the P25 suspension reactor, due to the mass transfer limitation of an immobilized system. It is known that H2O2 can enhance the reaction by providing additional hydroxyl radicals either through trapping of photogenerated electrons and/ or photolysis of H2O2. Adding larger concentrations of H2O2 may even result in a substantial decrease of the reaction rate, on the one hand because of the formation of less reactive hydroperoxyl radicals, and on the other hand due to competition of H2O2 with the contaminants for conduction band electrons. The reaction mechanisms involved are shown below. TiO2–UV → h+ +e−

H2O2 +e−→ OH− +OH•

H2O2 ←→ 2OH• hv,λ<300nm



The effect of the addition of H2O2 to the MPC system was found to be quite different. Coleman et al. found that the mineralization rate of 1,4-dioxane increases as the concentration of H2O2 increased up to 30 ppm H2O2. This probably can be attributed to the low activity of the MPC compared to P25 in generating hydroxyl radicals. Because of this, addition of H2O2 will cause the rates to increase due to the increased number of hydroxyl radicals present for reaction. •OH radicals can be generated when H2O2 traps photogenerated electrons. This trapping would also help suppress recombination of electron-hole pairs produced at the activated catalyst surface. Beyond the concentration of 30 ppm H2O2, no enhancement to the degradation process could be observed, and actually a decrease in the mineralization rate could be observed when the H2O2 concentration was increased beyond 60 ppm. This may be due to the racial scavenger properties of H2O2. Murray et al. (Murray 2006) used a sol-gel system with immobilized TiO2 and a UVA lamp with an emission wavelength of 365 nm. Compared with a slurry reactor the removals obtained were low, which can be explained from the fact that there only was a small surface area of coated slide being placed in an excess of humic acid. The actual photo catalytic rate seemed not to depend on the specific surface area of the catalyst, but on the available or active sites involved in the adsorption and subsequent oxidation process. Variations in the crystalline phase of TiO2 present in the thin films could also contribute to the variation in photo catalytic efficiency observed between different coatings. Anatase is more photoactive than the rutile phase (Murray 2006). This has been ascribed to the Fermi level of anatase being higher than that of rutile (Hussain 2010). The temperature at which the thin films were calcined affects which crystalline phase of TiO2 dominates. Furthermore, thin films seem to be more efficient than relatively thick films. Watts et al. (Watts 2008) applied a UV-TiO2- O3 process to wastewater containing anti-foaming agents and flame retardants (tri-n-butyl phosphate and tris(2-chloroethyl)phosphate, and found that the biodegradable fraction of DOC thus was increased.

2.6 Ultrasonication

Suri et al. (Suri 2007) studied the use of ultrasound to destroy estrogen compounds in water. The high acoustic energy generates physical and chemical reactions that can degrade organic chemicals present in the liquid. These reactions result from the creation and violent collapse of cavitation bubbles. These cavitation bubbles, produced acoustically in a matter of microseconds upon implosion result in extreme conditions (5000 K and 500 bar) in the gaseous phase at microscopic points in the solution. Cavitation produces high mechanical shear stresses that are exerted on the substance in the liquid. Thermal breakdown of volatile substances occurs in the gaseous phase and in the interfacial region. Sonochemical reactions are also caused by the generation of highly reactive radicals, such as •OH radicals, which cause chemical transformations in the bulk solution. This sonication process can be influenced by several parameters, such as ultrasound power density, power



intensity and reactor configuration. The type of reactor used can influence the imparted energy to the solution and, hence, the power density. Some of the factors which influence the power output from the ultrasound include solution viscosity, surface tension, vapor pressure, suspended solids, pressure, temperature, ultrasound frequency, power of the sonicator, and size and type of the reactor vessel. The authors showed that degradation kinetics and energy efficiency depend on the reactor configuration and power characteristics such as power intensity and power density. The estrogen degradation rates were observed to increase with increasing power intensity. A higher power intensity results in a higher pressure, causing a more complete implosion of the cavities (and thus higher reactivity), and more cavities in the solution. The choice of reactor and ultrasound power are important to achieve both optimized kinetics as well as energy efficiency. Reactors with high ultrasound intensity and low power density would be favorable for cost effective destruction of pharmaceutical compounds in water.

2.7 The Fenton process

The principle of the Fenton process is the catalytic cycle of the reaction between iron (catalyst) and hydrogen peroxide (oxidant) to produce hydroxyl radicals (IJpelaar et al., 2000b). The hydroxyl radical is produced according to the following reaction:

Fe2+ + H2O2 → Fe3+ + OH- + ●OH The system has its maximum catalytic effect at pH 3. At higher pH, competition between the reduction of ferric to ferrous iron and the precipitation of ferric iron as hydroxide occurs and reduces the catalytic cycle efficiency (IJpelaar et al., 2001). Moreover, at pH values above 5, it is possible that the Fenton reaction produces another intermediate than the OH-radical, which is assumed to be the Ferry species. This intermediate is more selective than OH-radicals and therefore not equally efficient towards oxidation of different organic micro pollutants and in general organic compounds (Hug and Leupin, 2003). Nevertheless, relevant organic pollutant degradation can be achieved by the ferrous iron concentration present in the groundwater and at ambient pH. At pH higher than 4 the single production of hydroxyl radicals may be sufficient to degrade the target compounds (IJpelaar et al., 2001). To enable an optimal production of radicals by reaction between Fe2+ and H2O2, the presence of oxygen should be prevented in order to prevent formation of iron(hydr)oxides, which take away the iron from the catalytic cycle. Thus, anaerobic water is required (IJpelaar et al., 2001).



2.8 Summary of literature review.

A summary of the literature review shown above is given in Table 2.1

Table 2.1: summary of literature review on AOPs

Subject Literature Application of AOP Cooper 2008

Baus 2007 Application of AOP; EDC Badriyha 2007 Application of AOP; organic micropollutants Wang 2009 Application of AOP; pharmaceuticals Ikehata 2006

Ikehata 2008 Application of AOP; pesticides Ikehata 2008 Fenton IJpelaar 2001

IJpelaar 2001b Heterogenous catalysis Coleman 2007

Hussain 2010 Yu 2010 Murray 2006 Murray 2007 Koryabkina 2007 Ikehata 2006 Watts 2008

modeling Badriyha 2007 Cooper 2008 Crittenden 1999 Bielski 1985 Schwarzenbach 1993 Stefan 2004 De Laat 1999 Glaze 1995 Liao 2001 Wols 2010

Ozone processes Ried 2006 Song 2009 Peyton 1988 Sonntag 2007 Kornmueller 2007 Rui 2006

Photolysis

Lee 2008 Rosario-Oritiz 2010

Ultrasonisation Suri 2007

UV/H2O2 Oppenländer 2002 Kwon 2009



UV/H2O2; aldehydes Jo 2009 UV/H2O2; byproducts Buchanan 2007

Goldstein 2008 Heringa 2010

UV/H2O2; disinfection Buchanan 2007 Mamane 2007

UV/H2O2; EDC Linden 2007 UV/H2O2; MTBE Tawabini 2009 UV/H2O2; naproxen Felis 2007 UV/H2O2; NDMA Plumlee 2008 UV/H2O2; NOM Murray 2006

Sarathy 2007 Pereira 2007 Sarathy 2009 Toor 2007 Badriyha 2007

UV/H2O2; organic micropollutants, HAA’s Nienow 2009 UV/H2O2; pesticides Chu 2009 UV/H2O2; pharmaceuticals Rosario-Oritiz 2010 UV/H2O2; pretreatment Li 2008 UV/H2O2; radical scavengers Rodriguez 2008 UV/H2O2; tertiairy Butyl alcohol Li 2008



3 Pilot experiments with emerging substances

3.1 Introduction

According to the recent findings described in literature (chapter 2) the UV/H2O2 process is very effective for the conversion of many emerging substances. In this research we tested whether the UV/H2O2 process can effectively be applied to convert a series of emerging contaminants. We also determined the Electrical Energy per Order (EEO; the amount of electrical energy required to obtain a conversion of 90% for a certain compound). In this way the effectivity of the UV/H2O2 process for all compounds can be compared.

3.2 Emerging substances A series of emerging substances was tested in the pilot experiments: Alachlor Atrazine Atrazine-desethyl Desisopropylatrazine Carbamazepine Sulfamethoxazole Para-chlorobenzoic acid (pCBA) MTBE pCBA often is used to study the effectiveness of radical formation in a UV/H2O2 system, as pCBA mainly is degraded by •OH radicals.

3.3 Experimental section Reagents Lab grade reagents were used: H2O2 (30%) and Na2SO3 A mixture of Alachlor, Atrazine, Atrazine-desethyl, Desisopropylatrazine, Carbamazepine, Sulfamethoxazole, Para-chlorobenzoic acid (pCBA), and MTBE was prepared in milliQ. The concentrations are shown in the appendix. The mixture was spiked to a stainless steel tank containing the water with a final concentration of 2 µg/L Desisopropylatrazine, Carbamazepine and Sulfamethoxazole, 5 µg/L Alachlor, Atrazine, Atrazine-desethyl and MTBE and 10 µg/L pCBA.



Test water UV Pilot tests were conducted in pretreated (coagulation, sedimentation, micro sieves and sand filtration) surface water from the river Meuse. The water quality during the different tests is shown in Table 1.

Table 3.1: Water quality during the various tests

Date pH HCO3 NPOC ÙV-T254nm mg/L mg C/L %

19-1-2010 7,8 185 4,1 75,5 21-1-2010 7,8 180 4,2 75,2

1 Non Purgable Organic Carbon Analyses Details of the analytical methods applied are shown in appendix . Flow through experiments; KWR pilot reactor The stainless steel pilot UV unit (1-5 m3/h) was equipped with four quartz sleeves perpendicular to the flow direction. The reactor can be equipped with LP lamps or MP lamps supplied by Philips Lighting BV (Roosendaal, the Netherlands). During the experiments the reactor was equipped with 4 MP lamps which were dimmed to 10 %. Lamp output and water flow were monitored frequently. The lamp intensity was measured with a sensor (MUV2.4WR UV-referenzradiometer with ‘SUV20.2A2Y1R/150/UVD6(RO001’ sensor, IL Metronic Sensortechnik GmbH) which was on the reactor wall of the UV reactor pointing towards one of the lamps (see Figure 1). Contaminants were spiked in a tank and stirred during the tests. H2O2 was dosed shortly before the pump that fed the water to the UV/H2O2-pilot reactor. By adjustment of the valves (1 and 5 m3) positioned after the pump the flow, during all the tests, was set to about 1.1 m3/h. Before and after the UV reactor static mixers were installed to establish adequate mixing prior to the reactor and the sampling points. The influent and effluent sampling points were positioned directly after the static mixers (see Figure 1 ).



Ø125

300

85

85

30

30

30

30

150

Stirred spiked

test water

Static mixer

Sample pointinfluent

Static mixer

Overflow to waste

H2O

2

Sensor

Sample pointeffluent

UV reactor

Lampposition

F

Tests were conducted on 2 different days with an applied H2O2 dose of 0 or 10 mg/L. Every day a fresh mixture of micropollutants was added to the tank with water prior to addition. During all tests 1 influent and effluent sample was taken. Samples taken for the tests with H2O2 were directly quenched with 300 mg/L sodium sulfite, after sampling to neutralize the residual H2O2.

Figure 1 Schematic depiction of the pilot UV test installation at KWR.

Figure 2: Schematic overview of the KWR pilot reactor. This reactor can be equipped with either four LP or four MP UV lamps.



3.4 Results and discussion Detailed results are shown in Appendix I. The UV dose distribution applied in the reactor was calculated according to CFD calculations. An extensive description of this method and its application to the KWR pilot reactor can be found in the PhD thesis of B. Wols (Wols 2010). The results are shown in figure 3.

From these data it was calculated that the mean UV dose in this reactor then was ca. 600 mJ/cm2. The experiments were carried out at a flow of 1.05 m3/h, and a lamp power of 0.804 kW. The degradation of the emerging substances at a UV-dose of ca. 600 mJ/cm2 is shown in Appendix II and graphically in figure 4.

Figure 3: CFD dose calculation for the KWR pilot reactor at a flow of 1.05 m3/h



0

10

20

30

40

50

60

70

80

90

100

Alachloor

atrazine

atrazine-desethyl

desisopropylatrazine

carbamazepine

sulfamethoxazool

pCBA

MTBE

degradation (%)

0 mg H2O2

10 mg H2O2

From these results it can be concluded that, in general, addition of 10 mg/L H2O2 results in a better degradation of the compounds. For some compounds, especially carbamazepine, alachlor, atrazine desisopropoyl atrazine, pCBA and MTBE the improvement was large. However, even with H2O2 the conversion of MTBE appeared not to exceed 35% in this experiment. This is in accordance with literature data (Baus 2007; also see chapter 2). The UV/H2O2 method in general seems to be an effective method to convert this series of emerging substances, except for atrazine-desethyl and MTBE, as for these compounds the conversion appeared to be less than 50%. This conversion probably can be increased by increasing the UV dose, but this means that a very high amount of energy is required. Increasing the H2O2 concentration is less effective to obtain a higher conversion. Although UV/H2O2 is an effective technique to convert micro pollutants into more biodegradable compounds, the disadvantage is that energy demand of the technique may be rather high. This energy demand, in a certain reactor, is described by the “Electrical Energy per Order” (EEO). This is the amount of energy, required to convert 90% of a certain compound in a certain reactor:

f

iEO

c

cF

PE

lg*

=

Figure 4: Conversion of micro pollutants by means of UV or UV/H2O2



In this formula, P is the electrical lamp power (kW), F the water flow (m3/h), C0 the concentration of a contaminant at the beginning of the process, and Ct the end concentration of this contaminant. EEO thus has the unit kWh/m3. It is specific for a certain compound. As reactor design and the flow profile may be different for various reactors, the EEO also may differ for various reactors. The EEO was calculated for this series of compounds in the KWR pilot reactor, as shown in figure 5.

0

1

2

3

4

5

Alachloor

atrazine

atrazine-desethyl


carbamazepine

sulfamethoxazool

pCBA

MTBE

EEO (kWh\m

3)

0 mg H2O2

10 mg H2O2

Figure 5: EEO for micro pollutants in the KWR pilot reactor

From this figure it can be concluded that in general addition of H2O2 results in a lower energy demand for obtaining a similar degradation result. As expected, for atrazine-desethyl and MTBE the EEO is rather high, as their conversion was relatively low. This means that relatively much energy will be required to convert 90% of these compounds. For the other compounds an average EEO of about 1 kWh/m3 is found. The results were compared with previous results, obtained in 2006 at a calculated UV dose of 633 mJ/cm2, as shown in figure 6 (Beerendonk 2009, Hofman 2009)



Remark:. The dose calculated by CFD for the situation in 2006 (as described in literature, Wols 2010) was a little higher than in the present situation. This is caused by a higher UV Transmission of the water used in 2006, and by a different flow profile entering the UV reactor in the set-up used in 2006. The influence of UV Transmission of the water is rather high, and changes as a function of the seasons. For application of UV/H2O2 processes for the conversion of organic micropollutants this has to be taken into account.

0

1

2

3

4

5

Alachloor

atrazine

atrazine-desethyl


carbamazepine

sulfamethoxazool

pCBA

MTBE

EEO (kWh\m

3)

Dunea 2010(600/10)

Dunea 2006(633/10)

Figure 6: Comparison of EEO at two different UV doses.

It seems that the EEO for the process in 2010 was a little higher than for the process in 2006. Both experiments were carried with a similar experimental setting, but as in 2006 the UV-T of the water happened to be higher, the mean UV dose was higher, or in other words, the process was more effective. This can also be seen from the EEO: the EEO of the process in 2006 seems to be lower than the EEO in 2010. This example illustrates the importance of the efficiency of the process. The UV-T of the influent cannot always be influenced, but using an optimized reactor design will certainly help in rendering the process more efficient and less expensive.



4 Conclusions

Advanced oxidation processes can be very effective for the conversion of emerging substances in water. Which type of AOP will be most efficient, strongly depends on the circumstances, e.g. the UV transmission, salt concentration, pH, oxygen content etc. In this investigation we studied the UV/H2O2 –process. The conversion of a set of emerging substances in a pilot reactor, equipped with 4 medium pressure mercury lamps was measured. The effect of adding H2O2 on the conversion of these substances was studied, and the Electrical Energy per Order was calculated. The UV/H2O2 method in general seems to be an effective method to convert this series of emerging substances, except for atrazine-desethyl and MTBE, as for these compounds the conversion appeared to be less than 50%. For carbamazepine, alachlor, atrazine desisopropoyl atrazine, pCBA and MTBE it was found that addition of H2O2 resulted in a strongly enhanced conversion. These data are in good accordance with literature. According to Tuhkanen (Tuhkanen, 2004), the EEO of a UV/H2O2 process is 0.5-2 kWh/m3. Our results indicate that the EEO in our pilot reactor is about 1 kWh/m3. However, it still means that this technique requires a large amount of electrical energy, and thus is relatively expensive. On the other hand, as also indicated by our results, the energy demand also strongly depends on the circumstances: in water with a high UV transmission, a lower EEO will be observed. Furthermore, our pilot plant has not been optimized. Using CFD modeling, it will be possible to construct a more efficient reactor, and thus also the EEO may be lowered.



5 Literature

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I Appendix

Analytical methods applied UV scan/UV254nm; The UV scan is performed with the Thermo Spectronic Unicam UV500 spectrophotometer at a wavelength of 254nm at a path length of 1 cm. pH; Prescription LAM-043, conform NEN 6411 (1981). The pH is measured with a PHM 83 autocal pH meter of Radiometer Copenhagen Peroxide; Prescription LAM-0048, conform KWR prescription 1-06-1 (1995) The peroxide analyses is performed with the Thermo Spectronic Unicam UV500 spectrophotometer at a wavelength of 400nm at a path length of 10 cm NPOC; prescription LAM-041 conform ISO 8245 and NEN-EN 1484. NPOC is analyzed using Infra Red gas. The analyzer is a Shimadzu TOC-5000A with TOC-control and an ASI-5000A auto sampler. Hydrogen carbonate; prescription LAM-042, conform NEN 6531 en 6532 Hydrogen carbonate is titrated up to a pH of 4,35. The pH is measured using a PHM 83 autocal pH meter of Radiometer Copenhagen Alachlor, Atrazine, Atrazine-desethyl are analyzed using GCMS (LOA protocol Scr.GCMS.SPE target compounds) Desisopropylatrazine, Carbamazepine and Sulfamethoxazole prescription LOA-600 using online LC/MS/MS (Orbitrap; Thermo TSQ-7000 triple-Qud LC-MS/MS). pCBA; prescription LOA-007 using a SPE online HPLC-UV (Waters 996 diode array detector). MTBE; prescription LOA-405 using purge and trap GC/MS



II Appendix

The water quality of the Dunea water from Bergambacht was tested:

water quality water after rapid sand filtration ps. Bergambacht (Dunea)

Date pH HCO3 NPOC ÙV-T254nm mg/L mg C/L % 19-1-2010 7,8 185 4,1 75,5 21-1-2010 7,8 180 4,2 75,2

Furthermore, the actual water flow through the reactor and the H2O2 concentration were measured:

flow H2O2

required measured measured required

measured influent influent effluent

Date test

m3/h m3/h mg/l mg/l mg/l

19-1-2010 1 1,1 1,05 0 0,24

19-1-2010 2 1,1 1,05 10 10,8 9,6

21-1-2010 3 1,1 1,05 0 0,24

21-1-2010 4 1,1 1,05 10 10,9 10,2 Applying a flow of 1.05 m3/h and a UV-T of 75% the following results were obtained by means of CFD calculations for this reactor: Dmin 249,2 mJ/cm2

D10 319,2 mJ/cm2 D50 477,7 mJ/cm2 Dmean 600,5 mJ/cm2 D50/Dmin 1,92 Dmean/Dmin 2,41 D50/D10 1,50 Dmean/D10 1,88 For an “ ideal” UV reactor, Dmean/D10 would be “1”. The closer this value approaches “1”, the more optimal the reactor will be. The results of the UV/H2O2 tests are shown in table 4.

Table 5.1: water quality used

Table 5.2: flow through the pilot reactor and H2O2 concentration during experiments



Compound H2O2

0 mg/L H2O2 10 mg/L

Influent effluent Conversion (%)

Influent effluent Conversion (%)

4,53 2,06 54,5 3,01 0,44 85,4 Alachlor

4,50 2,19 51,3 2,84 0,64 77,5 4,32 2,03 53,0 3,76 1,09 71,0

atrazine 4,33 2,11 51,3 3,74 1,27 66, 0 4,40 2,49 43,4 4,03 2,19 45,7

atrazine-desethyl 4,53 2,59 42,8 4,18 2,32 44,5 2,13 1,00 53,1 1,56 0,53 66,0

desisopropylatrazine 2,15 1,08 49,8 1,63 0,59 63,8 1,90 1,21 36,3 1,97 0,26 86,8

carbamazepine 2,23 1,25 43,9 1,81 0,28 84,5 1,85 0,212 88,6 1,69 0,05 97,0

sulfamethoxazole 1,83 0,212 88,5 1,81 0,06 96,7 0.27 5.85 37.1 8.33 1.29 8.45

pCBA1 8.68 6.65 34.9 7.54 1.26 8.33 5.05 5.18 -2.6 5.21 3.28 37.0

MTBE 4.81 4.35 9.6 4.41 3.06 30.6

1) The lower detection limit was temporarily lowered from 5 to 1 µg/L. Due to analytical problems the samples could only be analyzed after 14 days, whereas the official expiry date is 7 days. The samples had been stored in a dark room at 4

0C,

and the results obtained seemed to be reliable. 2) For these results the sulfamethoxazole concentration has been corrected for the recovery of sulfamethoxazole (4 resp. 6%). As a result, the deviation for these concentrations can be higher than usual. In tables 5-7 the EEO data of the compounds are shown.

Table 5.3: Conversion of micro pollutants in the KWR pilot reactor



ci cf

EEO

EEO MP-lamp dose 600 mJ/cm2 and 0 mg H2O2/l Flow 1.05 m3/h Power 0.804 kW (µg/l) (µg/l)

Alachlor 4,52 2,13 2,34 atrazine 4,33 2,07 2,39 atrazine-desethyl 4,47 2,54 3,13 desisopropylatrazine 2,14 1,04 2,44 carbamazepine 2,07 1,23 3,40 sulfamethoxazole 1,84 0,21 0,81 pCBA 8,98 5,74 3,94 MTBE 4,93 4,77 51,79

ci

cf

EEO

EEO MP-lamp dose600 mJ/cm2 and 10 mg H2O2/L Flow 1.05 m3/h Power 0.804 kW

(µg/l) (µg/l)

Alachlor 2,93 0,54 1,04 atrazine 3,75 1,18 1,52 atrazine-desethyl 4,11 2,26 2,94 desisopropylatrazine 1,60 0,56 1,68 carbamazepine 1,89 0,27 0,91 sulfamethoxazole 1,75 0,06 0,51 pCBA 7,94 1,28 0,96 MTBE 4,81 3,17 4,23

Table 5.4: EEO for the KWR pilot reactor at 600 mJ/cm2 and 0 mg H2O2/L

Table 5.5: EEO for the KWR pilot reactor at 600 mJ/cm2 and 10 mg H2O2/L:



ci

cf

EEO

EEO MP-lamp, dose 450 mJ/cm2 and 10 mg H2O2/L Flow 1.35 m3/h Power 0.804 kW

(µg/l) (µg/l)

Alachlor 1,60 0,23 0,71 atrazine 1,60 0,44 1,06 atrazine-desethyl desisopropylatrazine carbamazepine 1,90 0,32 0,77 sulfamethoxazole 1,20 0,09 0,53 pCBA MTBE 3,00 1,40 1,80

Table 5.6: EEO for the KWR pilot reactor at a dose of 450 mJ/cm2 and 10 mg H2O2/L

degradation of priority compounds by uv and uv- oxidation · “advanced oxidation techniques”...

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