Universidad de Cantabria Escuela Técnica Superior de Ingenieros Industriales y de Telecomunicación (ETSIIT) Departamento de Ingenierías Química y Biomolecular
“Assessment of polychlorinated dibenzo-p-dioxins and dibenzofurans, PCDD/Fs, in the application of advanced oxidation processes”
(Evaluación de dibenzo-p-dioxinas y dibenzofuranos policlorados, PCDD/Fs, en la aplicación de procesos de oxidación avanzada)
Memoria de Tesis Doctoral presentada para optar al título de Doctora por la Universidad de Cantabria. Doctorado en Ingeniería Química y de Procesos.
Marta Vallejo Montes
Supervisors:
Prof. Dr. Inmaculada Ortiz Uribe Dr. Mª Fresnedo San Román San Emeterio
Santander, Julio 2014.
Programa Oficial de Doctorado en Ingeniería Química y de Procesos (BOE núm. 36, de 10 de febrero de 2010. RUCT: 5311209) con Mención hacia la Excelencia (BOE núm. 253, de 20 de Octubre de 2011. Referencia: MEE2011-0031)
The research described in this thesis has been carried out at the Advanced Separation Processes Research group in the Chemical and Biomolecular Engineering Department at the University of Cantabria. Financial support from the projects CTQ2011-25262, CTQ2008-05545 and CTQ2008-00690 (Ministerio de Economía y Competitividad- MINECO/SPAIN and Fondo Europeo de Desarrollo Regional- FEDER) is gratefully acknowledged. Marta Vallejo is indebted to the “Ministerio de Economía y Competitividad” for the FPI fellowship BES-2009-025363.
“Assessment of polychlorinated dibenzo-p-dioxins and dibenzofurans, PCDD/Fs, in the application of advanced oxidation processes”
Marta Vallejo Montes
Summary
Chlorophenols (CPs) are a family of organic compounds, some of them listed as priority pollutants, characterized by their toxicity and persistence in the environment. As consequence of their extensive and long-term use in industry and daily life, CPs have been found ubiquitously in the environment, including soils, surface and groundwaters and industrial wastewaters. Advanced oxidation processes (AOPs), which are based on the formation of very active oxidant species, mainly hydroxyl radicals (OH), have been successfully applied to the treatment of wastewaters containing recalcitrant organic compounds such as chlorinated phenols. However, the partial mineralization of organic contaminants during their treatment may lead to the formation of reaction byproducts that could be more harmful than their parent compounds. On this way, CPs are among the most important and direct precursors of the formation of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs). PCDD/Fs are a family of unintentionally persistent organic pollutants regulated internationally by the Stockholm Convention (2001) and characterized by their persistence, bioaccumulative behavior and various degrees of inherent toxicity.
The potential formation of toxic byproducts, especially if the highly toxic PCDD/Fs are produced, is therefore an important issue to take into account in the assessment of treatment technologies. To date, a thorough understanding of the potential formation of PCDD/Fs during the application of AOPs for the abatement of chlorinated organic pollutants is lacking. In this regard, this thesis aims to assess the reaction intermediate products, focusing on the potential formation of PCDD/Fs, in the advanced oxidation treatment of aqueous solutions containing chlorinated organic compounds, studying the relevant role of the operating conditions. Concretely two broadly applied AOPs, namely Fenton and electrochemical oxidation, have been studied.
Despite the high effectiveness of electrochemical oxidation in the oxidation of recalcitrant compounds, most studies have focused on the degradation of major contaminants without considering the reaction transformation products. As such, the distribution of intermediate products, with special attention to the potential formation of PCDD/Fs, in the electrochemical oxidation of 2- chlorophenol (2-CP) solutions using boron doped diamond anodes was reported in chapter 3.1. The role of two commonly applied electrolytes, namely NaCl and
I
Na2SO4, was studied. Results depicted that at times when near complete mineralization was achieved, the use of NaCl resulted in a far higher formation of PCDD/Fs than Na2SO4, emphasizing the importance of the concomitant presence of chloride ions and PCDD/Fs precursors in the reaction medium.
In chapter 3.2 the performance of the Fenton oxidation of 2-CP was assessed in terms of the reaction intermediate products, highlighting the possible formation of PCDD/Fs. The effect of H2O2 dose, which constitutes one important limitation in the process economy, and the working temperature was evaluated. In addition, since chloride may be one of the products of 2-CP degradation and it is frequently found in a wide variety of wastewater matrices, the presence of chloride ions in the reaction medium was studied. The obtained results showed that sub-stoichiometric amounts of H2O2 led to the formation of PCDD/Fs, which was accentuated when chloride was present in the reaction medium.
In a further step, chapter 3.3 focuses on the potential formation of PCDD/Fs when electrochemical and Fenton oxidation were applied to the remediation of real wastewaters, particularly the leachates from a municipal solid waste landfill, characterized by a concomitant presence of high concentrations of organic matter and chloride ions. Although a high effectiveness in the degradation of the organic matter content was reported with both technologies, they depicted different behavior in the oxidation of PCDD/Fs. Whereas the concentration of PCDD/Fs in the leachate samples was reduced as a result of their electrochemical treatment, the opposite trend, an increase in PCDD/Fs concentration, was quantified in the Fenton oxidized samples.
This thesis reports novel results dealing with the potential formation of PCDD/Fs in the oxidation treatment of aqueous solutions containing chlorinated organic pollutants, emphasizing the importance of transformation byproducts assessment as well as the correct selection of operating conditions in extending the understanding of the overall effectiveness of AOPs.
IV Resumen
Los clorofenoles (CPs) son una familia de compuestos orgánicos, algunos de los cuales se encuentran listados como contaminantes prioritarios, caracterizados por su toxicidad y persistencia en el medio ambiente. Como consecuencia de su uso extensivo durante largos periodos de tiempo tanto en la industria como en la vida diaria, los CPs se han encontrado ampliamente en el medio ambiente, incluyendo suelos, aguas superficiales y aguas subterráneas, además de en aguas residuales industriales. Los procesos de oxidación avanzada (AOPs), basados en la formación de especies oxidantes, principalmente de radicales hidroxilo (OH), se han aplicado satisfactoriamente en el tratamiento de aguas residuales contaminadas con compuestos orgánicos recalcitrantes como los clorofenoles. Sin embargo, en ocasiones la incompleta mineralización de la materia orgánica, durante los tratamientos de oxidación avanzada de aguas residuales, puede conducir a la formación de intermedios o subproductos de reacción que pueden presentar mayor toxicidad que los contaminantes de partida.
En este sentido, los clorofenoles se encuentran entre los más importantes y directos precursores de la formación de dibenzo-p-dioxinas y dibenzofuranos policlorados (PCDD/Fs). Los PCDD/Fs son una familia de contaminantes orgánicos persistentes regulados internacionalmente por el Convenio de Estocolmo (2001). Los PCDD/Fs se caracterizan por su persistencia y su carácter lipofílico, lo que hace que tiendan a bioacumularse a través de la cadena alimentaria, y por su inherente toxicidad. Es por ello que la evaluación de la formación de posibles subproductos tóxicos, como es el caso de los PCDD/Fs, es un punto muy importante a considerar en la evaluación de las tecnologías para el tratamiento de aguas residuales. Hasta la fecha, la potencial formación de PCDD/Fs durante la aplicación de AOPs para el tratamiento de compuestos orgánicos clorados no ha sido evaluada en profundidad. En este sentido, la Tesis Doctoral aquí presentada tiene por objetivo la evaluación de los productos de reacción, con especial interés en la posible formación de PCDD/Fs, durante la aplicación de AOPs al tratamiento de soluciones acuosas que contienen compuestos organoclorados. Concretamente dos procesos ampliamente utilizados, la oxidación electroquímica y la oxidación Fenton han sido estudiados.
III A pesar de la alta eficacia de la oxidación electroquímica en el tratamiento de compuestos recalcitrantes, la mayoría de los trabajos se centran en el estudio de los contaminantes mayoritarios sin tener en cuenta los productos de reacción. Es este sentido, su evaluación, prestando especial interés en la posible formación de PCDD/Fs, durante la oxidación electroquímica de disoluciones acuosas de 2-clorofenol (2-CP), utilizando ánodos de diamante dopado con boro se llevó a cabo en el capítulo 3.1, estudiando la influencia del electrolito. Para ello se seleccionaron los electrolitos NaCl y Na2SO4, comúnmente utilizados debido a su extendida presencia en gran variedad de matrices acuosas. Los resultados obtenidos mostraron que para tiempos de reacción en los que prácticamente el COT se mineralizó por completo, el empleo de NaCl dio lugar a una considerable formación de PCDD/Fs en comparación con el Na2SO4, subrayando la importancia de la presencia simultánea de NaCl y precursores de PCDD/Fs en el medio de reacción.
Un paso más en el desarrollo de esta investigación se llevó a cabo en el capítulo 3.2., donde se estudió la influencia de las variables de operación: dosis de H2O2, factor limitante en la economía del proceso, y temperatura, así como la presencia de cloruro en el medio de reacción, en la oxidación Fenton de disoluciones acuosas de 2-CP, evaluando la formación de PCDD/Fs como posibles subproductos de reacción. Los resultados obtenidos mostraron que cantidades subestequiométricas de H2O2 condujeron a la formación de PCDD/Fs, lo que se vio acentuado cuando se añadió cloruro al medio de reacción.
Finalmente, el capítulo 3 tuvo como objetivo el análisis de la evolución de la concentración de PCDD/Fs en muestras de lixiviados de vertedero de residuos sólidos urbanos, caracterizados por una relativa alta concentración de materia orgánica y cloruros, los cuales fueron tratados mediante oxidación electroquímica y Fenton. Aunque ambos procesos mostraron altas efectividades en la degradación del contenido orgánico expresado como COD, se observó un comportamiento diferente respecto a los PCDD/Fs. Mientras que la concentración de PCDDFs en las muestras de lixiviados se redujo como consecuencia del tratamiento electroquímico, la tendencia opuesta, un aumento en la concentración de PCDD/Fs, se cuantificó en las muestras de lixiviados tratadas mediante oxidación Fenton.
IV Contents
SUMMARY/ RESUMEN I
1. INTRODUCTION
1.1. Water contamination with chlorinated organic compounds 3
1.1.1. Chlorophenolic compounds 4
1.1.2. Polychlorinated dibenzo-p-dioxins and dibenzofurnas 6 (PCDD/Fs)
1.1.3. Legal framework for chlorophenols and PCDD/Fs 10
1.1.4. Case study: the problematic of landfill leachates 13
1.2. Advanced oxidation processes for wastewater treatment 15
1.2.1. Electrochemical oxidation 18
1.2.2. Fenton oxidation 22
1.3. PCDD/Fs formation during the advanced oxidation of polluted 25 waters
1.4. Thesis scope and outline 30
1.5. References 31
2. MATERIALS AND METHODS
2.1. Chemical reagents 45
2.2. Electrochemical oxidation experiments 47
2.3. Fenton oxidation experiments 49
2.4. Analytical measurements 50
2.4.1. Analysis of chemical oxygen demand 50
2.4.2. Analysis of total organic carbon 51
2.4.3. Analysis of organic acids and inorganic ions 52 2.4.4. Analysis of ammonium nitrogen concentration 53
2.4.5. Analysis of 2-chlorophenol and related aromatic 54 compounds
2.4.6. Qualitative screening of organics in the advanced 55 oxidation of 2-chlorophenol
2.4.7. Qualitative screening of organic in leachates 56
2.5. Analysis of PCDD/Fs 58
2.6. PCDD/Fs analytical methodology setup 62
2.7. Quality control in the analysis of PCDD/Fs 64
2.8. References 65
3. RESULTS AND DISCUSSION 3.1. Electrochemical oxidation of 2-chlorophenol 69
3.1.1. Major intermediate products in the electrochemical 69 oxidation of 2-chlorophenol
3.1.2. Mass balances of total organic carbon and chlorine 78
3.1.3. Formation of PCDD/Fs during the electrochemical 79 oxidation of 2-chlorophenol
3.1.4. Proposed reactions pathway for the electrochemical 83 oxidation of 2-chlorophenol
3.2. Fenton oxidation of 2-chlorophenol 87
3.2.1. Effect of H2O2 dose on Fenton degradation of 2- 88 chlorophenol
3.2.2. Effect of temperature on Fenton degradation of 2- 92 chlorophenol
3.2.3. Effect of choride ions on Fenton degradation of 2- 96 chlorophenol
3.2.4. Mass balances of total organic carbon and chlorine 99
3.2.5. Formation of PCCD/Fs during the Fenton degradation of 2- 102 chlorophenol 3.2.6. Proposed reactions pathway for the Fenton oxidation of 2- 105 chlorophenol
3.3 Electrochemical and Fenton oxidation of landfill leachates 108
3.3.1. Characterization of landfill leachate samples 108
3.3.2. Determination of PCDD/Fs in landfill leachates 116
3.3.3. Electrochemical oxidation of landfill leachates 131
3.3.4.Fenton oxidation of landfill leachates 135
3.4. References 145
4. CONCLUSIONS 161
ANNEXES 163
ANNEX I. PCDD/Fs NOMENCLATURE 163 ANNEX II.GENERAL NOMENCLATURE 165 ANNEX III. RESULTS FROM PCDD/Fs ANALYSES DISCUSSED IN 167 CHAPTER 3
Introduction
1.1. Water contamination with chlorinated organic compounds
With agriculture, livestock and energy consuming more than 80% of all water for human use, demand for fresh water is expected to increase with population growth and higher living standards (Shannon et al. 2008). According to the latest estimates of the WHO/UNICEF Joint Monitoring Programme for Water Supply and Sanitation (JMP), released in early 2013 (collected in 2011), 36% of the world’s population–2.5 billion people–lack improved sanitation facilities, and 768 million people still use unsafe drinking water sources (WHO/UNICEF, 2014).
Moreover, with growing population, water pollution has become increasingly serious because domestic and industrial activities generate high amounts of residual wastewater, whose direct disposal to natural channels causes a negative impact in the environment. Amongst possible compounds found in wastewaters, chlorinated organic compounds include some of the most toxic and largest groups of hazardous chemicals. Due to their resistance to chemical and biological degradation these compounds have been found to persist in a range of environments like lakes, rivers, groundwater systems, sediments and soils (Stringer and Johnston, 2001).
According to The European Pollutant Release and Transfer Register (E- PRTR), 3,706.7 tons of chlorinated organic substances were released to water in Europe in 2011. The main focus of water pollution with these compounds has its origin in the paper and wood production processing sector (54.5%), where chlorine is used for the bleaching of cellulose fibers (Figure 1.1). Waste and wastewater management activities, especially urban wastewater treatment plants, constitute additional emission sources of chlorinated organic compounds (37%) (E-PRTR, 2014). Besides, chlorine, which is widely used as disinfectant in treatment plants, may react with natural organic matter to produce many disinfection byproducts with harmful long term effects (Sim et al. 2009).
3
Chapter 1
3.0% 0.5% 5.0% Paper and wood production processing Waste and wastewater management 37.0% 54.5% Chemical industry Production and processing of metals Energy sector
Figure 1.1. Chlorinated organic substances released to water in 2011 per activity (E-PRTR, 2014).
Among chlorinated organic pollutants, chlorophenols (CPs) constitute a particular group of priority toxic pollutants that, as a result of their wide use, have been broadly identified in industrial wastewater, soils and surface waters (Detomaso et al. 2003).
1.1.1 Chlorophenolic compounds
CPs comprise a group of nineteen organic compounds that differ in degree and position of chlorination, ranging from monochlorophenols to the highest chloro-substituted, pentachlorophenol (PCP). They are characterized by acute toxicity, emission of strong odor, bioaccumulative behaviour and resistance to biodegradation; they are also suspected mutagens and carcinogens (Methatham et al. 2011). Consequently, some of them have been listed as priority pollutants by the U.S. EPA in the Clean Water Act (CWA) (EPA, 1980) and by the European Decision 2455/2001/CE. Toxicity of CPs, which depends on the number and location of chlorine atoms, generally increments with the increase in the number of chloro-substituents (Ozkaya 2005). At the same time, toxicity can be lowered by the presence of chlorine in the ortho position of the aromatic ring (Cañizares et al. 2003). The higher toxicity of the more chlorinated CPs may be ascribed to an increase in lypophility, which leads to a greater potential for their uptake into the organism (Pera-Titus et al. 2004).
As a result of their antimicrobial activity, CPs serve as intermediates in the synthesis of insecticides, herbicides and fungicides. In addition CPs have been applied in the manufacture of pharmaceuticals and dyes. Besides, they have been used as wood preservatives and disinfectants (Olaniran & Igbinosa 2011).
4 Introduction
On the other hand, CPs may be formed during the incineration of waste, the bleaching of pulp and paper, and the disinfection of water (Pera-Titus et al. 2004). Furthermore, large scale coal gasification and carbonization plants also generate wastewaters containing large quantities of highly toxic phenolic compounds (Gangula 2010). As consequence of their extensively use over years, and the past practice of disposal chemical waste in ordinary landfills, causing their leachate from waste dumps, CPs have been identified in industrial wastewater, surface and groundwater, soils and even in finished drinking waters (Detomaso et al. 2003; Muna et al. 2004; Oturan et al. 2009). The principal point sources of water pollution by CPs are considered to be industrial waste discharges and leaching of CPs from landfills (Olaniran & Igbinosa 2011). On the other hand, the primary nonpoint source pollution of CPs comes from the application of chlorophenol made pesticides (Czaplicka 2004).
Once released into the environment, CPs can constitute an environmental and health hazard because of their persistence and toxicity to numerous organisms. Besides, CPs can undergo reactions leading to the formation of even more dangerous products. On this way, CPs formulations are known to contain many chlorinated impurities such as polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) (Koistinen et al. 2007). Moreover, CPs and chlorobenzenes (CBzs) are thought to be the most important and direct precursors of the formation of PCDD/Fs (Weber 2007; Altarawneh et al. 2009). Therefore, it is very important to reduce the input of these compounds into the environment, as well as to develop effective methods for their removal from the receiving medium. In fact, nowadays several limitations are imposed on their use and production in many countries. In this work, we have focused on the study of 2-chlorophenol (2-CP) as model chlorinated phenol since it is one of the most applied CPs, mainly as intermediate in the manufacture of fungicides, herbicides, disinfectants and wood preservatives. In addition, the high water solubility of 2-CP contributes to make worse the problem of its potential water pollution.
5
Chapter 1
1.1.2 Polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs)
Persistent organic pollutants (POPs) represent a group of carbon-based chemicals of particular concern because they are resistant to degradation in the environment (persistent), accumulate in human and animal tissue, biomagnify along food chains and have an inherent toxicity (Fiedler 2003). PCDD/Fs are a family of unintentionally POPs that comprise two groups of planar tricyclic ethers, which can have up to eight chlorine atoms attached at carbon atoms 1 to 4 and 6 to 9, Figure 1.2 (Rappe 1994). This substitution pattern results in different homologue groups (Table I.2, Annex I) containing those congeners with the same level of chlorination. Among each group of homologues, chlorine atoms can be bound at different carbon atoms giving rise to a total of 210 possible PCDD/Fs congeners, 75 PCDDs and 135 PCDFs.
Figure 1.2. Molecular structure of PCDD/Fs.
The main physical and chemical properties of PCDD/Fs encompass high thermal stability, low vapor pressure, low solubility in water, high solubility in organic/fatty matrices and preference to bind to organic matter in soil and sediments (Fiedler 2003). As persistent compounds they are resistant to biological and chemical degradation, remaining in the environment for long times and being transported through long distances away from their emission sources (Altarawneh et al. 2009). The lypophilic nature of PCDD/Fs cause their biomagnification through the food chain, and therefore high tissue concentrations can often be found in top predator species (Van den Berg et al. 2006). These properties along with their inherent toxicity represent a serious concern with respect to the health of the environment, wildlife and humans.
6 Introduction
However, not all PCDD/Fs are toxic at extremely low concentrations; among the 210 possible congeners of PCDD/Fs, those having the positions 2,3,7,8 chlorinated, accounting for a total of 17 congeners (whose abbreviated names are listed in Table I.1, Annex I), have received the greatest public and scientific attention due to their toxicity and adverse effects (van Bavel & Abad 2008). It is worth to mention the congener 2,3,7,8-TCDD, which has one of the lowest known LD50 (lethal dose for 50% of the population) values and is frequently highlighted as “the most toxic man made chemical” (Hites 2011). Short term exposure to high levels of PCDD/Fs causes liver damage and chloracne (Marinković et al. 2010). Long term exposure is related with endocrinological, immunological, neurological, developmental, and reproductive effects (Srogi 2007; Marinković et al. 2010). The U.S. EPA has regulated 2,3,7,8-TCDD as a carcinogen based on the positive animal data and the compatible epidemiological findings of occupational exposure (Watanabe et al. 1999).
Normally, PCDD/Fs are present in samples as a mixture of isomers including toxic, 2,3,7,8-PCDD/Fs congeners, and non-2,3,7,8 chlorosubstituted compounds, which do not show toxic activity to the same extent. In order to estimate the toxicity of complex mixtures containing PCDD/Fs, the toxicity of the 2,3,7,8 congeners is related to that of the most toxic isomer, 2,3,7,8-TCDD, through the use of toxicity equivalency factors (TEFs). There are two different TEFs scales; the international toxicity equivalency factors (I-TEFs) that were established by a working group on dioxins and related compounds of the NATO Committee on the Challenge of Modern Society (CCMS), and a most recent scale, the World Health Organization toxicity equivalency factors (WHO-TEFs), established by a WHO working group after the re-evaluation of the I-TEFs (NATO 1988; Van den Berg et al. 2006). The I-TEFs are frequently used in regulatory documents such as toxic release inventories, whereas WHO-TEFs are preferred in toxicology/risk assessment studies. TEFs values for PCDD/Fs are shown in the Table 1.1.
The levels of the individual 2,3,7,8-PCDD/Fs are converted into one value of toxic equivalents (TEQ) of the isomer 2,3,7,8-TCDD. The TEQ value gives an idea of the potential toxicity of the sample and is calculated as the sum of the concentration of each congener multiplied by its TEF, TEQ = Σ [PCDD/F]i x TEFi.
7
Chapter 1
Table 1.1. Toxicity Equivalency Factors (TEFs) for the 17 PCDD/Fs congeners.
WHO-TEF WHO-TEF Compound I-TEF 1998 2005 PCDDs 2,3,7,8-TCDD 1 1 1 1,2,3,7,8-PeCDD 1 1 0.5 1,2,3,4,7,8-HxCDD 0.1 0.1 0.1 1,2,3,6,7,8-HxCDD 0.1 0.1 0.1 1,2,3,7,8,9-HxCDD 0.1 0.1 0.1 1,2,3,4,6,7,8-HpCDD 0.01 0.01 0.01 OCDD 0.0001 0.0003 0.001 PCDFs 2,3,7,8-TCDF 0.1 0.1 0.1 1,2,3,7,8-PeCDF 0.05 0.03 0.05 2,3,4,7,8-PeCDF 0.5 0.3 0.5 1,2,3,4,7,8-HxCDF 0.1 0.1 0.1 1,2,3,6,7,8-HxCDF 0.1 0.1 0.1 1,2,3,7,8,9-HxCDF 0.1 0.1 0.1 2,3,4,6,7,8-HxCDF 0.1 0.1 0.1 1,2,3,4,6,7,8-HpCDF 0.01 0.01 0.01 1,2,3,4,7,8,9-HpCDF 0.01 0.01 0.01 OCDF 0.0001 0.0003 0.001
PCDD/Fs have never been produced intentionally for any industrial use, but they occur as unwanted byproducts in many industrial and thermal processes (Fiedler 1996). According to the literature, three main categories of PCDD/Fs sources can be distinguished, namely chemical-industrial sources, thermal or combustion sources and reservoirs (Fiedler 2003). Chemical and industrial processes require the presence of carbon, hydrogen, oxygen and chlorine to generate PCDD/Fs. In addition, their formation is favored at high temperatures (> 150ºC), alkaline conditions, presence of metal catalysts and UV irradiation or other radical producing substances (Ruiz 2007). Chemical and industrial processes have resulted in the contamination of pesticides and technical products including CPs, chlorophenoxy herbicides and polychlorinated biphenyls (PCBs). The production and use of most of these chemicals are nowadays banned or strictly regulated in most countries, but during the 1960s and 1970s they were widely used, and hence became a major source of PCDD/Fs contamination in the environment (Rappe 1994). Other processes generating PCDD/Fs comprise the production of metals, the bleaching of wood pulp using chlorine and the production of chlorine gas using graphite electrodes.
8 Introduction
Thermal processes are a predictable source of PCDD/Fs when they involve chlorinated organic and inorganic compounds within a temperature range of 200-650ºC. In addition, their formation is favored in presence of carbonaceous matter and metal catalysts (Rappe 1994). Among thermal-combustion sources, of special importance is the incineration of various types of waste (municipal, hospital, hazardous), sewage sludge, sintering plants as well as diffuse sources such as domestic heating systems, automobile exhaust and combustions in landfills (Fiedler, 2003).
Since PCDD/Fs are persistent compounds and once emitted to the receiving environment, atmospheric transport moves them away from their emission sources, they are able to remain in the environment for long times (Altarawneh et al. 2009; Lee et al. 2004). As a result, they can be transferred to additional matrices called reservoirs, which are considered as secondary source of PCDD/Fs and have the potential to redistribute and recirculate them in the environment (Kulkarni et al. 2008). Reservoirs in the environment include landfills and waste dumps, contaminated soils and contaminated sediments (Fiedler 2003; Weber et al. 2008; Weber et al. 2011). Furthermore, PCDD/Fs can be produced during the treatment of sewage sludge (Öberg et al. 1993) and as a result of the biochemical and photolytic treatment of precursors such as CPs (Öberg & Rappe 1992; Vollmuth et al. 1994).
PCDD/Fs can be formed by two fundamental pathways; on the one hand, the formation via the degradation of carbon species in the presence of a chlorine source at temperatures within 200-600ºC, the so called “de novo synthesis” (Addink & Olie 1995). On the other hand, they can be formed via precursor compounds, which include compounds such as CPs, CBzs, chlorinated diphenylethers (CDEs) and PCBs (Weber 2007). Among them, it is thought that the most important and direct precursors are monocyclic aromatic compounds such as CPs and CBzs (Altarawneh et al. 2009). Two potent PCDD/Fs precursors are PCP, used as fungicidal and bactericidal agent in wood and lather treatment, and polychlorinated phenoxy phenols (PCPPs), such as triclosan, which can be also found as impurities in PCP (Holt et al. 2008). The formation of PCDD/Fs from their precursors is generally favored under alkaline conditions, heat (150- 600ºC), with even ambient temperatures in the presence of catalysts, e.g. copper, iron, aluminum salts, or radicals, as well as during irradiation with UV light (Holt et al. 2012).
9
Chapter 1
PCDD/Fs are ubiquitous contaminants in the environment and have been found throughout the world in all primary medium comprising air, soil, water, sediments, and secondary medium such as food including fish and shellfish, meat, dairy products, and consumer goods (Mukerjee 1998). In spite of their low water solubility, PCDD/Fs have been found in the aquatic environment, where they enter from atmospheric deposition, the use of agricultural chemicals and as direct discharges from industrial sources and sewage treatment plants, among others (Kim et al. 2002). Besides, PCDD/Fs leaching into the aquatic environment may be enhanced by co-disposed organics and colloidal transport (Ham et al. 2008). Since PCDD/Fs are persistent in the environment and their toxic effects can linger for long periods of time after contamination (Sim et al. 2009), the effects of low and continuous exposure to these compounds are extremely important for humans. Over time, continuous low-level exposures accumulate in the human body until subtle, adverse effects begin to occur (Mitrou et al. 2001).
1.1.3 Legal framework for chlorophenols and PCDD/Fs
The Water Framework Directive (WFD) 2000/60/EC sets out a legal framework with the aim of improving water quality across Europe. The main objective of the WFD is to maintain water of lakes, rivers, streams and groundwater aquifers in a healthy state by 2015. The WFD in its article 16, establishes criteria for the selection of priority substances which present a significant risk to or via the aquatic environment, as well as their environmental quality standards (EQS). The list of priority substances and its related EQS was established by Decision 2455/2001/EC and listed in Annex X of the WFD. The original list of 33 priority substances was expanded in Directive 2008/105/EC to include eight additional substances. Recently, the European Council passed the Directive 2013/39/EU which modifies the Directives 2000/60/EC and 2008/105/EC with regard to contaminants, defining new priority substances and determining and/or revising their corresponding EQS. Among the 19 possible CPs congeners, PCP is included in the WFD as a priority substance with an EQS value, expressed as the maximum allowable concentration in inland surface waters and other surface waters, of 1 µg L-1.
Regarding the U.S. regulation, the CWA of the EPA, enacted in 1948 and reorganized and expanded in 1972, establishes the basic structure for regulating discharges of pollutants into the waters of the U.S. and quality standards for
10 Introduction
surface waters. The list of priority pollutants in the CWA includes next CPs: 2- CP, 2,4-dichlorophenol (2,4-DCP), 2,4,6-trichlorophenol (2,4,6-TCP) and PCP. On the other hand, the Safe Drinking Water Act (SDWA) of the U.S.EPA is the main federal law that ensures the quality of drinking water. SWDA requires the U.S. EPA to establish primary drinking-water regulations for contaminants in public water systems that may have adverse effects on people’s health. Such regulations typically include a media quality standard that defines legally allowable concentrations of toxic chemicals, called maximum contaminant level (MCL). MCLs are established in order to be as close to a level that is without known or anticipated adverse health effects as it is technically or economically feasible. Among CPs, a MCL of 1 µg L-1 has been established for PCP.
With respect to PCDD/Fs, their environmental management and control is addressed at a global level through the Stockholm Convention on POPs, which was adopted in Stockholm, Sweden, on 2001 and entered into force on 17 May 2004, when it was ratified by Spain. The objective of the Stockholm Convention is to protect human health and the environment from POPs by reducing or eliminating releases to the environment (Stockholm Convention 2001). For intentionally produced POPs, this will be achieved by stopping their production and use, whereas for unintentionally POPs, such as PCDD/Fs, measures have to be taken to reduce the total releases derived from anthropogenic sources, considering if possible their total elimination (Stockholm Convention, 2001). The Stockholm Convention started with 12 initial POPs, the so called “dirty dozen”. At the fourth conference of the parties (COP-4, 2009) 9 new POPs were added, while at the fifth meeting (COP-5, 2011) one additional compound was added to the list. The total 22 POPs included in Annexes A, B or C of the Convention are shown in Table 1.2. (Fiedler et al. 2013).
11
Chapter 1
Table 1.2. POPs listed in the Stockholm Convention.
Initial 12 POPs Aldrin Chlordane Dichlorodiphenyltrichloroethane (DDT) Dieldrin Endrin Hexachlorobenzene (HCB) Heptachlor Mirex PCBs PCDDs PCDFs Toxaphene New listed POPs (COP-4, 2009) Chlordecone α-hexachlorocyclohexane (α-HCH) Hexachlorocyclohexane (β-HCH) Lindane (γ-HCH) Hexabromobiphenyl (HBB) Pentachlorobenzene (PeCBz) Tetra- and pentabromodiphenyl ether Hexa- and heptabromodiphenyl ether (c-penta PBDE) (c-octa PBDE) Perfluorooctane sulfonic acid (PFOS) New listed POPs (COP-5, 2011) Endosulfan
The European Community ratified the Stockholm Convention and approved the Regulation 850/2004/EC of 29 April 2004 on POPs. The Regulation implements the provisions of the Stockholm Convention, emphasizing the aim to eliminate the production and use of the internationally recognized POPs. Furthermore, Member States must set up emission inventories for unintentionally produced POPs, national implementation plans (NIPs) and monitoring and information exchange mechanisms. In Spain, the NIP was approved by the Spanish Cabinet with the agreement of 2 February 2007. Under the direction of the National Coordination Group (GNC), the National Reference Centre for Persistent Organic Pollutants (CNRCOP) has the task of coordinating the NIP implementation.
On the other hand, dioxins and related compounds are included as new priority substances in the new Directive 2013/39/EU that modifies the Directives 2000/60/EC and 2008/105/EC. Nevertheless, just one EQS value of 6.5 ng Kg-1 related to biota (fish) has been reported. Such EQS should not be exceeded in order to protect human health and the environment.
12 Introduction
Regarding the U.S. EPA regulation, the SDWA has set a maximum contaminant level goal (MCLG) for 2,3,7,8-TCDD in drinking water of zero, based on the best available science to prevent potential health problems. Besides, a MCL for 2,3,7,8-TCDD in drinking water of 30 pg L-1 has been set as close to the health goals as possible, considering cost, benefits and the ability of public water systems to detect and remove contaminants using suitable treatment technologies.
1.1.4 Case of study: the problematic of landfill leachates
Increasingly affluent lifestyles and industrial growth in many countries around the world has gone together with rapid increases in both municipal and industrial waste production. As consequence, municipal solid waste (MSW) management constitutes today a major environmental, economical and social problem worldwide (Renou et al. 2008). Deposit in landfills is one of the most commonly used methods for the management of MSW, and the cheapest option in terms of exploitation and capital costs (Hermosilla et al. 2009). Up to 34% of total MSW treated in Europe (EU-28 member states) in 2012 was disposed using the landfilling method, whereas this percentage was higher in the case of Spain, where 63% of the total treated MSW was landfilled, Figure 1.3 (Eurostat, 2014).
Deposit onto or into landfill Total incineration Material recycling Composting and digestion
15% 10% 10% 34% 24% 17% 63% 27%
Europe (EU-28) Spain Figure 1.3. Management alternatives of treated MSW in Europe and Spain in 2012 (Eurostat, 2014).
The Directive 99/31/EC, which intended to prevent or reduce the adverse effects of waste landfilling on the environment, in particular on surface water, groundwater, soil, air and human health, defines leachate as any liquid
13
Chapter 1
percolating through the deposited waste and emitted from or contained within a landfill. Basically, leachates are generated as a result of the physicochemical and biological degradation of waste along with the percolation of rainwater through waste layers (Hermosilla et al. 2009). Landfill leachate is a high-strength wastewater containing a large number of compounds, some of which can be expected to create a threat to health and nature if released into the natural environment. Leachates may contain great amounts of organic matter (biodegradable and recalcitrant), where humic-type substances constitutes an important group, as well as ammonia-nitrogen, heavy metals, chlorinated organic and inorganic salts (Oman & Junestedt 2008).
There are many factors affecting the composition of leachates such as landfill age, precipitation, seasonal weather variation, waste type and composition of wastes. In particular, the composition of landfill leachates varies greatly depending on the age of the landfill (Table 1.3) (Ahmed & Lan 2012). In young landfills, the characteristic large amount of biodegradable organic matter is transformed into volatile fatty acids. This early phase is called the acidogenic phase and leads to the release of biodegradable organic acids as the major constituents of the organic content (Renou et al. 2008). Generally, young leachates are characterized by high COD (> 10,000 mg L-1), high biodegradability -1 (BOD5/COD > 0.5-1), moderately content of ammonium (< 400 mg L ) and low pH values (as low as 4) (Ahmed & Lan 2012). In a later stage, in the methanogenic phase, landfill matures and the organic acids are converted to biogas. In this phase, refractory substances such as humic acids dominate the organic content (Renou et al. 2008). Mature or stabilized leachates are characterized by a relatively low COD (< 4000 mg L-1), slightly basic pH (7.5-8.5), low biodegradability (BOD5/COD < 0.1), and high molecular weight compounds such as humic substances (Ahmed & Lan 2012).
Table 1.3. Classification of landfill leachates as a function of the landfill age.
Young Intermediate Mature Age (years) < 5 5-10 > 10 pH < 6.5 6.5-7.5 > 7.5 COD (mg L-1) > 10,000 4,000-10,000 < 4,000
BOD5/COD 0.5-1 0.1-0.5 < 0.1 Ammonium (mg L-1) < 400 - > 400
14 Introduction
Although the discharge of solid waste from chemical industries into landfills was a common practice in industrial countries until the 1970-1980s, when hazardous waste incinerators increased their use for waste management, nowadays, hazardous chemical waste, especially from production of organochlorinated chemicals, is still landfilled in many countries (Weber et al. 2008). In addition, a considerable quantity of chlorinated and persistent toxic substances contained within consumer goods finishes in household waste and therefore is still widely deposited in landfills (Weber et al. 2011). Contaminants from landfills are released into leachates becoming into an important source of ground and surface water contamination.
As consequence of the extended use of CPs over the years and the past practice of their dumping in ordinary landfills, their leaching from landfills is considered as an important source of water pollution (Olaniran & Igbinosa 2011). In the case of PCDD/Fs, the defined low POPs level (15 µg TEQ kg-1) is high compared with the PCDD/Fs level found in most contaminated waste, and therefore, a large part of waste containing PCDD/Fs is excluded from the obligations and continue being disposed in landfills (Weber et al. 2011).
1.2 Advanced oxidation processes for wastewater treatment
The need to restore contaminated sites in order to avoid further risks to the environment and threats to human health, along with stricter water quality control and regulations against hazardous pollutants, has aroused in the last few years the development of effective methods for the removal of pollutants from wastewater. Conventional biological processes do not always provide adequate results, especially for industrial wastewaters, since many of them contain toxic or resistant to biological treatment substances such as CPs (Oller et al. 2011). The same drawback must to be considered regarding the handling of mature leachates since their low biodegradability and the presence of high quantities of refractory material make ineffective their biological treatment. On the other hand, physical/chemical treatments such as flocculation, precipitation, adsorption or reverse osmosis require additional post-treatment processes to remove the pollutants from the newly contaminated stream (Oturan et al. 2009).
15
Chapter 1
As an alternative to conventional treatments, advanced oxidation processes (AOPs) have already been used for the treatment of wastewaters containing recalcitrant organic compounds such as pesticides, pharmaceuticals and endocrine disrupting chemicals, dyes and phenolic compounds (Brillas et al. 2004; Pera-Titus et al. 2004; Cañizares et al. 2007; Dalrymple et al. 2007; Bautista et al. 2008; Martínez-Huitle & Brillas 2009). Furthermore, they have been applied as pretreatment methods to enhance the biodegradability of several wastewaters (Wiszniowski et al. 2006) and for the mineralization of recalcitrant organics in landfill leachates (Primo et al. 2008; Anglada et al. 2009b). The main advantages of AOPs include fast reaction rates and non- selective oxidation, allowing the treatment of many pollutants at the same time, and potential to reduce the toxicity of contaminants (Vilhunen & Sillanpää 2010).
AOPs were defined by Glaze et al. (1987) as near ambient temperature and pressure water treatment processes that involve the generation of highly reactive radicals, specially hydroxyl radicals (OH), in sufficient quantity to effect water purification. OH are powerful oxidants with an oxidation potential of 2.8 V vs NHE (normal standard hydrogen electrode), which react non-selectively and rapidly with most organic compounds (Gogate & Pandit 2004; Pera-Titus et al. 2004). Once generated, OH can react with organic compounds by radical addition, hydrogen abstraction or electron transfer, depending on the structure and ionization potential of the organic pollutants (Bautista et al. 2008). One possible classification of AOPs, which is based on the way to generate the OH, is shown in Figure 1.4.
16 Introduction
Photolysis, Ultraviolet (UV) Photocatalysis Vacuum Ultraviolet (VUV)
O3 based O3/UV /H2O2 processes
H2O2 based UV/H2O2; Fenton; Fenton-like; processes Photo-Fenton
Oxidation Supercritical oxidation Subcritical oxidation under hard Wet oxidation (WO) conditions Wetcatalytical oxidation (WCO)
Electrochemical oxidation
Sonolysis
Electromagnetic radiations
Figure 1.4. Classification of advanced oxidation processes.
Although the total mineralization of organic pollutants by AOPs can be achieved, it is generally expensive because the produced intermediates tend to be more resistant to chemical degradation, and therefore, energy and chemical reagents consumption increases with time (Oller et al. 2011). As alternative to cut down treatment costs, AOPs can be applied as a pre-treatment stage to improve the biodegradability of initially persistent compounds, which then can be treated with a biological process from a more economical point of view (Pera-Titus et al. 2004; Vilhunen & Sillanpää 2010; Oller et al. 2011). Another step on this way is the use of renewable energy sources to power AOPs as in the case of solar photocatalysis (Comninellis et al. 2008).
Among different AOPs, Fenton and electrochemical oxidation have been applied in this thesis. Fenton oxidation is considered one of the most popular technologies for wastewater treatment as shown by the large amount of data available in the literature (Bautista et al. 2008; Babuponnusami & Muthukumar 2014; Umar et al. 2010), whereas electrochemical oxidation has received special
17
Chapter 1
attention in recent years because it exhibits high treatment efficiencies (Anglada et al. 2009a; Martínez-Huitle & Brillas 2009; Sirés et al. 2014).
1.2.1 Electrochemical Oxidation
Electrochemical oxidation has been proved to be an environmentally benign technology able to mineralize completely non-biodegradable organic matter and to eliminate nitrogen species (Anglada et al. 2009a). Electro- oxidation is a process in which a chemical reaction is forced to occur at an electrode by an imposed voltage. The reaction takes place in an electrolytic reactor where the solution to be treated is in contact with the cell electrodes, where the oxidation (anode) and reduction (cathode) occur.
Electrochemical oxidation of pollutants can take place through two different mechanisms: (i) direct anodic oxidation, where the pollutants are destroyed at the anode surface (Figure 1.5); (ii) indirect oxidation, where a mediator is electrochemically generated to carry out the oxidation (Figure 1.6). It must to be highlighted that during the electro-oxidation of aqueous effluents, both oxidation mechanisms may coexist (Anglada et al. 2009a).
Mass transport
RH RH e-
Anode RH-O RH-O
Electrochemical reaction
Figure 1.5. Mechanism scheme of anodic oxidation of organic compounds on the anode surface.
Direct electrochemical oxidation (Figure 1.5) of contaminants is considered to take place mainly in two steps: (i) diffusion of pollutants (RH) from the bulk solution to the anode surface and (ii) oxidation of pollutants at anode surface. The oxidation of pollutants occurs once they have been adsorbed on the anode
18 Introduction
surface. The oxidized compound (RH-O) either can continue being adsorbed and oxidized, or can be desorbed and pass to the bulk solution.
Mass transport
H2O e- RH RH
Anode OH· RH-O RH-O
Electrochemical reaction
Figure 1.6. Mechanism scheme of the oxidation of organic compounds by means of OH· generated on the anode surface.
In the indirect electrochemical oxidation, electro-active species generated at the anode surface act as intermediary for shuttling electrons between the electrode and the pollutants (Panizza 2010). A great deal of electrochemical reactions involved in the treatment of waters take place by means of OH· produced as a result of the water discharge on the electrode surface (Figure 1.6.). OH are characterized by a high oxidation potential and non-selective character. Once generated, they have a short half-life and therefore react quickly with the organic matter, as well as with other species present in the solution to give rise to additional oxidants. Besides, OH· can combine between themselves leading to the formation of oxygen.
Electrochemical oxidation is generally characterized by simple equipment, easy operation and brief retention time. Other advantages include robustness, versatility and amenability to automation (Anglada 2010). On the other hand, the main drawbacks comprise high operating costs due to the high energy consumption, the external addition of electrolytes when the solution does not have sufficient conductance and electrode fouling.
The nature of the electrode material has strong influence in the selectivity and efficiency of the electrochemical process. Whereas some anodes encourage the partial and selective oxidation of the pollutants, any others lead to the
19
Chapter 1
complete mineralization of the pollutants (Chen 2004). The competition between the oxidation of the pollutants on the anode surface and the inevitable side reaction of water oxidation, which is also called oxygen evolution reaction
(r1), must be considered in the selection of the proper anodic material (Anglada et al. 2009a). The current efficiency depends on the extent in which organic oxidation takes place with respect to the oxygen evolution reaction.
+ − 2H2O → O2 + 4H + 4e (r1)
Electrodes can be classified according to their interaction with OH as non- active electrodes (weak interaction) or active electrodes (strong interaction). As a general rule, the weaker is the interaction between the electrode and OH, the smaller is the electrochemical activity towards the oxygen evolution and the higher is the activity towards the oxidation of the organic matter (Kapałka et al. 2007). On the other hand, electrodes can be classified depending on the oxygen evolution overpotential. This is defined as the difference between the real and the thermodynamic value (1.2 V vs NHE) at which the oxidation of water takes place. In Table 1.4, the oxygen evolution overpotential for different electrodes is shown.
Table 1.4. Oxygen evolution overpotential for different electrode materials.
Material Potential (V) Conditions
Pt 1.3 0.5 M H2SO4
Pt 1.6 0.5 M H2SO4
IrO2 1.6 0.5 M H2SO4
Graphite 1.7 0.5 M H2SO4
PbO2 1.9 1 M H2SO4
SnO2 1.9 0.5 M H2SO4
TiO2 2.2 1 M H2SO4
Si/BDD 2.3 0.5 M H2SO4
Ti/BDD 2.7 0.5 M H2SO4
By this way, it can be distinguished between high and low overpotential electrodes. Electrodes with high oxygen evolution overpotential are high oxidizing power anodes, characterized by non-active behavior and high current efficiency for the oxidation of the organic matter. On the contrary, low overpotential anodes are low oxidizing power anodes, characterized by active
20 Introduction
behavior and low current efficiency for the oxidation of organic compounds (Kapałka et al. 2007; Kapałka et al. 2010).
Boron-doped diamond (BDD) has reported to yield higher oxidation rates and greater current efficiencies than additional materials such as PbO2 and
Ti/SnO2-Sb2O5 (Anglada et al. 2009a). This fact can be explained due to its properties such as high thermal conductivity, electrochemical stability, inert surface and a high oxygen evolution overpotential, which minimizes the not desired oxygen evolution side reaction (Chen 2004).
Depending on the applied potential, the oxidation of the organic matter at BDD anodes can follow two: (i) direct electron transfer in the potential region prior to the oxygen evolution (water stability); (ii) oxidation by electrogenerated OH· in the region of water evolution (water decomposition) (Panizza & Cerisola 2005). In addition, it must be taken into account that depending on the inorganic salts present in the reaction medium, additional oxidizing species such as free chlorine or peroxodisulphate ions can be electrochemically generated on the anode. These oxidizing species diffuse into the bulk solution contributing to the overall oxidation by indirect oxidation (Palmas et al. 2007).
Regarding the two main oxidation mechanisms, it was found that working in the water stability region, BDD has not electrocatalytic activity for the direct oxidation of aliphatic alcohols. However, this behavior is quite different during the oxidation of aromatic and heterocyclic compounds, since these compounds are susceptible to be oxidized by single electron transfer in the region of water stability. Nevertheless, such oxidation results in the deactivation of the electrode as a result of the deposition of a film on the anode surface (Panizza & Cerisola 2005).
On the other hand, it has been established that if the oxidation of the organic compounds with BDD takes place in the oxygen evolution region, OH are involved in the reaction (Polcaro et al. 2004). Comninellis and coworkers proposed a mechanism for the oxidation of organic compounds on BDD with the oxygen evolution reaction simultaneously. The model assumes that both reactions, the oxidation of the organic matter (r3) and the oxygen evolution (r4), take place with OH as intermediates, which are generated from the decomposition of water (r2) (Panizza & Cerisola 2005):
21
Chapter 1
. + − BDD + H2O → BDD(OH ) + H + e (r2)
. BDD(OH ) + R → BDD+ mCO2 + nH2O (r3)
. 1 + − BDD(OH ) → BDD+ O + H + e (r4) 2 2 The reaction between the organic matter and OH competes with the anode discharge of these radicals to lead oxygen. The BDD is a high oxidizing power anode, with a high potential for the oxygen evolution reaction (2.3 to 2.7 V vs. NHE). Working at this potential, it is thermodynamically feasible to form OH (2.38 V vs. NHE) as shown by Marselli et al. (2003) who detected OH· on BDD electrodes by electron spin resonance. Therefore, during the electrolysis using BDD in the potential region of oxygen evolution a large amount of OH are produced. OH are weakly adsorbed on the anode surface and consequently have a high reactivity for the oxidation of the organic compounds, providing the possibility of an efficient wastewater treatment (Panizza and Cerisola, 2005). Finally, it is noteworthy that working at high potentials, in the region of water decomposition, no evidence of electrode deactivation has been found, and even the formed polymeric film was destroyed restoring the activity of the electrode (Panizza & Cerisola 2005).
1.2.2 Fenton oxidation
· The Fenton oxidation process is based on the generation of OH from H2O2 using Fe2+ as catalyst at acidic pH. Despite having been studied in detail, there is some controversy about the exact mechanism of the Fenton process and the nature of the species generated during the process (Bossmann et al. 1998). The most commonly accepted reaction scheme establishes that OH are produced 2+ according to reaction r5, whereas the catalyst, Fe , is either regenerated trough 3+ reaction r6 or from the reaction between Fe and the intermediate organic radicals, R , (r7-r9) (Bautista et al. 2008; Pignatello et al. 2006). However, such reactions, which allow the regeneration of Fe2+, are several orders of magnitude 2+ slower than reaction r5, decreasing the amount of Fe available in the reaction medium, and therefore becoming the rate-limiting step. In addition, Fe3+ can form complexes with many organic and inorganic ligands and scavenge iron from the chain reactions. It is noteworthy to mention the formation of complexes between Fe3+ and carboxylic acids, since these compounds are
22 Introduction
frequently present as intermediates during the oxidation process of organic compounds.
2+ 3+ . - Fe + H2O2 → Fe + OH + OH r5
3+ 2+ . + Fe + H2O2 → Fe + HO2 + H r6
. . . RH+ OH → R + H2O r7
. - + 2+. R + OH → R + Fe r8
+ - . R + OH → R -OH r9
Besides, additional competitive reactions may take place (r10-r13):
2+ . 3+ - Fe + OH → Fe + OH r10
. . H2O2 + OH → HO2 + H2O r11
. . . HO2 + OH → O2 + H2O r12
.. . OH + OH → H2O2 r13
The OH·react with the organic matter by three different pathways depending on the nature of the organic compounds and leading to the formation of carbon-centered radicals (Pignatello et al. 2006):
i) Abstraction of H from C-H, N-H and O-H bonds (r14):
. . RH+ OH → R + H2O r14
ii) Adding to double bonds (C=C) (r15):
. . R −C = C + OH → R -C -C -OH r15
iii) Adding to aromatic rings (r16): OH
+ OH. r16
↔
23
Chapter 1
The organic intermediates formed (R) may react with OH, oxygen and additional oxidant species, with the overall process leading eventually to the mineralization to CO2 and H2O, and the formation of inorganic acids if the pollutant contains heteroatoms.
The major advantages of the Fenton process are: i) the Fenton reagent is easy to handle and environmentally benign; ii) there are no mass transfer limitations due to its homogeneous catalytic nature; and iii) the process is technologically simple and requires relatively mild operating conditions (Esplugas et al. 2002; Lopez et al. 2004; Du et al. 2007). The efficiency of the Fenton process depends on different operating conditions, such as working pH, temperature and H2O2 and catalyst concentrations. Optimum working pH is in the range 2-4 (Pera-Titus et al. 2004). At pH above 4 the degradation efficiency decreases because iron begins to precipitate as Fe(OH)3, thereby lowering its ability to catalyze H2O2 (Kavitha & Palanivelu
2003). On the other hand, at pH lower than 2, H2O2 is solvated to form oxonium + 2+ ion (H3O2 ) reducing the reactivity of H2O2 with Fe , and thereby lowering the concentration of OH (Kavitha & Palanivelu 2003). In addition, Fe2+ regeneration 3+ by the reaction of Fe with H2O2 is inhibited at acidic pH values (Bautista et al. 2008).
Regarding the reaction temperature, it should be expected that its increase enhanced the process kinetics (Bautista et al. 2008). However, some authors have claimed that the use of Fenton oxidation at high temperatures can be limited due to the thermal instability of H2O2 (Gogate & Pandit 2004). Nevertheless, Zazo and coworkers found that increasing the temperature in the · range 25 to 130ºC resulted in faster Fe-catalyzed conversions of H2O2 into OH , enhancing mineralization rates (Zazo et al. 2011).
One important factor affecting the Fenton reaction is H2O2 and iron concentration. H2O2 dose is important in order to obtain better degradation efficiency, while iron concentration is important for the reaction kinetics
(Chamarro 2001). H2O2 dose is frequently established according to the initial pollutant concentration. It is expected that as the H2O2 dose increases, more OH· are available in the medium and the degradation of contaminants increases. However, there is a maximum value above which an increment in
H2O2 quantity does not enhance pollutant degradation since H2O2 starts to react with OH acting as a free-radical scavenger, decreasing the OH concentration
24 Introduction
· and generating HO2 radicals, which much less reactive (r11 and r12) (Primo et al. 2+ 2008). On the other hand, it is desirable to have ratios Fe /H2O2 as small as possible in order to reduce the sludge production from iron complexes, and to avoid OH recombination. Besides, higher Fe2+ concentrations hinder the efficiency of the system as it competes with the target substrate for OH while 2+ at the same time terminating the reaction cycle by oxidizing Fe (r10) (Xu et al.
2003; Lopez et al. 2005; Duesterberg & Waite 2006). Different optimal H2O2 to Fe2+ mass ratios have been found in the literature such as 20 (Xu et al. 2003), 5- 25 (Pera-Titus et al. 2004) and 61-607 (Pignatello et al. 2006).
1.3 PCDD/Fs formation during the advanced oxidation of pollutes waters
In general, the basis for the destruction of organic pollutants using green chemistry processes includes: high destruction efficiency processes operated preferentially at room temperature and atmospheric pressure and starting materials and any reagents that are non-toxic and environmentally benign. Moreover, processes should release non-toxic, environmental compatible substances and contain any residual waste or keep any hazardous byproducts from escaping to the environment (Laine & Cheng 2007).
However, sometimes the partial oxidation of organic contaminants during the remediation of wastewaters may produce transformation products being more toxic than their parent compounds (Fatta-Kassinos et al. 2011). Therefore, the only evaluation of conventional physicochemical parameters in wastewater treatment such as COD and TOC, remains insufficient to assess the biocompatibility of AOP-treated effluents before they can be safely discharged into receiving waters (Fatta-Kassinos & Michael 2013; Karci 2014). Accordingly, wastes released to aquatic systems should be subjected to more stringent standards (Sedlak & Schnoor 2013) and the assessment of transformation products and toxicity must be considered in extending the understanding of the overall efficiency of AOPs for pollutants degradation (Karci 2014). In fact, the transformation products are increasingly being taken into account in studies relating to the ecotoxicological impact of AOPs (Karci, 2014).
The potential formation of POPs and other toxic byproducts, concretely, if the highly toxic PCDD/Fs are formed and under which operating conditions their formation is relevant, is consequently one important criterion to take into
25
Chapter 1
account in the assessment of treatment technologies (Weber 2007). Deep knowledge regarding the formation of PCDD/Fs during the application of AOPs for the abatement of chlorinated organic compounds considered as potential PCDD/Fs precursors is still lacking to date. Therefore, despite the high effectiveness of AOPs in the oxidation of major contaminants and conventional parameters, scarce references deal with the monitoring of PCDD/Fs in the course of the oxidation processes. By this way, Table 1.5 summarizes the works reported in the scientific literature regarding formation of PCDD/Fs within the advanced oxidation treatment of wastewaters containing mainly chlorinated phenols and related organic compounds considered as potential PCDD/Fs precursors. As it is depicted in Table 1.5, photolysis based processes are among the most studied treatments. In relation to Fenton and electrochemical oxidation, studies are more limited to date and data regarding the formation of PCDD/Fs is covered basically from a qualitative approach. Therefore, more studies are needed, including their extent to the treatment of real wastewaters, in order to take into account the effect of coexisting organic and/or inorganic species on the generation of transformation products, as well as a detailed and quantitative assessment of the influence of the operating variables.
26 Introduction
Table 1.5. Summary on the formation of PCDD/Fs during the advanced oxidation treatment of wastewaters.
Formation of PCDD/Fs in the treatment of landfill seepage waters Reference AOP Matrix Summary of main results • Formation of 1,2,3,4,6,7,8-HpCDD, 1,2,3,6,7,8-HxCDD and Vollmuth et al. Seepage water from a Photolysis: UV (254 nm) OCDF. (1994) landfill • 69.3% reduction for OCDD. • Not PCDD/Fs degradation was observed. Vollmuth and Seepage water from a UV (254 nm)/O (960 mg h-1) • A concentration increase for 2,3,7,8-TCDD, OCDD, 2,3,7,8- Niessner (1995) 3 landfill TCDF and OCDF was observed. Formation of PCDD/Fs in the treatment of waters containing dioxin precursors Reference AOP Matrix Summary of main results • Purified PCP: formation of Hepta- and Octa-CDD/Fs (1,2,3,4,6,7,8-HpCDD, 1,2,3,4,5,7,9-HpCDD, OCDD, and Photolysis: UV (254 nm) Synthetic waters 1,2,3,4,6,7,8-HpCDF) Vollmuth et al. -1 [PCP]0= 1 mg L (purified) containing purified • Technical PCP (PCDD/Fs present as impurities): formation (1994) -1 [PCP]0= 0.92 mg L (technical) and technical PCP of 1,2,3,4,7,8-HxCDD and 1,2,3,4,6,7,8-HpCDF congeners. Degradation between 80.6 and 100% for remaining congeners. Hong et al. Photolysis: UV Reagent water -1 • Formation of 1,2,3,4,6,7,8-HpCDD and OCDD. (2000) [PCP]0= 5 mg L containing PCP • Formation of OCDD: Photolysis: UV (λ > 370nm), - In the presence of Fe(III) only: OCDD was detected in Fe3+= 560 mg L-1 Aqueous solution Fukushima et al. the order of several hundred nM (pH=3: 150 nM; pH=5: -1 containing PCP, Fe(III) (2000) HA= 50 mg L 890 nM). -1 and Humic Acid (HA) [PCP]o= 2665 mg L - In the presence of Fe(III)-HA: OCDD peak decreased too low to be quantified.
27
Chapter 1
Table 1.5. Summary on the formation of PCDD/Fs during the advanced oxidation treatment of wastewaters (cont.)
Formation of PCDD/Fs in the treatment of waters containing dioxin precursors Reference AOP Matrix Summary of main results Liquid formulation: 2,4-D • 93 PCDD/Fs congeners increased their concentration by Photolysis: solar (2,4-dihclorophenoxyacetic 5600% and 3000% respectively in two chlorinated Holt et al. (2012) [PCNB]0= 30 g acid). pesticides formulations containing PCNB and 2,4-D after
[2,4-D]0= 24 g Solid formulation: PCNB their exposure to natural light during 833 h and 1.6 h, (pentachloronitrobenzene). respectively. Reagent water and • 2,7/2,8-DCDD was the major degradation product of TCS: wastewater from an urban - Reagent water (pH=7): [2,7/2,8-DCDD] ≈ 2 mg/l Mezcua et al. max Photolysis: solar wastewater treatment after 1000 min. (2004) plant spiked with triclosan - Wastewater (pH=8): [2,7/2,8-DCDD]max≈ 8.5 mg/l (TCS). after 1000 min. Photolysis: Aranami & artificial white light Pure water, fresh water and • 2,7/2,8-DCDD detected after 3 days of irradiation in fresh Readman (2007) -1 sea water spiked with TCS. water (≈ 0.025 mg/l) and sea water (0.045 mg/l). [TCS]0= 9.4 mg L Photolysis: • SPME fiber: 0.4-1.5 % of 2,7/2,8-DCDD was produced. SPME fiber coating and Lores et al. (2005) UV (254 nm) • Pure water: photoformation of 2,7/2,8-DCDD was -1 pure water spiked with TCS. [TCS]0= 5-200 µg L reported a basic media is not required at 254 nm. Photolysis: UV (254 nm) Wastewater from a sewage Sanchez-Prado et Simulated sunlight treatment plant. • The formation of 2,8-DCDD was confirmed in all al. (2006) irradiation SPME fiber coating from experiments at different working pH (from 3 to 8.5). -1 wastewater extraction. [TCS]0= ng mL level
28 Introduction
Table 1.5. Summary on the formation of PCDD/Fs during the advanced oxidation treatment of wastewaters (cont.)
Formation of PCDD/Fs in the treatment of waters containing dioxin precursors Reference AOP Matrix Summary of main results Photocatalysis: Sankoda et al. Reagent water containing • Production of 2,8-DCDD: [2,8-DCDD]max ≈ 0.25 µg/l after UV (TiO2) (2011) -1 TCS. 120 min. [TCS]0= 1 mg L Photo-Fenton: • H O /Fe(III) system: OCDD was formed at pH 5 (Not UV (λ < 370 nm) Aqueous solution of PCP: 2 2 Fukushima & 3+= observed at pH 3). Fe 5.6 mg L-1 [H O /Fe(III) system]. Tatsumi (2001) 2 2 • H O /Fe(III)/HA system: not formation of OCDD was HA= 50 mg L-1 [H O /Fe (III)/HA system]. 2 2 2 2 observed. [PCP]o= 13.3 mg L-1 Fenton-like: H2O2= 1375.9 mg L-1; • Formation of PCBs, PCDDs and dichlorodiphenyl ethers 3+ Reagent water containing Munoz et al. (2011) Fe = 10 mg L-1 working with substoichiometric H O doses and low iron monochlorophenols. 2 2 [4-CP, 3-CP, 2-CP]o= quantities. 2000 mg L-1 Electrolysis: -2 • Formation of minor compounds such as J= 1000 mA cm Synthetic solution of Hong et al. (2003) hydroxylated/chloro-diphenylethers, hydroxylated IrO2/Ti anode 4-chlorophenol. -1 dibenzodioxins/furans and hydroxylated chlorobiphenyls. [4-CP]o= 30 mg L
29
Introduction
1.4 Thesis scope and outline
This thesis focuses on the study of intermediate products, paying special attention to the potential formation of PCDD/Fs, during the electrochemical oxidation and Fenton treatment of model solutions containing 2-CP, a known precursor of the formation of PCDD/Fs, and real wastewaters such as the leachates from a MSW landfill.
Chapter 3.1 examines the electrochemical oxidation of aqueous solutions containing 2-CP on BDD anodes. The influence of the supporting electrolyte is determined using two different and commonly employed electrolytes, NaCl and
Na2SO4. The distribution of major oxidation byproducts, as well as the formation of PCDD/Fs, is quantitatively assessed in the course of the electrooxidation treatment. Finally, a scheme of the reactions pathways in the oxidation of 2-CP based on the identified species is shown.
In chapter 3.2 the Fenton oxidation of aqueous solutions of 2-CP is reported. The effect of H2O2 dose, working temperature and the presence of chloride ions in the reaction medium is evaluated. The formation of oxidation products, focusing on the possible production of PCDD/Fs, as a result of the Fenton treatment is assessed. A possible reaction pathway is then proposed.
Chapter 3.3 explores the advanced oxidation treatment of real complex wastewaters such as leachates coming from a MSW landfill. First, leachates are characterized in terms of major physicochemical parameters. In addition, qualitative analysis of organic compounds in leachates is also performed. Afterwards, the electrochemical and Fenton oxidation of leachates is evaluated through the analysis of major parameters, such as COD and TOC. Furthermore, the monitoring of PCDD/Fs during the treatment of leachates is displayed.
Finally, chapter 4 summarizes the main results obtained in this thesis and highlights the main aspects that need to be faced.
30
Introduction
1.5 References
Addink, R. & Olie, K., 1995. Mechanisms of Formation and Destruction of Polychlorinated Dibenzo-p-dioxins and Dibenzofurans in Heterogeneous Systems. Environmental science & technology, 29(6), pp.1425–35.
Ahmed, F.N. & Lan, C.Q., 2012. Treatment of landfill leachate using membrane bioreactors: A review. Desalination, 287, pp.41–54.
Altarawneh, M. et al., 2009. Mechanisms for formation, chlorination, dechlorination and destruction of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs). Progress in Energy and Combustion Science, 35(3), pp.245–274.
Anglada, Á., 2010. Electro-oxidation on Boron-doped Diamond Anodes of Ammonia and Organic Pollutants in Landfill Leachate. Universidad de Cantabria.
Anglada, A., Urtiaga, A. & Ortiz, I., 2009a. Contributions of electrochemical oxidation to waste-water treatment: fundamentals and review of applications. Journal of Chemical Technology & Biotechnology, 84(12), pp.1747–1755.
Anglada, A., Urtiaga, A. & Ortiz, I., 2009b. Pilot Scale Performance of the Electro-Oxidation of Landfill Leachate at Boron-Doped Diamond Anodes. Environmental Science & Technology, 43(6), pp.2035–2040.
Aranami, K. & Readman, J.W., 2007. Photolytic degradation of triclosan in freshwater and seawater. Chemosphere, 66(6), pp.1052–6.
Babuponnusami, A. & Muthukumar, K., 2014. A review on Fenton and improvements to the Fenton process for wastewater treatment. Journal of Environmental Chemical Engineering, 2(1), pp.557–572.
Bautista, P. et al., 2008. An overview of the application of Fenton oxidation to industrial wastewaters treatment. , 1338(April), pp.1323–1338.
Van Bavel, B. & Abad, E., 2008. Long-Term Worldwide QA/QC of Dioxins and Dioxin-like PCBs in Environmental Samples. Analytical Chemistry, 80(11), pp.3956–3964.
Van den Berg, M. et al., 2006. The 2005 World Health Organization reevaluation of human and Mammalian toxic equivalency factors for dioxins
31
Chapter 1
and dioxin-like compounds. Toxicological sciences : an official journal of the Society of Toxicology, 93(2), pp.223–41.
Bossmann, S.H. et al., 1998. New Evidence against Hydroxyl Radicals as Reactive Intermediates in the Thermal and Photochemically Enhanced Fenton Reactions. The Journal of Physical Chemistry A, 102(28), pp.5542–5550.
Brillas, E. et al., 2004. Electrochemical destruction of chlorophenoxy herbicides by anodic oxidation and electro-Fenton using a boron-doped diamond electrode. Electrochimica Acta, 49(25), pp.4487–4496.
Cañizares, P. et al., 2003. Electrochemical oxidation of several chlorophenols on diamond electrodes - Part I. Reaction mechanism. In Journal of Applied Electrochemistry. pp. 917–927.
Cañizares, P. et al., 2007. Electrochemical Oxidation of Wastewaters Polluted with Aromatics and Heterocyclic Compounds. Journal of The Electrochemical Society, 154(11), p.E165.
Chamarro, E., 2001. Use of fenton reagent to improve organic chemical biodegradability. Water Research, 35(4), pp.1047–1051.
Chen, G., 2004. Electrochemical technologies in wastewater treatment. Separation and Purification Technology, 38(1), pp.11–41.
Comninellis, C. et al., 2008. Advanced oxidation processes for water treatment: advances and trends for R&D. Journal of Chemical Technology & Biotechnology, 83(6), pp.769–776.
Czaplicka, M., 2004. Sources and transformations of chlorophenols in the natural environment. The Science of the total environment, 322(1-3), pp.21–39.
Dalrymple, O.K., Yeh, D.H. & Trotz, M.A., 2007. Removing pharmaceuticals and endocrine-disrupting compounds from wastewater by photocatalysis. Journal of Chemical Technology & Biotechnology, 82(2), pp.121–134.
Detomaso, A. et al., 2003. Practical Applications of the Fenton Reaction to the Removal of Chlorinated Aromatic Pollutants: Oxidative Degradation of 2,4- Dichlorophenol. In Environmental Science and Pollution Research. pp. 379–384.
Du, Y., Zhou, M. & Lei, L., 2007. The role of oxygen in the degradation of p- chlorophenol by Fenton system. Journal of hazardous materials, 139(1), pp.108–15.
32 Introduction
Duesterberg, C.K. & Waite, T.D., 2006. Process Optimization of Fenton Oxidation Using Kinetic Modeling. Environmental Science & Technology, 40(13), pp.4189–4195.
EC, 1999. Council Directive 1999/31/EC of 26 April 1999 on the landfill of waste. Official Journal of the European Communities 16.7.1999: L 182, 1-19.
European Commission (EC), 2000. Directive 2000/60/EC of the European Parliament and of the Council establishing a framework for the Community action in the field of water policy. Official Journal of the European Communities 22.12.2000: L 327, 1 -72.
EC, 2001. Decision No 2455/2001/EC of the European Parliament and of the Council of 20 November 2001 establishing the list of priority substances in the field of water policy and amending Directive 2000/60/EC. Official Journal of the European Communities 15.12.2001: L 331, 1-5.
EC, 2008. Directive 2008/105/EC of the European Parliament and of the Council of 16 December 2008 on environmental quality standards in the field of water policy, amending and subsequently repealing Council Directives 82/176/EEC, 83/513/EEC, 84/156/EEC, 84/491/EEC, 86/280/EEC and amending Directive 2000/60/EC of the European Parliament and of the Council. Official Journal of the European Union 24.12.2008: L 348, 84-97.
European Union, 2013. Directive 2013/39/EU of the European Parliament and of the Council of 12 August 2013 amending Directives 2000/60/EC and 2008/105/EC as regards priority substances in the field of water policy. Official Journal of the European Communities. Official Journal of the European Communities 24.8.2013: L 226, 1-17.
E-PRTR, 2014. European Pollutant Release and Transfer Register. Coppenaghen, Denmark: European Environment Agency (EEA). Available at http://prtr.ec.europa.eu/. Accessed May 2014.
Esplugas, S. et al., 2002. Comparison of different advanced oxidation processes for phenol degradation. Water Research, 36(4), pp.1034–1042.
EUROSTAT, 2014. Environmental data center on waste. Luxembourg: European Commission. Available at http://epp.eurostat.ec.europa.eu/. Accessed May 2014.
33 Chapter 1
Fatta-Kassinos, D. et al., 2011. The risks associated with wastewater reuse and xenobiotics in the agroecological environment. The Science of the total environment, 409(19), pp.3555–63.
Fatta-Kassinos, D. & Michael, C., 2013. Wastewater reuse applications and contaminants of emerging concern. Environmental science and pollution research international, 20(6), pp.3493–5.
Fiedler, H., 2003. Dioxins and Furans (PCDD/PCDF). ChemInform, 34(26), pp.11–13. Available at: http://doi.wiley.com/10.1002/chin.200326261.
Fiedler, H., 1996. Sources of PCDD/PCDF and impact on the environment. Chemosphere, 32(1), pp.55–64.
Fiedler, H. et al., 2013. The need for capacity building and first results for the Stockholm Convention Global Monitoring Plan. TrAC Trends in Analytical Chemistry, 46(i), pp.72–84.
Fukushima, M. & Tatsumi, K., 2001. Degradation pathways of pentachlorophenol by photo-Fenton systems in the presence of iron(III), humic acid, and hydrogen peroxide. Environmental science & technology, 35(9), pp.1771–8. A
Fukushima, M., Tatsumi, K. & Morimoto, K., 2000. Influence of iron(III) and humic acid on the photodegradation of pentachlorophenol. Environmental Toxicology and Chemistry, 19(7), pp.1711–1716.
Gangula, S ., 2010. Degradation of Chlorophenols in Swine Waste. Western Kentucky University.
Glaze, W.H., Kang, J.W. & Chapin, D.H., 1987. Chemistry of water treatment processes involving ozone, hydrogen peroxide and ultraviolet radiation. In Ozone: Science and Engineering. pp. 335–352.
Gogate, P.R. & Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment I: oxidation technologies at ambient conditions. Advances in Environmental Research, 8(3-4), pp.501–551.
Ham, S.-Y., Kim, Y.-J. & Lee, D.-H., 2008. Leaching characteristics of PCDDs/DFs and dioxin-like PCBs from landfills containing municipal solid waste and incineration residues. Chemosphere, 70(9), pp.1685–93.
34 Introduction
Hermosilla, D., Cortijo, M. & Huang, C.P., 2009. Optimizing the treatment of landfill leachate by conventional Fenton and photo-Fenton processes. The Science of the total environment, 407(11), pp.3473–81.
Hites, R.A., 2011. Dioxins: an overview and history. Environmental science & technology, 45(1), pp.16–20.
Holt, E. et al., 2008. Assessing dioxin precursors in pesticide formulations and environmental samples as a source of octachlorodibenzo-p-dioxin in soil and sediment. Environmental science & technology, 42(5), pp.1472–8.
Holt, E. et al., 2012. Formation of dioxins during exposure of pesticide formulations to sunlight. Chemosphere, 88(3), pp.364–70.
Hong, J. et al., 2003. Identification of electrical degradation products of 4- chlorophenol in water. Analytical sciences : the international journal of the Japan Society for Analytical Chemistry, 19(4), pp.537–42.
Hong, J. et al., 2000. Identification of photolytical transformation products of pentachlorophenol in water. Analytical Sciences, 16(6), pp.621–626.
Kapałka, A., Fóti, G. & Comninellis, C., 2010. Basic Principles of the Electrochemical Mineralization of Organic Pollutants for Wastewater Treatment. In C. Comninellis & G. Chen, eds. Electrochemistry for the Environment. Springer New York Dordrecht Heidelberg London, pp. 1–23.
Kapałka, A., Fóti, G. & Comninellis, C., 2007. Kinetic modelling of the electrochemical mineralization of organic pollutants for wastewater treatment. Journal of Applied Electrochemistry, 38(1), pp.7–16.
Karci, A., 2014. Degradation of chlorophenols and alkylphenol ethoxylates, two representative textile chemicals, in water by advanced oxidation processes: the state of the art on transformation products and toxicity. Chemosphere, 99, pp.1–18.
Kavitha, V. & Palanivelu, K., 2003. Degradation of 2-Chlorophenol by Fenton and Photo-Fenton Processes—A Comparative Study. Journal of Environmental Science and Health, Part A, 38(7), pp.1215–1231.
Kim, H. et al., 2002. Removal efficiency and homologue patterns of dioxins in drinking water treatment. Water Research, 36(19), pp.4861–4869.
35 Chapter 1
Koistinen, J. et al., 2007. Bioaccumulation, bioavailability and environmental fate of chlorophenol impurities, polychlorinated hydroxydiphenylethers and their methoxy analogues. Chemosphere, 68(7), pp.1382–91.
Kulkarni, P.S., Crespo, J.G. & Afonso, C.A.M., 2008. Dioxins sources and current remediation technologies--a review. Environment international, 34(1), pp.139–53.
Laine, D.F. & Cheng, I.F., 2007. The destruction of organic pollutants under mild reaction conditions: A review. Microchemical Journal, 85(2), pp.183–193.
Lee, J.E. et al., 2004. Theoretical Study on the Reaction of OH Radicals with Polychlorinated Dibenzo- p -dioxins. The Journal of Physical Chemistry A, 108(4), pp.607–614..
Lopez, A. et al., 2004. Fenton’s pre-treatment of mature landfill leachate. Chemosphere, 54(7), pp.1005–1010.
Lopez, A. et al., 2005. Temperature activated degradation (mineralization) of 4-chloro-3-methyl phenol by Fenton’s reagent. Chemosphere, 59(3), pp.397– 403.
Lores, M. et al., 2005. Confirmation of the formation of dichlorodibenzo-p- dioxin in the photodegradation of triclosan by photo-SPME. Analytical and bioanalytical chemistry, 381(6), pp.1294–8.
Marinković, N. et al., 2010. Dioxins and human toxicity. Arhiv za higijenu rada i toksikologiju, 61(4), pp.445–53.
Marselli, B. et al., 2003. Electrogeneration of Hydroxyl Radicals on Boron- Doped Diamond Electrodes. Journal of The Electrochemical Society, 150(3), p.D79.
Martínez-Huitle, C. a. & Brillas, E., 2009. Decontamination of wastewaters containing synthetic organic dyes by electrochemical methods: A general review. Applied Catalysis B: Environmental, 87(3-4), pp.105–145.
Methatham, T., Lu, M.-C. & Ratanatamskul, C., 2011. Removal of 2,4- dichlorophenol as herbicide’s by-product by Fenton’s reagent combined with an electrochemical system. Desalination and Water Treatment, 32(1-3), pp.42–48.
36 Introduction
Mezcua, M. et al., 2004. Evidence of 2,7/2,8-dibenzodichloro-p-dioxin as a photodegradation product of triclosan in water and wastewater samples. Analytica Chimica Acta, 524(1-2), pp.241–247.
Mitrou, P.., Dimitriadis, G. & Raptis, S.., 2001. Toxic effects of 2,3,7,8- tetrachlorodibenzo-p-dioxin and related compounds. European Journal of Internal Medicine, 12(5), pp.406–411.
Mukerjee, D., 1998. Health impact of polychlorinated dibenzo-p-dioxins: A critical review. In Journal of the Air and Waste Management Association. Air & Waste Management Assoc, pp. 157–165.
Muna, G.W., Tasheva, N. & Swain, G.M., 2004. Electro-oxidation and Amperometric Detection of Chlorinated Phenols at Boron-Doped Diamond Electrodes: A Comparison of Microcrystalline and Nanocrystalline Thin Films. Environmental Science & Technology, 38(13), pp.3674–3682.
Munoz, M. et al., 2011. Assessment of the generation of chlorinated byproducts upon Fenton-like oxidation of chlorophenols at different conditions. Journal of hazardous materials, 190(1-3), pp.993–1000.
NATO (North Atlantic Treaty organization), 1988. International Toxicity Equivalency Factor (I-TEF) Method of Risk Assessment for Complex Mixtures of Dioxins and Related Compounds. Pilot study on international information exchange on dioxins and related compounds. Committee on Challenges of Modern Society.
Öberg, L.G. & Rappe, C., 1992. Biochemical formation of PCDD/Fs from chlorophenols. Chemosphere, 25(1-2), pp.49–52.
Öberg, L.G. et al., 1993. De novo formation of PCDD/Fs in compost and sewage sludge—a status report. Organohalogen Compounds, 11, pp. 297–302.
Olaniran, A.O. & Igbinosa, E.O., 2011. Chlorophenols and other related derivatives of environmental concern: properties, distribution and microbial degradation processes. Chemosphere, 83(10), pp.1297–306.
Oller, I., Malato, S. & Sánchez-Pérez, J. a, 2011. Combination of Advanced Oxidation Processes and biological treatments for wastewater decontamination--a review. The Science of the total environment, 409(20), pp.4141–66.
37 Chapter 1
Oman, C.B. & Junestedt, C., 2008. Chemical characterization of landfill leachates--400 parameters and compounds. Waste management (New York, N.Y.), 28(10), pp.1876–91.
Oturan, N., Panizza, M. & Oturan, M.A., 2009. Cold incineration of chlorophenols in aqueous solution by advanced electrochemical process electro-Fenton. Effect of number and position of chlorine atoms on the degradation kinetics. The journal of physical chemistry. A, 113(41), pp.10988– 93.
Ozkaya, B., 2005. Chlorophenols in leachates originating from different landfills and aerobic composting plants. Journal of hazardous materials, 124(1- 3), pp.107–12.
Palmas, S. et al., 2007. Influence of the operating conditions on the electrochemical disinfection process of natural waters at BDD electrodes. Journal of Applied Electrochemistry, 37(11), pp.1357–1365.
Panizza, M., 2010. Importance of Electrode Material in the Electrochemical Treatment of Wastewater containing Organic Pollutants. In C. Comninellis & G. Chen, eds. Electrochemistry for the Environment. Springer New York Dordrecht Heidelberg London, pp. 25–54.
Panizza, M. & Cerisola, G., 2005. Application of diamond electrodes to electrochemical processes. Electrochimica Acta, 51(2), pp.191–199.
Pera-Titus, M. et al., 2004. Degradation of chlorophenols by means of advanced oxidation processes: a general review. Applied Catalysis B: Environmental, 47(4), pp.219–256.
Pignatello, J.J., Oliveros, E. & MacKay, A., 2006. Advanced Oxidation Processes for Organic Contaminant Destruction Based on the Fenton Reaction and Related Chemistry. Critical Reviews in Environmental Science and Technology, 36(1), pp.1–84.
Polcaro, A.M. et al., 2004. Electrochemical degradation of diuron and dichloroaniline at BDD electrode. Electrochimica Acta, 49(4), pp.649–656.
Primo, O., Rivero, M.J. & Ortiz, I., 2008. Photo-Fenton process as an efficient alternative to the treatment of landfill leachates. Journal of hazardous materials, 153(1-2), pp.834–42.
38 Introduction
Rappe, C., 1994. Dioxin, patterns and source identification. Fresenius’ Journal of Analytical Chemistry, 348(1-2), pp.63–75.
Renou, S. et al., 2008. Landfill leachate treatment: Review and opportunity. Journal of hazardous materials, 150(3), pp.468–93.
Ruiz, M.L., 2007. Determinación y evaluación de las emisiones de dioxinas y furanos en la producción de cemento en España. Universidad Complutense de Madrid.
Sanchez-Prado, L. et al., 2006. Monitoring the photochemical degradation of triclosan in wastewater by UV light and sunlight using solid-phase microextraction. Chemosphere, 65(8), pp.1338–47.
Sankoda, K. et al., 2011. Identification of triclosan intermediates produced by oxidative degradation using TiO2 in pure water and their endocrine disrupting activities. Bulletin of environmental contamination and toxicology, 86(5), pp.470–5.
Sedlak, D.L. & Schnoor, J.L., 2013. The challenge of water sustainability. Environmental science & technology, 47(11), p.5517.
Shannon, M.A. et al., 2008. Science and technology for water purification in the coming decades. Nature, 452(7185), pp.301–10.
Sim, W.-J. et al., 2009. Distribution and formation of chlorophenols and bromophenols in marine and riverine environments. Chemosphere, 77(4), pp.552–8.
Sirés, I. et al., 2014. Electrochemical advanced oxidation processes: today and tomorrow. A review. Environmental science and pollution research international.
Srogi, K., 2007. Levels and congener distributions of PCDDs, PCDFs and dioxin-like PCBs in environmental and human samples: a review. Environmental Chemistry Letters, 6(1), pp.1–28.
Stockholm Convention on Persistent Organic Pollutants, 2001. Stockholm Convention, Geneva, Switzerland < http://www.pops.int/documents/convtext/convtext_en.pdf>.
Stringer, R. & Johnston, P., 2001. Chlorine and the Environment: An Overview of the Chlorine Industry. Kluwer Academic Publishers. Dordrecht.
39 Chapter 1
Umar, M., Aziz, H.A. & Yusoff, M.S., 2010. Trends in the use of Fenton, electro-Fenton and photo-Fenton for the treatment of landfill leachate. Waste management (New York, N.Y.), 30(11), pp.2113–21.
UNEP, 2003. United Nations Environment Programme. Available at http://www.unep.org/. Accessed May 2014.
UNICEF, 2014. United Nations International Children's Emergency Fund. Available at http://www.unicef.org. Accessed May 2014.
Unites States Environmental Protection Agency. Health and Environmental Effects Profile No. 135; USEPA: Washington DC, 1980.
Vilhunen, S. & Sillanpää, M., 2010. Recent developments in photochemical and chemical AOPs in water treatment: a mini-review. Reviews in Environmental Science and Bio/Technology, 9(4), pp.323–330.
Vollmuth, S. & Niessner, R., 1995. Degradation of PCDD, PCDF, PAH, PCB and chlorinated phenols during the destruction-treatment of landfill seepage water in laboratory model reactor (UV, Ozone, and UV/Ozone). Chemosphere, 30(12), pp.2317–2331.
Vollmuth, S., Zajc, A. & Niessner, R., 1994. Formation of Polychlorinated Dibenzo-p-dioxins and Polychlorinated Dibenzofurans during the Photolysis of Pentachlorophenol-Containing Water. Environmental science & technology, 28(6), pp.1145–9.
Watanabe, S., Kitamura, K. & Nagahashi, M., 1999. Effects of dioxins on human health: a review. Journal of epidemiology / Japan Epidemiological Association, 9(1), pp.1–13.
Weber, R. et al., 2008. Dioxin- and POP-contaminated sites--contemporary and future relevance and challenges: overview on background, aims and scope of the series. Environmental science and pollution research international, 15(5), pp.363–93.
Weber, R., 2007. Relevance of PCDD/PCDF formation for the evaluation of POPs destruction technologies--review on current status and assessment gaps. Chemosphere, 67(9), pp.S109–17.
40 Introduction
Weber, R. et al., 2011. Review Article: Persistent organic pollutants and landfills - a review of past experiences and future challenges. Waste management & research : the journal of the International Solid Wastes and Public Cleansing Association, ISWA, 29(1), pp.107–21.
Wiszniowski, J. et al., 2006. Landfill leachate treatment methods: A review. Environmental Chemistry Letters, 4(1), pp.51–61.
Xu, X.H. et al., 2003. 2-chlorophenol oxidation kinetic by photo-assisted Fenton process. Journal of Environmental Sciences, 15(4), pp.475–481.
Zazo, J. a. et al., 2011. Intensification of the Fenton Process by Increasing the Temperature. Industrial & Engineering Chemistry Research, 50(2), pp.866– 870.
41
Materials and Methods
2.1. Chemical reagents
Leachate samples used in this work were collected from a municipal solid waste landfill located in Cantabria, Spain, from June 2010 to November 2011, and kept under refrigeration until their analysis. On the other hand, the chemicals used in Fenton and electrochemical experiments as well as the standards utilized for the analytical measurements are listed in Table 2.1. On the other hand, the specific reagents applied for PCDD/Fs analysis are detailed in Table 2.2.
Table 2.1. Summary of the chemical reagents utilized in this research.
Name Chemical formula Purity Supplier
Acetic acid C2H4O2 100% Merck
Acetonitrile C2H3N 99.9% Panreac
Ammonium hydroxide NH4OH 25% Panreac Ammonium Iron (II) (NH ) Fe(SO ) ·6H O 98% Panreac sulphate (FAS) 4 2 4 2 2
Ammonium Molibdate (NH4)6Mo7O24·4H2O Panreac
Benzoquinone C6H4O2 98% Merck
Bisphenol A C15H16O2 97% Sigma-Aldrich
Catechol C6H6O2 98% Merck
2-Chlorobenzoquinone C8H3ClO3 95% Sigma-Aldrich
2-Chlorophenol (2-CP) C6H5ClO 99% Sigma-Aldrich
2,4-dichlorophenol C6H4Cl2O 99% Sigma-Aldrich
2,6-dichlorophenol C6H4Cl2O 99% Sigma-Aldrich 2,6- C H Cl O 98% Sigma-Aldrich dichlorobenzoquinone 6 2 2 2
Formic acid CH2O2 98% Panreac
Fumaric acid C4H4O4 99% Panreac
Hydrogen peroxide H2O2 40% Merck
Hydroquinone C2H6O2 99% Panreac IC standard 0.5 g L-1 Panreac - - -1 Inorganic Ions Cl , NO2 1000 mg L Merck - 2- NO3 , SO4 Iron sulphate 7-hydrate FeSO4 ·7H2O 99.5% Panreac
Maleic acid C4H4O4 99% Sigma-Aldrich
Malonic acid C3H4O4 99% Merck
45
Chapter 2
Table 2.1. Summary of the chemical reagents utilized in this research (Cont.).
Name Chemical formula Purity Supplier
Methanol CH4O Suprasolv Merck
Methyl tertiary-butyl C5H12O 99% Riedel-de ether (MTBE) Haën
n-Butyl C10H15NO2S 99% Sigma-Aldrich benzenesufonamide (NBBS)
Nitric acid HNO3 65% Merck Oxalic acid H₂C₂O₄ 99% Panreac
Phenol C6H6O 99% Panreac
Phosphoric acid H3PO4 85% Panreac
Potasium dichromate K2Cr2O7 0.1 N Merck
potassium hydrogen C8H5KO4 Nacalai phthalate Tesque Potasium iodide KI 99% Panreac -1 Silver sulphate solution Ag2SO4 10 g L in Panreac
H2SO4
Sodium bisulfite NaHSO3 40% w/v Panreac
Sodium carbonate Na2CO3 99.5% Sigma-Aldrich Sodium chloride NaCl 99.5% Panreac Sodium hydroxide NaOH 32% Merck
Sodium thiosulfate Na2S2O3 0.1 N Panreac
Sodium sulfate Na2SO4 99% Merck
Sulfuric acid H2SO4 98% Merck
2,3,6-TCP C6H3Cl3O 98% Sigma-Aldrich TOC standard 1 g L-1 Panreac
Table 2.2. Reagents used for PCDD/Fs analysis.
Reagent Supplier Acetone (Suprasolv) Merck Dichloromethane (Unisolv) Merck Ethyl acetate (Pestinorm) Pestinorm Toluene (Unisolv) Merck EPA 1613 CSL-CS5: calibration standard solutions Wellington EPA 1613 ISS: internal standard spiking solution Wellington EPA 1613 LCS: labelled compound stock solution Wellington EPA 1613 PAR: precision and recovery stock solution Wellington
46 Materials and Methods
2.2. Electrochemical oxidation experiments
An schematic diagram of the experimental set-up used in the electrochemical experiments is depicted in Figure 2.1. Besides, a laboratory photograph of this experimental system is shown in Figure 2.2.
Electrochemical cell
Magnetic pump
Feed tank Power Supply
Refrigeration system
Figure 2.1. Schematic diagram of the electrochemical oxidation experimental set-up.
Electrochemical cell
Magnetic pump
Feed tank
Figure 2.2. Experimental set-up employed in the electrochemical oxidation experiments.
47 Chapter 2
The experimental set-up depicted in Figure 2.1 and Figure 2.2 consisted of a feed tank (2 L of capacity) in which the sample was stored before it was recirculated through the electrolytic cell by means of a magnetic pump. To remove the heat released due to the continuous solution pumping and the electrical resistance within the electrochemical cell, the system was refrigerated. The electrochemical cell used that is shown in Figure 2.3, is a single compartment electrolytic flow-cell (DiaCell) manufactured by Adamant Technologies (Switzerland).
Figure 2.3. DiaCell type electrochemical cell (Adamant technologies).
The cell was comprised of two circular electrodes disposed in parallel with a geometric area of 0.007 m2 each and an electrode gap of 5 mm. The anode is made on boron-doped diamond (BDD) on silicon and the cathode is made on stainless steel. Experiments were ran in discontinuous mode under galvanostatic mode at a current density (J) equal to 900 A m-2 and the working temperature was maintained at 20ºC by means of the refrigeration bath. Once the solution to be treated was loaded in the feed tank, it was pumped through the experimental system selecting the target flow-rate. The refrigeration system was then turned on and when the solution temperature was set at 20ºC the power supply was connected selecting the desired intensity, in this case 6.3 A (equivalent to J = 900 A m-2). Samples were withdrawn at regular time intervals and analyzed according to methods described below. The electrochemical experiments were developed under the experimental conditions summarized in Table 2.3.
48 Materials and Methods
Table 2.3. Experimental conditions of electrochemical experiments.
2-CP Landfill leachates
experiments experiments
[2-CP]o (mM) 15.56 - J (A m-2) 900 Volume (L) 1 Tª (ºC) 20 Flow-rate (L min-1) 9 11 NaCl: 56.3 Electrolyte (mM) - Na2SO4 : 21.1 Sample conductivity 7.5 14.6 (mS cm-1)
2.3. Fenton oxidation experiments
Fenton oxidation experiments were performed in batch mode in glass reactors magnetically stirred (700 rpm) with a heating magnetic stirrer with temperature control (RET basic IKA safety control). Batch experiments were carried out adjusting the initial pH of the solutions to be treated to 3.0 with nitric acid (HNO3) in the case of 2-CP samples or sulphuric acid (H2SO4) in the case of leachate samples. Iron sulphate 7-hydrate (FeSO4·7H2O) was added to the solution to yield the desired Fe2+ dose. The Fenton reaction was initiated by adding the corresponding amount of hydrogen peroxide (H2O2) to the solution. Samples were withdrawn from the reaction medium when the oxidation time was reached and the excess of H2O2 was removed by sodium bisulfite (NaHSO3), after tritation with sodium thiosulfate (Na2S2O3). Experiments were done under the operating conditions listed in Table 2.4.
49 Chapter 2
Table 2.4. Experimental conditions of Fenton experiments.
2-CP experiments 20% of the stoichiometric t 100% of the stoichiometric H O (mM) 2 2 40.44 40.44 40.44 202.28 202.28 202.28 Fe+2 (mM) 0.18 0.18 0.18 7.22 7.22 7.22 Tª (ºC) 20 20 70 20 70 70 NaCl (mM) - 56.34 - - - 56.34 Landfill leachates experiments
H2O2/COD (w/w) 5.86 Fe+2/COD (w/w) 0.29 Tª (ºC) 20
2.4. Analytical measurements
2.4.1. Analysis of chemical oxygen demand
The followed procedure to analyze the chemical oxygen demand (COD), which is based on the oxidation of the organic matter by a boiling mixture of potassium dichromate (K2Cr2O7) and silver sulfate (Ag2SO4), is described by the
Standard Methods 5220 B (APHA 1998). Firstly, 1.5 mL of K2Cr2O7 ( 0.04 M), 3.5 -1 mL of Ag2SO4 (10 g L ) in sulphuric acid and 2.5 mL of sample were added into a vial, mixed thoroughly and digested at 148ºC during 2 h. The color of the solution changed from orange to green. After digestion and cooling up to room temperature, the excess of K2Cr2O7 was determined by tritation against 0.02 M ammonium Iron (II) sulphate ((NH4)2Fe(SO4)2·6H2O) (FAS) using ferroin as indicator. The dichromate consumed by the sample is equivalent to the amount of O2 required to oxidize the organic matter. The end point of the titration was taken as the first sharp color change from blue-green to reddish-brown that persisted for 1 min or longer. The COD was calculated according to:
(A -B) M 8000 COD (mg O L-1 ) = 2 mL of sample
Where A is the FAS volume used for the blank (mL), B is the FAS volume used for the sample (mL) and M is the FAS molarity.
50 Materials and Methods
2.4.2. Analysis of total organic carbon
Total organic carbon (TOC) was analyzed using a TOC analyzer Shimadzu TOC-V CPH with auto-sampler ASI-V (Figure 2.4) according to the international regulation UNE-EN 1484:1998.
Figure 2.4. TOC analyzer Shimadzu TOC-V CPH with auto-sampler ASI-V.
For this propose, 100 µL of the sample were automatically injected into the oven where they were combusted to CO2. The produced CO2 was carried by means of an air current through a bubbler containing H3PO4 up to the non- dispersive infrared (NDIR) detector, where the CO2 was measured. The integration of this measure during the time gives the total carbon (TC) present in the sample. The inorganic carbon (IC) value is obtained from the H3PO4 acid bubbler. TOC is obtained as the difference between TC and IC. The curve calibrations were made with potassium hydrogen phthalate (C8H5KO4) for TC and sodium carbonate (Na2CO3) for IC. The operational conditions used for the analysis of TOC are summarized in Table 2.5.
Table 2.5. Operational conditions used in the analysis of TOC.
Nº of measurements 2/3 or 3/5 Injection volume 50 µL Oven temperature 680ºC Detection range 0-1000 mg L-1
51 Chapter 2
2.4.3. Analysis of organic acids and inorganic ions
- - The analysis of chloride (Cl ), chlorate (ClO3 ) and aliphatic organic acids including acetic, formic, malonic, maleic, oxalic and fumaric acids was carried out by means of ion chromatography with anionic suppression using the chromatograph Dionex ICS-1100 with a conductivity cell detector (ASR-ULTRA model) (Figure 2.5).
Figure 2.5. Ion chromatograph Dionex ICS-1100 equipped with a conductivity cell detector (ASR-ULTRA model).
The system was run using a 25μL sampling loop. Separation occurred with an IonPac AS9-HC 4mm anion separation column at 30°C. The suppressing system eliminates the conductivity due to the produced anions from the mobile phase using a column model P/N 53946. The samples were supplied to the system by means of a Dionex AS40 auto-sampler. The eluent was a 9 mM -1 Na2CO3 solution at 1 mL min . Anions were retained in the column and eluted at different retention times being registered in the detector. The signal from the detector was transformed into concentration by means of the calibration curves of the measured compounds. The entire system was controlled by Chromeleon 6.8 software.
52 Materials and Methods
2.4.4. Analysis of ammonium nitrogen concentration
+ The ammonium nitrogen (N-NH4 ) concentration was determined by distillation and subsequent tritation according to the analytical procedure 4500 from Standards Methods (APHA 1998). The distillation was carried out in alkaline medium using a Buchi UK-355 distillator (Figure 2.6).
Figure 2.6. Distillation system Buchi UK-355 distillator.
The distilled ammonium, collected in H3PO4, was tritated with HCl using a mixture of methyl red and methylene blue indicator. The end point of the titration was taken as the first sharp color change from green to violet that + persisted for 1 min or longer. The N-NH4 was calculated according to:
+ -1 (VHCl NHCl )PMN1000 N- NH4 (mg L ) = Vsample
Where VHCl is the HCl volume spent (mL), NHCl is the HCl normality and PMN is the molar mass of N.
53 Chapter 2
2.4.5. Analysis of 2-chlorophenol and related aromatic compounds
2-CP and aromatic reaction intermediates were measured by high performance liquid chromatography (HPLC). 2-CP, catechol, hydroquinone, p- benzoquinone and phenol, were analyzed using a Waters 2695 HPLC coupled to a photo diode array (PDA) detector (Figure 2.7a). Separation occurred with a
Gemini C-18 separation column (L = 150 mm; Øi = 3 mm) using H2SO4 4 mM as mobile phase with a flowrate of 1 mL min-1. The selected wavelengths were 210 and 245 nm and the column temperature was kept at 50ºC. On the other hand, 2,4-dichorophenol (2,4-DCP), 2,6-dichlorophenol (2,6-DCP), 2,3,6- trichlorophenol (2,3,6-TCP), 2-chlorobenzoquinone and 2,6- dichlorobenzoquinone were measured using an Agilent series 1100 HPLC coupled to a PDA detector (Figure 2.7b). Separation occurred with a Gemini C-
18 column (L = 150 mm; Øi = 3 mm) and 20% Acetonitrile-80% H3PO4 0.01 M as mobile phase at 1 mL min-1 with a flow gradient up to 80% Acetonitrile–20%
H3PO4 in 2.5 min. The selected wavelengths were 210, 254 and 280 nm and the column temperature was 30ºC. The signal from the detector was transformed into concentration by means of the calibration curves made with standards of the different compounds.
Figure 2.7. a) HPLC Waters 2695; b) HPLC Agilent series 1100.
54 Materials and Methods
2.4.6. Qualitative screening of organics in the advanced oxidation of 2-chlorophenol
Qualitative screening (non-target analysis) for intermediate organic compounds formed during the advanced oxidation treatment of 2-chlorophenol solutions was performed by gas chromatography-mass spectrometry (GC-MS). Samples (100 mL) were extracted twice with 50 ml of dichloromethane using a separatory funnel. The organic extract was concentrated up to 1 mL using a rotatory evaporator in a first step (Laborota 4000) and a N2 stream in a second step. Finally, the concentrated extract was analyzed using a GC-MS Shimadzu QP2010 equipped with an auto-sampler (Figure 2.8).
Figure 2.8. GC-MS Shimadzu QP2010.
The analytes were separated using a HP5 MS column (L = 30m; Øi = 0.25 mm) with a film thickness of 0.1 mm. The temperature program of the GC–MS oven was 60 ºC, hold 3 min, rate 10 ºC min-1 to 270ºC, hold 2.5 min. He was used as a carrier gas at a flow rate of 1 mL min-1. A volume sample of 2µL was injected in splitless mode and the injector temperature was set at 285ºC. The mass spectrometer (MS) was operated in the electron impact ionization mode (70 eV). The transfer line and ion source temperatures were 290 and 230ºC, respectively. Data acquisition was in full scan mode with a range from m/z 35 to 400. Confirmation of all structural assignments for the identified compounds was made using the NIST08 spectra library.
55 Chapter 2
2.4.7. Qualitative screening of organics in leachates
Extraction and concentration
Qualitative screening (non-target analysis) for organic compounds in landfill leachates was performed by GC-MS after previous solid phase extraction (SPE). The organic compounds were extracted from the leachate matrix using a SPE procedure with Oasis HLB cartridges (Waters) (Figure 2.9a). The cartridges were first conditioned with 3 mL of methyl tertiary-butyl ether (MTBE), 3 mL of methanol and 3 mL of ultrapure water. Afterwards, leachate samples (500 mL) at pH 2 were loaded onto the cartridges and filtered dropwise with a vacuum pump. The cartridges were then washed with 3 mL of 40% methanol in ultrapure water to remove organic interferences, re-equilibrated with 3 mL of ultrapure water, washed with 3 mL of 10% methanol/2% NH4OH to remove humic interferences. Finally the organic analites were eluted with 6 mL of 10% methanol/90% MTBE and this extract was evaporated to 200 µL by means of a
N2 stream using a TurvoVap evaporator with a temperature bath at 30ºC (Figure 2.9b).
a b
Figure 2.9. a) SPE device; b) TurvoVap evaporator.
Qualitative analysis
The concentrated extract was then analyzed using a GC-MS Agilent 6890N equipped with an auto-sampler (Figure 2.10). The analytes were separated using a HP5 MS column (L = 30m; Øi = 0.25 mm) with a film thickness of 0.1 mm. The temperature program of the GC–MS oven was 50 ºC, hold 2 min, rate 8 ºC min-1 to 350ºC, hold 30s. He was used as a carrier gas at a flow rate of 1 mL min-1. The injector temperature was set at 270ºC, and the MS was operated in the electron impact ionization mode (70 eV). The transfer line and ion source
56 Materials and Methods
temperatures were 290 and 220ºC, respectively. Scan runs were made with a range from m/z 35 to 400. Confirmation of all structural assignments for the identified compounds was made using the NIST08 spectra library as well as analytical standards for some compounds. A match percentage was obtained by comparing the mass spectrum of a peak with that of a known compound from the library. The compound was deemed identified and reported if the match percentage was higher than 70%.
Figure 2.10. GC-MS Agilent 6890 N.
Quality control
The reliability of the analytical methodology was assured by means of its application to ultrapure water samples spiked with n-butyl benzenesulfonamide (NBBS) and bisphenol A (BPA) standards at similar concentrations to those found in landfill leachates. External calibration was used for their quantitation The average recoveries were in the range of 65% for NBBS and 107% for BPA. Blanks covering the whole sample preparation methodology were also carried out.
57 Chapter 2
2.5. Analysis of PCDD/Fs
PCDD/Fs were determined according to the Standard Method U.S. EPA 1613 by isotope dilution and high resolution gas chromatography/high resolution mass spectrometry (HRGC-HRMS) (U.S. EPA 1613 1994). This method consists of successive rigorous stages, as is summarized in Figure 2.11, that include extraction, concentration and purification steps in order to prepare the samples previous to their final analysis by HRGC-HRMS. A detailed description of the analytical methodology applied for the determination of PCDD/Fs is shown below.
Sample Concentration (0,5-1 L)
H2SO4 treatment L-L extraction L-L extraction with with DCM 13C Labelled n-hexane Standard
Concentration Concentration and filtration
HRGC - HRMS Clean-up
Figure 2.11. Analytical methodology scheme for the determination of PCDD/Fs.
Extraction and pre-treatment of samples
Samples (0.5-1 L) are spiked with 10 µL of a 15 13C-labeled PCDD/Fs solution (EPA 1613 LCS) dissolved in acetone and mixed by careful shaking. Afterwards, PCDD/Fs are extracted from the aqueous phase with three portions of 60 mL aliquots of dichloromethane using a 2 L separatory funnel. The organic extract is concentrated in a rotatory evaporator Büchi R-210 (Figure 2.12) near to dryness at Tª = 35ºC and P = 700-750 mbar. At this point, the extract is rinsed with dichloromethane and concentrated again near to dryness in triplicate.
58 Materials and Methods
Figure 2.12. Rotatory evaporator Büchi R-210.
Samples with high quantity of co-extracted organic matter (most cases in this work) needed to be treated with H2SO4. For this propose, the evaporated extracts are transferred to n-hexane, treated with H2SO4 (100 mL) and extracted with 75 mL of n-hexane in duplicate. The organic phase is then dried with
Na2SO4 and concentrated in the rotatory evaporator to approximately 1-2 mL (n-hexane: Tª = 35ºC, P = 600-500 mbar). Next, the concentrated extract is transferred to 12 mL of n-hexane and filtered through a 0.45 µm PTFE filter.
Extract cleanup
The purification of the filtered extracts is performed by means of solid- liquid adsorption chromatography using the automated system Power-PrepTM (Fluid Management Systems Inc., Waltham) shown in Figure 2.13
59 Chapter 2
Figure 2.13. Purification automated system Power-PrepTM.
The automated cleanup system Power-Prep is based on the sequential use of multilayer silica, basic alumina and PX-21 carbon adsorbents respectively, pre-packed in disposable Teflon and sealed columns (Technospec). The filtered extract is loaded and pumped through the columns using next solvents and their mixtures: n-hexane, n-hexane/dichloromethane (98/2 v/v), n- hexane/dichloromethane (50/50 v/v), toluene/ethyl acetate (50/50 v/v) and toluene. Organic interferences are eliminated through the different columns and the purified extract is collected from the carbon column in 75 mL of toluene. Toluene is evaporated firstly by means of the rotatory evaporator (toluene: Tª = 50ºC, P = 150-87 mbar) and once the extract is near to dryness, it is transferred into a vial to be concentrated to dryness under a nitrogen stream.
HRGC-HRMS analysis
Purified samples are analyzed by the Chromatography Service (SERCROM) of the University of Cantabria. Before the chromatographic analysis, internal syringe standards (EPA 1613 ISS) were added to the sample. The analysis is carried out by HRGC-HRMS on a TRACE GC UltraTM gas chromatograph equipped with a split/splitless injector (Thermo Electron S.p.A.) and a DB-5 MS fused silica
60 Materials and Methods
capillary column (J&W Scientific). The initial temperature of the column, 120ºC, is kept constant for 2 min and then it is increased sequentially in 3 steps to 210ºC, 230ºC and 310ºC at 15, 1 and 3ºC/min respectively. The column is connected through a heated transfer line kept at 270°C to a DFS high-resolution magnetic sector mass spectrometer with a BE geometry, Figure 2.14 (Thermo Fisher Scientific). A positive electron ionization (EI+) mode with ionization energy of 45 eV is used in the source and its temperature is set at 270ºC. The mass spectrometer is operated in the SIM mode at 10,000 resolution power (10% valley definition). Detection limits are calculated as the concentration values that gave instrumental responses within a signal-to-noise ratio of 3.
Figure 2.14. DFS double focusing sector system (Thermo Fisher Scientific).
PCDD/Fs quantification
Quantitative determination is carried out by the isotopic dilution method. Relative response factors (RRFs), obtained from the calibration curve by analyzing CS-1 to CS-5 standard solution mixtures, are used to determine the target compounds concentration in the samples. The recoveries of labelled standards are calculated using a mixture of two labelled PCDD (ISS) that were added to the samples before the chromatographic analysis. Final results were expressed both in pg L-1 and in toxic equivalents (TEQ), both pg I-TEQ L-1 and pg WHO-TEQ L-1.
61 Chapter 2
2.6. PCDD/Fs analytical methodology setup
Before to PCDD/Fs analysis, it was necessary to set up the analytical methodology for the determination of PCDD/Fs in aqueous matrices. For this proposal, analyses of ultrapure water spiked with PCDD/Fs standards, in order to verify the accuracy, the recoveries and the absence of contamination during the analysis were carried out.
In order to achieve this goal, several precision and recovery experiments were performed using ultrapure water spiked with LCS standard (1-2 ng), which contains the PCDD/Fs isotopically labelled with carbon 13, and PAR standard (20-200 pg) containing the native PCDD/Fs to be determined in the samples. Besides, blank experiments were carried out with ultrapure water spiked with the LCS standard (1-2 ng). The main goals of the addition of the LCS standard to the sample are to quantify the native PCDD/Fs contained in the sample by the isotope dilution method, as well as to check the losses during the whole sample preparation methodology for their final analysis by HRGC-HRMS.
In the precision and recovery experiments, the relative standard deviations (RSD) in the analyses of ultrapure water spiked with a known amount of native PCDD/Fs ranging from 40 pg L-1 for 2,3,7,8-TCDD to 400 pg L-1 for OCDD, were in the range of 10.68% for OCDF to 17.49% for 1,2,3,7,8,9-HxCDD (Table 2.6). Besides, the average recoveries of labelled PCDD/Fs ranged from 75.6% for 2,3,7,8-TCDF13C to 93% for 1,2,3,7,8-PeCDD13C, all within the EPA method established limits (Table 2.7). Therefore, the ability to generate acceptable precision and recoveries in the analysis of PCDD/Fs was verified. On the other hand, results from blank experiments showed average recoveries of labelled PCDD/Fs in the range of 76.8% for 2,3,7,8-TCDF13C to 91.2% for1,2,3,4,7,8,9- HpCDF13C. In addition, the concentrations of native PCDD/Fs were in most cases below the detection limits verifying the absence of contamination. All results met the acceptance criteria established by the EPA 1613 Standard Method and therefore the analytical methodology setup was successfully achieved.
62 Materials and Methods
Table 2.6. Concentration of native PCDD/Fs in precision and recovery experiments and in method blanks.
Precision and recovery Blank experiments experiments Average Average (DL) concentration RSD (%) concentration -1 (pg L ) (pg L-1) (pg L-1) 2,3,7,8-TCDD 42.97 15.37
63 Chapter 2
Table 2.7. Recoveries of labelled PCDD/Fs in precision and recovery experiments and in method blanks.
Precision and recovery Blank experiments experiments Average Average RSD RSD (%) recovery (%) recovery (%) (%) 2,3,7,8-TCDD13C 91.60 9.57 86.60 16.90 1,2,3,7,8-PeCDD13C 93.00 7.41 81.60 17.81 1,2,3,4,7,8-HxCDD13C 90.60 11.75 86.20 16.15 1,2,3,6,7,8-HxCDD13C 87.80 18.12 85.00 16.26 1,2,3,7,8,9-HxCDD13C 100.00 0.00 100.00 0.00 1,2,3,4,6,7,8-HpCDD13C 90.60 11.03 88.00 16.87 OCDD13C 81.60 15.80 85.50 14.37 2,3,7,8-TCDF13C 75.60 10.85 76.80 14.83 1,2,3,7,8-PeCDF13C 89.40 8.69 85.20 20.15 2,3,4,7,8-PeCDF13C 91.00 4.04 88.00 18.50 1,2,3,4,7,8-HxCDF13C 87.40 17.36 84.60 13.57 1,2,3,6,7,8-HxCDF13C 85.00 20.73 82.40 15.60 1,2,3,7,8,9-HxCDF13C 89.00 6.65 86.40 12.07 2,3,4,6,7,8-HxCDF13C 86.60 12.61 85.20 10.96 1,2,3,4,6,7,8-HpCDF13C 79.60 19.64 81.20 14.24 1,2,3,4,7,8,9-HpCDF13C 91.40 13.12 91.20 16.44
2.7. Quality control in the analysis of PCDD/Fs
The ongoing reliability of the PCDD/Fs analytical methodology was assured by means of the use of the EPA 1613 LCS standard and the development of method blanks. The average recoveries were in the range of 68.2% to 103.4% for 13C-labeled 2,3,7,8-CDD/Fs congeners, within the ranges established in the EPA 1613 method. Besides, blanks covering the whole sample preparation methodology showed that congeners were either not detected or below the detection limits.
64 Materials and Methods
2.8. References
APHA, AWWA, WPCF, Standard Methods for Examination of Water and Wastewater, 20th ed., American Public Health Association, American Water Works Association, Water Pollution Federation, Washington DC, USA, 1998.
EPA Method 163. Tetra-trough octa-chlorinated dioxins and furans by isotope dilution HRGC-HRMS, 1994. U.S. Environmental Protection Agency. Office of Water. Engineering and Analysis Division (4303). 401 Street S.W. Washington, D.C.
UNE-EN-ISO 1484: 1998. Análisis del agua. Directrices para la determinación del carbono orgánico (COT) y del carbono orgánico disuelto (COD).
65
3. Results and discussion
Results and Discussion
3.1. Electrochemical oxidation of 2-chlorophenol
In this chapter, the assessment of the intermediate products, paying special attention to the potential formation of PCDD/Fs, during the electrochemical oxidation of aqueous solutions containing 2-CP using a boron doped diamond (BDD) anode is reported. The initial concentration of 2-CP in the solutions to be treated was set at 15.56 Mm (2000 mg L-1). Such relatively high concentration of 2-CP is not frequently found in environmental samples, since various industrial effluents and municipal waste discharges typically contain chlorophenols (CPs) concentrations from 1 to 21 mg L-1 (Terashima et al. 2002). Nevertheless, large concentrations of phenolic compounds have also been reported in wastewaters from oil refineries and coking plants (500 to 1,500 mg L-1)(El-Ashtoukhy et al. 2013) and even higher, up to 10,000 mg L-1, in olive oil mills wastewaters (Aissam et al. 2007). In this work, a relative high initial concentration of 2-CP has been used for the correct assessment of the formed byproducts during the treatment (Poerschmann et al. 2009; Munoz et al. 2011). Experiments have been carried out working with a current density (J) of 900 A m-2, according to the results reported in previous works developed by the research group, where the viability of the electrochemical oxidation using BDD electrodes to treat landfill leachates in terms of COD and ammonia removal was assessed (Cabeza et al. 2007; Anglada et al. 2009; Anglada et al. 2010; Anglada et al. 2011). Another important variable in the electrochemical oxidation is the selection of the proper electrolyte. The influence of two commonly used electrolytes, namely NaCl and Na2SO4, which are commonly found in wastewater matrices, has been studied (Comninellis & Nerini 1995; Kapałka et al. 2010; Govindaraj et al. 2013). Particularly, the oxidation mediated by active chlorine has attracted attention due to the wide presence of chloride in wastewaters, providing the necessary conductivity and avoiding the use of additional chemicals.
3.1.1. Major intermediate products in the electrochemical oxidation of 2-chlorophenol
The electrochemical oxidation of 2-CP model solutions with an initial concentration of 15.56 mM (2000 mg L-1) was performed under galvanostatic conditions (J = 900 A m-2) using a BDD anode and at room temperature, 20ºC.
Two broadly applied electrolytes, NaCl (56.3 mM) and Na2SO4 (21.1 mM),
69
Chapter 3
resulting in the same initial conductivity (7.5 mS cm−1), were used. Experimental data are depicted together with error bars obtained after replication of the experiments. The changes in normalized 2-CP concentration with time during the electrochemical treatment are depicted in Figure 3.1. Experimental data are depicted together with error bars obtained after replication of the experiments.
2-CP 1.0
0.8 o CP] - 0.6
CP] / [2 0.4 - [2 0.2
0.0 0 50 100 150 200 250 time (min) Figure 3.1. Changes in normalized 2-CP concentration with time.
Solid dots: NaCl; empty dots: Na2SO4.
As can be observed in Figure 3.1, the concentration of 2-CP decreased to undetectable values after 1 h using NaCl, whereas the use of Na2SO4 needed longer times (180 min) to degrade 2-CP completely. The kinetics of the electrochemical oxidation of 2-CP were satisfactorily described by pseudo first- order rate equations for both electrolytes, NaCl and Na2SO4. During the electrochemical oxidation of organic matter by means of BDD anodes in the potential region of oxygen evolution, OH are expected to be formed in high amounts (Panizza & Cerisola 2005). Therefore, assuming pseudo-stationary -1 2 state for OH , kinetic constant values of kNaCl = 0.06 min (R = 0.95) and kNa SO = -1 2 0.03 min (R = 0.97) were calculated, respectively for NaCl and Na2SO4. ₂ ₄ Two different operating regimes can be distinguished depending on the applied J with respect to the limiting current density (Jlim). The limiting current density establishes the limit between the control of the operating regime by the applied charge or by mass-transfer and was calculated according to eq 1:
Jlim = 4FKm COD (eq 1)
70 Results and Discussion
-1 where F is the Faraday’s constant (96500 C mol ), Km is the mass transport coefficient (m s-1), and the chemical oxygen demand, COD, is expressed as mol -3 O2 m . Depending on the applied J the following operating regimes are identified (Panizza et al. 2001):
• J < Jlim, the electrolysis is under current control. In this operating regime the current efficiency is 100% and COD decreases linearly with time.
• J > Jlim, the electrolysis is under mass-transport control. Secondary reactions such as oxygen evolution commence, resulting in a decrease of the current efficiency. Under these conditions COD removal follows and exponential trend.
The value of the mass transport coefficient is determined by means of the Sherwood (Sh), Reynolds (Re) and Schmidt (Sc) numbers, which are calculated through equations 2-4:
K d Sh = m h (eq 2) D
uρ d Re = h (eq 3) µ
µ Sc = (eq 4) ρD
-10 2 -1 In these equations D is the diffusivity of 2-CP (9.5 x 10 m s ), dh is the equivalent diameter (9.5 x 10-3 m), u is the fluid velocity (0.32 m s-1), µ is the viscosity of the fluid (1002 x 10-6 Kg m-1 s-1 (20ºC)), and ρ is the density of the fluid (998.2 Kg m-3 (20ºC)). Furthermore, the effect of the geometry of the cell is included according to equation 5.
Sh = a RebScc (eq 5)
The values of the parameters a, b, and c can be determined mathematically or experimentally. For circular electrodes, expression on the form of equation 6 has been proposed (Anglada 2010).
Sh = 1.67 Re0.385 Sc0.2 (eq 6)
71 Chapter 3
-5 -1 As a result a Km of 1.46·10 m s was obtained. Therefore, for a CODo of -3 104.9 mol m , corresponding to the initial concentration of 2-CP, the Jlim was -2 -2 591 A m . Since the applied J, 900 A m , was higher than the Jlim, the process should be mass-transfer-controlled. Under these conditions, the reaction between 2-CP and OH in the anode diffuse layer is favored because of the high concentration of OH on the anode surface compared with the pollutant concentration (Iniesta 2001). Therefore, due to the high concentration of OH, the generic reaction of 2-CP oxidation may be described by a pseudo first-order kinetic equation according to what has been observed with experimental data.
Additionally to the main oxidation mechanism mediated by OH, evidence of indirect oxidation by means of different electro-generated oxidant species, such as peroxodisulphates and active chlorine, has been reported in the literature (Cañizares et al. 2005; Mascia et al. 2010; Kapałka et al. 2010; Pérez et al. 2012). Moreover, such indirect oxidation mechanisms are favored at higher current densities (Chiang et al. 1995). According to Figure 3.1, considerable lower time was needed in order to decrease the concentration of 2-CP to undetectable levels with NaCl than with Na2SO4, suggesting a positive contribution of indirect oxidation by means of active chlorine to 2-CP degradation.
When chloride ions are present in the solution, they will be directly oxidized on the anode surface with the subsequent formation of chlorine radicals (Cl) that combine together to produce chlorine at the anode surface, which further hydrolyzes in water to form hypochlorous acid and hypochlorite (Anglada 2010; Bergmann 2010). The resulting mixture of active chlorine species (chlorine, hypochlorite and hypochlorous acid) is highly reactive with many organics, being efficient for their mineralization (Sirés et al. 2014). However, this reaction mixture can promote the formation of organochlorinated intermediates and final products that can be even more harmful than the raw pollutants (Comninellis & Nerini 1995). Besides, the consequent oxidation of this reaction mixture can promote the formation of chlorate, which is usually an unwanted product in treated effluents (Pérez et al. 2012; Azizi et al. 2011).
72 Results and Discussion
The mineralization efficacy, expressed in terms of total organic carbon (TOC) abatement, and COD removal during the electrochemical oxidation of 2- CP solutions versus Q is shown in Figure 3.2.
COD TOC 1 o 0.8
0.6
0.4
0.2 [TOC, COD] / [TOC, COD] [TOC, / [TOC, COD] 0 0 50 100 150 200 250 time (min)
Figure 3.2. Changes in normalized COD and TOC concentrations with time.
Solid dots: NaCl; empty dots: Na2SO4.
According to the results depicted in Figure 3.2, there was no influence of the electrolyte type on the kinetic change of either TOC or COD, since the difference in results may be within the experimental error. TOC mineralization required higher times than did the reduction of both COD and 2-CP, independently of the used electrolyte. This pointed out to the formation of oxidation intermediate products that need longer times to be mineralized. The hydroxyl functional group in 2-CP is a highly active ortho/para-directing group, and OH are strong electrophilic radicals (Huang & Chu 2012). In addition, although the chlorine atom of the aromatic ring is a deactivating group, it is also an ortho/para-directing group (Morrison & Boyd 1998). Therefore, hydroxylation and chlorination products with predominantly ortho/para substitutions are expected. Below, the major intermediate products detected during the electrochemical oxidation of 2-CP are shown. Figure 3.3 displays the changes in the concentration of the detected chlorinated aromatic intermediate compounds with time, in the runs performed with both electrolytes, NaCl and
Na2SO4.
73 Chapter 3
2,4-DCP 2,6-DCP 2,3,6-TCP 2,6-dichlorobenzoquinone 1.8
) 1.5 0.03
mM 0.02 ( 1.2 0.01
0.9 0 0 50 100 150 200 250 0.6 Concentration Concentration 0.3
0 0 50 100 150 200 250 time (min) a
0.02 2-chlorobenzoquinone ) mM
0.01 Concentration ( Concentration
0 0 50 100 150 200 250 time (min) b
Figure 3.3. Changes in the concentration of chlorinated aromatic intermediates with
time: a) NaCl; b) Na2SO4.
As can be observed in Figure 3.3a, when NaCl was used as electrolyte, three chlorosubstituted phenols, 2,4-dichlorophenol (2,4-DCP), 2,6-dichlorophenol (2,6-DCP) and 2,3,6-trichlorophenol (2,3,6-TCP), were formed during the first min of the electrooxidation treatment. They showed a typical accumulation- destruction cycle with maximum concentrations at 30 min, and negligible values after 2h of treatment. The highest concentrations of chlorinated intermediates were noticed for 2,4-DCP, highlighting the preference of 2-CP chlorination in the para position (Huang & Chu 2012). Finally, the presence of 2,6- dichlorobenzoquinone was noticed at very low concentrations (< 0.03 mM).
74 Results and Discussion
With respect to the experiments carried out with Na2SO4, whose results are reported in Figure 3.3b, chlorinated intermediates were not identified with the exception of 2-chlorobenzoquinone that was detected at a very low concentration (< 0.01 mM); this can be attributed to the low concentration of chlorine species in the solution, only coming from C-Cl bond cleavage of 2-CP. In addition to chlorinated derivatives, the formation of non-chlorinated aromatic intermediates was reported. The changes in their concentration with the treatment time are detailed in Figure 3.4.
Hydroquinone Catechol Benzoquinone Phenol 0.8 ) 0.6 mM (
0.4
0.2 Concentration Concentration
0 0 50 100 150 200 250 time (min) a
0.8 ) 0.6 mM
0.4
0.2 Concentration ( Concentration
0 0 50 100 150 200 250 time (min) b
Figure 3.4. Changes in the concentration of non-chlorinated aromatic intermediates with
time: a) NaCl; b) Na2SO4.
75 Chapter 3
According to data from Figure 3.4a, when NaCl was used as electrolyte, non-chlorinated aromatic products were identified at lower concentrations than chlorinated derivatives. Catechol (arising from the hydroxylation of the aromatic ring in the ortho position) was the main product reaching a peak concentration after 2 h, and then decreasing with treatment time. On the other hand, hydroxylation in the para position led to the formation of hydroquinone, which showed higher concentrations initially, but then decreased leading to its derivative benzoquinone. The latter reached a maximum concentration at 2h and then decreased until negligible values.
On the other hand, when Na2SO4 was used instead of NaCl, Figure 3.4b, hydroquinone was the main intermediate product reaching a maximum in its concentration at 30 min, and then decreasing to minimal concentrations after 4 h. The same trend, but with lower concentration was observed for catechol, meaning that hydroxylation preferentially took place in the para position. Moreover, the oxidation of hydroquinone to benzoquinone did not take place to a significant extent. Additionally, the presence of phenol, again in low concentrations, was detected with peak concentrations at initial times.
The cleavage of the aromatic ring led to the formation of aliphatic carboxylic acids, whose concentrations with the electrochemical treatment time using both electrolytes, NaCl and Na2SO4, are displayed in Figure 3.5. In general, longer times were needed for the destruction of organic acids compared to intermediate aromatic compounds, due to their much lower reactivity with OH (Dirany et al. 2012; Lopez et al. 2005).
76 Results and Discussion
Acetic Formic Malonic Oxalc Maleic Fumaric 3
) 2.5 mM ( 2
1.5
1
Concentration Concentration 0.5
0 0 50 100 150 200 250 time (min) a 3
) 2.5
mM 2
1.5
1
Concentration ( Concentration 0.5
0 0 50 100 150 200 250 time (min) b
Figure 3.5. Changes in aliphatic organic acids concentration with time:
a) NaCl; b) Na2SO4.
Results from Figure 3.5a showed that with NaCl, formic and acetic acids, which were the predominant organic acids, reached a maximum concentration at 1 h, and then began to decrease with the treatment time. Besides, the presence of maleic, acetic and malonic acids was also reported, but at lower concentrations. The use of Na2SO4 led to the formation of the same aliphatic carboxylic acids, with similar trends regarding oxidation time (Figure 3.5b), but with the exceptions that malonic acid could not be detected, and fumaric acid was detected at low concentrations. The observed dominance of formic and oxalic acids with both electrolytes suggested that they are formed from different routes to become the ultimate reaction resulting from the aromatic ring cleavage (Zazo et al. 2005).
77 Chapter 3
3.1.2. Mass balances of total organic carbon and chlorine
Theoretical TOC values were calculated accounting for the identified intermediate products together with the remaining concentration of 2-CP, and the obtained values were compared to experimentally measured TOC data, Figure 3.6.
MeasuredTOC Calculated TOC --Initial Cl Calculated Cl 120 80
100 60 80 ) ) mM mM 60 40 ( Cl
TOC ( 40 20 20
0 0 0 5 30 60 120 240 time (min) a
120 80
100 60 )
80 ) mM mM 60 40 ( Cl TOC ( 40 20 20
0 0 0 5 30 60 120 240 time (min) b
Figure 3.6. Changes in measured and calculated TOC and chlorine balance with time.
a) NaCl; b) Na2SO4.
As can be observed in Figure 3.6, the differences between measured and calculated TOC during the first 2 h, which were comparable for both electrolytes, could be explained by the presence of unidentified oxidation
78 Results and Discussion
intermediates. After 4 h of treatment, the TOC balance was almost 100% satisfied when Na2SO4 was employed as electrolyte, Figure 3.6b, whereas with NaCl a difference between measured TOC (5 mM) and theoretical TOC (1 mM) was determined (Figure 3.6a). This difference, although low in absolute value, could be due to the presence of minor byproducts formed by further oxidation and/or chlorination reactions of the aromatic ring. On the other hand, the calculated concentration of chlorine as the sum of its content in the analyzed species, 2-CP, chlorinated intermediates, chloride and chlorate, was also shown in Figure 3.6. As observed for TOC, the chlorine balance was far to be closed when NaCl was used as electrolyte. As mentioned above, the oxidation of chloride ions at the anode surface form a mixture of active chlorine species that can be further oxidized to chlorate and perchlorate (Pérez et al. 2012; Azizi et al. 2011; Sirés et al. 2014). Since the contribution of chloride, chlorate and chlorinated identified compounds to the chlorine balance has been quantified, and perchlorate ions were not detected in the oxidized samples, the unbalanced chlorine was attributed to the formation of active chlorine species along with non-identified chlorinated compounds.
3.1.3. Formation of PCDD/Fs during the electrochemical oxidation of 2-chlorophenol
Since 2-CP and its di- and tri-chlorinated derivatives are known to be potential precursors of PCDD/Fs formation, these compounds were assessed as possible minor byproducts as a result of the electrochemical oxidation of 2-CP. Once the analytical methodology for the determination of PCDD/Fs was satisfactorily established (chapter 2.6), the analysis of PCDD/Fs in the electrochemical oxidation of 2-CP experiments was performed. The homologue profiles of total PCDD/Fs after 4 h of treatment using either NaCl or Na2SO4 electrolytes, together with the profile found in the untreated 2-CP solution, are shown in Figure 3.7.
79 Chapter 3
Untreated 2-CP 4h (NaCl) 4h (Na₂SO₄) 1E+06
) 1E+05 1 -
1E+04
1E+03
1E+02 Concentration (Pg L (Pg Concentration 1E+01
1E+00
Figure 3.7. Comparison between the homologue profiles of total PCDD/Fs in the untreated 2-CP solution and in the electrochemical oxidized solutions (t= 4 h).
The possible contamination of the experimental system by the accumulation of PCDD/Fs was determined before each experiment. For this purpose, a blank solution was pumped through the electrochemical cell under the same experimental conditions used for 2-CP oxidation, and the concentration of PCDD/Fs was then determined. The concentrations of PCDD/Fs reported in Figure 3.7 were blank corrected by means of the subtraction of the PCDD/Fs background levels measured in the blank before each experiment. The total removal of trace concentrations of PCDD/Fs from the electrochemical cell was not possible, since the cleaning protocol applied does not allow the use of organic solvents to avoid damaging some parts of the equipment.
According to data reported in Figure 3.7, three groups of homologues, OCDD, TCDF and HpCDF, were detected at very low concentrations, from 1.6 to 7 pg L-1, in the untreated solutions containing 2-CP. When NaCl was used as supporting electrolyte, a considerable formation of PCDD/Fs was observed after 4 h of treatment. The total PCDD/Fs concentration in the oxidized sample (391.05 ng L-1) was 2.68 x 104 times higher than in the untreated solution. PCDFs contributed 88% to the total PCDD/Fs concentration. Among them, the group of HxCDFs was the dominant one, accounting for 68% of the total PCDD/Fs
80 Results and Discussion
concentration, and followed by PeCDF and TCDF which represented 12.5% and 5.7% of the total PCDD/Fs, respectively. On the other hand, PCDDs explained the remaining 12%, being the two major homologue groups PeCDD (5.4%) and TCDD (3.5%). The dominance of PCDFs over PCDDs indicates that the formation of PCDFs intermediates is favored against the corresponding ones for PCDDs.
With respect to the use of Na2SO4 as electrolyte, as it is depicted in Figure 3.7, the concentration of total PCDD/Fs increased by 200 times related to the untreated sample, but was 134 times lower than in the presence of NaCl. TCDD, HxCDD and TCDF were the only homologue groups that showed an increase in their concentration after 4 h of treatment. Their contribution to the total concentration was over 51.6%, 2.5% and 39.8%, respectively. For the remaining homologue groups, their concentrations were close to the background levels.
Regarding the most toxic congeners, 2,3,7,8-PCDD/Fs, their concentrations in both the untreated 2-CP model solutions and the oxidized samples, after 4 h and using either NaCl or Na2SO4 as electrolytes, are shown in Figure 3.8.
Untreated 2-CP 4h (NaCl) 4h (Na₂SO₄) 10000 ) 1 - 1000
100
10 Concentration (Pg L (Pg Concentration 1
0.1
Figure 3.8. Congener profiles of 2,3,7,8-PCDD/ in the untreated 2-CP solution and in the electrochemical oxidized solutions (t= 4 h).
81 Chapter 3
As reported in Figure 3.8, only two 2,3,7,8-chlorinated congeners, OCDD and 1,2,3,4,6,7,8-HpCDF, were detected at very low concentration (1.6-6 pg L-1) in the untreated solution of 2-CP. Regarding the electrochemically oxidized samples, when Na2SO4 was used, three 2,3,7,8-PCDD/Fs congeners were observed: 1,2,3,7,8-PeCDF, 2,3,4,7,8-PeCDF and 2,3,4,6,7,8-HxCDF, but at lower concentrations (0.7-1.34 pg L-1), similar to those of the untreated 2-CP solution. The remaining congeners showed concentration values near background levels.
On the other hand, according to Figure 3.8, when NaCl was present in the reaction medium the formation of 2,3,7,8-PCDD/Fs was notable. The total concentration of 2,3,7,8-PCDD/Fs (6311.10 pg L-1), which represented the 1.6% of the total PCDD/Fs concentration, increased by 828 times compared with the untreated 2-CP solution. The 2,3,7,8-congener profile was dominated by 1,2,3,4,6,7,8-HpCDF (32.6% of the total 2,3,7,8-PCDD/Fs concentration). Moreover, OCDD (17.7%), 1,2,3,4,6,7,8-HpCDD (17%) and OCDF (12%) contributed in a high percentage to the total concentration of 2,3,7,8-PCDD/Fs. Contrary to what was shown with homologue groups, where the ratio PCDFs/PCDDs was 7.4, the ratio between 2,3,7,8-PCDDs and 2,3,7,8-PCDFs is close to 1, indicating that both families of congeners contribute in equal quantity to the total 2,3,7,8-PCDD/Fs concentration.
The toxic equivalents (TEQ), which quantify the exposure risk of 2,3,7,8- PCDD/Fs-containing samples, were calculated for samples in which NaCl was used as electrolyte, since they were those reporting a relevant 2,3,7,8-PCDD/Fs concentration. Both the international (I-TEFs) and the World Health Organization toxicity equivalency factors (WHO-TEFs) were utilized. When the I- TEFs were applied, sample TEQ increased from values close to zero for the untreated solution to 220 pg I-TEQ L-1 when almost complete mineralization (t = 4 h) was achieved. On the other hand, when the WHO-TEFs values were used, the total TEQ increased even higher, up to 243 pg WHO-TEQ L-1, mainly due to the congener 1,2,3,7,8-PeCDD that posses a WHO-TEF value twice as high as its corresponding I-TEF. These values are far higher than the maximum contaminant level (MCL), 30 pg L-1 of 2,3,7,8-TCDD, established by the U.S. EPA based on potential health effects from water ingestion.
Up to our knowledge, this is the first time that an increase in the concentration of PCDD/Fs as a result of the electrochemical oxidation of chlorinated organic compounds has been quantified. In a previous work, the formation of dioxins and furans during the electrochemical oxidation of 4-CP
82 Results and Discussion
solutions was qualitatively reported by Hong et al. (2003). They studied the decomposition of 4-CP with an iridium dioxide anode and working at 400 A m-2 using a mixture of H3PO4 and NaH2PO4 as electrolyte. The authors identified the presence of hydroxylated dibenzodioxins/furans and hydroxylated chlorobiphenyls as minor byproducts after 60 min of treatment, but without their quantification.
The comparison between the results obtained using two broadly applied electrolytes, NaCl and Na2SO4, emphasizes the role and relevance of the chloride concentration in the formation of highly substituted chlorinated byproducts during the electrochemical oxidation of 2-CP. NaCl, which exerted a positive kinetic influence in the removal of 2-CP, showed a strong propensity to form polychlorinated derivatives. Thus, we emphasize the importance of the electrolyte selection in electrooxidation processes beyond their influence on the degradation of the primary pollutant or major environmental parameters such as TOC or COD, especially when PCDD/Fs precursors are initially present or may be formed in the samples to be treated. Furthermore, it must be highlighted that high concentrations of NaCl are ubiquitous in a wide variety of wastewaters, such as those generated during the manufacture of pesticides, pharmaceuticals, dyes, and from landfill leachates (Primo et al. 2008; Anglada et al. 2009; Bacardit et al. 2007). Therefore, the concomitant presence of chloride and organic compounds in wastewater is an important issue to assess since the formation of harmful chlorinated organic byproducts could take place as a result of their electrochemical treatment.
3.1.4. Proposed reactions pathway for the electrochemical oxidation of 2-chlorophenol
Summarizing the information collected from the detected (chlorinated)derivatives during the electrochemical oxidation of 2-CP and including data from the literature, a scheme containing various reaction pathways is proposed in Figure 3.9, where three different routes, A, B and C, are distinguished.
83 Results and Discussion
OH Cl
2-CP + OH OH (A) Chlorinated intermediates (B) Non-chlorinated intermediates (C) PCDD/Fs + OH Cl + Cl OH - Cl - Cl OH· - H Cl Cl OH + Cl-phenoxy radical -H2O OH OH OH OH OH O OH OH Cl Cl Cl OH Cl Cl Cl - H DHCD radical* Cl-phenoxy radical*
Cl Coupling reactions 2-CP Cl OH O catechol benzoquinone phenol Cl 2,6-DCP 2,4-DCP 2,3,6-TCP hydroquinone O O + OH + Cl OH O OH O OH Cl Cl H Cl Cl Cl Cl + OH Cl Cl Cl-phenoxyphenol* -HCl +H· Cl-diphenyl ether* - H - H +Cl -Cl Cl O O O OH OH O 2,6-Dichlorohydroquinone* 2,6-dichlorobenzoquinone 2-chlorohydroquinone* 2-chlorobenzoquinone O · Cl Cl H PCDDs +Cl O O O O O O O O HO HO O OH H3C H Ring OH HO OH HO OH O O OH OH Cleavage PCDFs Cl Fumaric Maleic Malonic Oxalic Acetic Formic Cl
Figure 3.9. Formation mechanisms of intermediate products in the electrochemical oxidation of 2-CP (*: non detected in this work)
84
Results and Discussion
Route A from Figure 3.9 led to the formation of chlorinated aromatic products by nucleophilic addition of chlorine radicals to the aromatic ring. Chloride ions present in the reaction medium and/or generated from the degradation of 2-CP may have been oxidized at the anode surface into Cl (Cl- → Cl + e-) (Boudenne & Cerclier 1999; Bergmann 2010). The substitutive nucleophilic aromatic reaction involving Cl as electrophilic radical resulted in the formation of 2,4-DCP and 2,6-DCP, as a result of the replacement of one hydrogen of the aromatic ring by Cl in para and ortho positions, respectively. When two Cl take part in the substitution, 2,3,6-TCP is formed. Afterwards, 2- chlorobenzoquinone and 2,6-dichlorobenzoquinone were formed by hydroxylation in para position of 2-CP and 2,6-DCP, respectively, followed by consecutive dehydrogenation.
Otherwise, route B led to the formation of non-chlorinated aromatic intermediates. Phenol, which was only detected using Na2SO4, was formed from the cleavage of the C-Cl bond of 2-CP. The remaining intermediates may have been formed from phenol hydroxylation or since phenol was not detected in the reaction medium using NaCl, they could also be formed through OH·attack, catechol being formed by hydroxylation in ortho position and hydroquinone in para position, along with loss of chlorine (Wang & Wang 2009). Indeed, a substitutive nucleophilic reaction may have occurred, where OH came and replaced an atom of chlorine (Boudenne & Cerclier 1999). Hydroquinone was subsequently dehydrogenated to benzoquinone (Wang & Wang 2009).
On the other hand, the route C, which deals with the formation of PCDD/Fs, has been proposed according to results previously reported in the literature. CPs are known to be the most direct precursors in the formation of PCDD/Fs congeners (Ryu et al. 2005; Pan et al. 2013), which may be formed even at room temperature (Weber 2007). The formation of PCDD/Fs from single aromatic molecules (phenols and benzenes) requires biaromatic intermediates (Weber & Hagenmaier 1999). It has been demonstrated that the electrophilic addition of OH to CPs results in the formation of chlorosubstituted dihydroxycyclohexadienyl (DHCD) radicals (Duesterberg & Waite 2007). Afterwards, the elimination of unimolecular water from the Cl-substituted DHCD radical leads to Cl-phenoxy radical (Duesterberg & Waite 2007), in which the unpaired electron resonates among different positions. The oxidative coupling reactions of resonance-stabilized radicals with other radicals or
85
Chapter 3
molecules give rise to biaromatic intermediates such as biphenyls, diphenyl ethers and phenoxyphenols (Altarawneh et al. 2009; Poerschmann et al. 2009).
(Chlorinated)hydroxybipehnyls and also (chlorinated)diphenyl ethers are known to be the most important intermediates in the formation of PCDFs from CPs (Weber & Hagenmaier 1999; Liu et al. 2004), whereas (chlorinated)phenoxyphenols are key intermediates in the formation of PCDDs (Holt et al. 2012). According to Altarawneh et al. (2009), PCDFs formation is considered to take place exclusively from the condensation of two radicals, whereas PCDDs involve additionally molecule/radical and molecule/molecule coupling reactions. The chlorination of the condensation products may also contribute to the higher chlorinated congeners that were not expected from direct condensation reactions between the reactants (Briois et al. 2007). Finally, the aromatic ring cleavage led to the formation of aliphatic carboxylic acids.
86 Results and Discussion
3.2. Fenton oxidation of 2-chlorophenol
This chapter assesses the distribution of intermediate products, highlighting the potential formation of PCDD/Fs, as a result of the Fenton treatment of solutions containing 2-CP. As in the case of the electrochemical experiments depicted in chapter 3.1, the initial concentration of 2-CP was set at a relatively high value, 15.56 mM (2000 mg L-1), in order to guarantee a correct assessment of the formed byproducts during the oxidation process (Munoz et al. 2011; Poerschmann et al. 2009). As it was mentioned in the introduction chapter, the efficiency of the Fenton process depends on different operating conditions such as working pH, temperature, and H2O2 and catalyst concentrations. In this work, the influence of two important operating variables, namely H2O2 dose and temperature was assessed.
2+ Regarding the concentration of H2O2 and Fe , H2O2 dose is important in order to obtain better degradation efficiency, while iron concentration is more important for the reaction kinetics (Chamarro 2001). Since the concentration of
H2O2 is the variable more directly related with the extent in the mineralization of the organic compounds, its influence in the Fenton oxidation of 2-CP has been studied in this work. H2O2 dose is frequently established according to the initial pollutant concentration. On this way, high molar ratios of H2O2/pollutant have been widely used in order to achieve high mineralization degrees. However, one important limitation in the treatment of heavily polluted wastewaters by the Fenton process could be the high H2O2 consumption, which accounts for 90% of the process costs (Munter et al. 2006). Therefore, to maintain economic feasibility, lower oxidant concentrations have been proposed; nevertheless, target contaminants may be degraded under these conditions but stable, and potentially toxic, aromatic intermediates may also be formed as a result of the treatment process (Poerschmann et al. 2009). On this way, two H2O2 doses corresponding to 100% (202.28 mM) and 20% (40.44 mM) of the theoretical stoichiometric amount to oxidize 2-CP have been assessed.
With respect to Fe2+ concentration, it was kept at catalytic amounts 2+ (Fe /H2O2 << 1) in order to make the most of H2O2 consumption avoiding OH 2+ recombination and OH scavenging by Fe (r10) (Duesterberg & Waite 2006; 2+ Lopez et al. 2005), and to prevent any coagulation taking place when Fe /H2O2 ratio is greater than 2 (Neyens & Baeyens 2003). Besides, low iron quantities ensure a slower H2O2 consumption making the identification of intermediate
87 Chapter 3
products easier (Munoz et al. 2011). Although there is not an established
H2O2/Fe ratio, typical mass ratios applied in the literature are in the ranges 5-25 (Pera-Titus et al. 2004; Xu et al. 2003) and 61-607 (Pignatello et al. 2006). In this work, and based on previous studies found in the literature (Munoz et al. 2011; 2+ Munoz et al. 2013), a H2O2/Fe molar ratio of 28 was used for H2O2 at 100% of the stoichiometric amount, whereas a higher molar ratio, 225, was applied with
H2O2 at 20% due to the low dose of this reagent. On the other hand, with respect to the working temperature, different opinions have been found in the literature. Whereas some authors have claimed the thermal instability of H2O2 at high temperatures (Gogate & Pandit 2004), Zazo and coworkers found that the increase of the temperature within the range of 25 to 130ºC enhanced mineralization rates (Zazo et al. 2011). In this work, the influence of the operating temperature was assessed working at two different values, 20ºC and 70ºC. Finally, since chloride may be one of the final mineralization products of 2-CP and it has been detected in a wide variety of aqueous samples (Kapałka et al. 2010; Bacardit et al. 2007), the influence of the presence of chloride ions (56.3 mM) in the reaction medium on the oxidation yield was evaluated.
3.2.1. Effect of H2O2 dose on Fenton degradation of 2- chlorophenol
The effect of two different concentrations of H2O2 corresponding to 20% (40.44 mM) and 100% (202.28 mM) of the theoretical stoichiometric amount to 2+ oxidize 2-CP was investigated. H2O2/Fe molar ratios of 225 and 28 were used for H2O2 at 20% and 100% of the stoichiometric amount, respectively. Experimental data are depicted together with error bars obtained after replication of the experiments. The changes in normalized 2-CP, H2O2, COD and TOC concentrations with time during the Fenton oxidation of 2-CP using both
H2O2 doses, 20% and 100% of the stoichiometric amount, are depicted in Figure 3.10.
88 Results and Discussion
2-CP H2O2 1 1 o ]
2 0.8
O 0.8 2 0.6 0.4 CP, H 0.6 - 0.2 ]/[2 2 0.4 0 O
2 0 10 20 30
CP, H 0.2 - [2 0 0 50 100 150 200 250 time (min) a TOC COD 1 o 0.8
0.6
0.4
0.2 [TOC, COD]/[TOC, COD]/[TOC, [TOC, COD] 0 0 50 100 150 200 250 time (min) b
Figure 3.10. Effect of H2O2 dose on: a) 2-CP; b) COD and TOC. Empty dots: 20% stoichiometric, solid dots: 100% stoichiometric.
As can be observed in Figure 3.10a, the oxidation of 2-CP was effective for both concentrations (20 and 100% of the stoichiometric H2O2 dose), and resulted in more than 85% of 2-CP degradation after 5 min. The experiments with 100% H2O2 resulted in faster degradation of 2-CP due to the higher OH availability. Under these conditions, the concentration of 2-CP was reduced by more than 99% within the first min of reaction.
89 Chapter 3
COD and TOC profiles displayed in Figure 3.10b suggest that their degradation proceeds via a first step, within the first 10 min with a faster rate. Afterwards, the organic matter started to decay more slowly. COD degradation and TOC mineralization were far from being completed at both H2O2 doses, even when 2-CP was entirely depleted. Using 100% of the stoichiometric amount of H2O2, higher COD and TOC removals were achieved (72% and 23% respectively, after 4 h), whereas with H2O2 at 20% of the stoichiometric value, COD and TOC were only reduced by 15% and 4% respectively, remaining constant after 30 min because of the total H2O2 consumption. These observations are consistent with the Fenton reaction scheme reported in chapter 1, where in a first step, the OH formed through r5 react very fast with the organic matter (107-1010 M-1 s-1) (Benitez et al. 2000). With 2+ the course of the reaction, the consumption of H2O2 along with the slower Fe 3+ regeneration from Fe (r6 and r8) that limited the production of OH from H2O2 (Bautista et al. 2008), lead to a slower degradation of the organic pollutants. Besides, the reaction progress involves intermediate oxidized byproducts, likely oxygenated compounds, perhaps containing carboxylic groups, with relatively slow degradation by OH· (Lopez et al. 2005). In addition, the presence of certain organic products (carboxylic acids such as oxalic acid) could result in the formation of complexes with Fe3+ deactivating its capacity to regenerate Fe2+ (Nakagawa & Yamaguchi 2012). These observations, along with a decrease in pH, suggest the formation of byproducts with low degradation rates, such as aliphatic organic acids. The evolution of the detected oxidized byproducts during the Fenton treatment of 2-CP with both H2O2 dose is shown in Figure 3.11.
90 Results and Discussion
2-Chlorobenzoquinone Catechol 6 1
) 5 0.8 mM ( 4 0.6 3 0.4 2
Concentration 1 0.2
0 0 0 50 100 150 200 250 time (min) a
Acetic Formic Oxalic X Fumaric 13 1 ) 11 0.8 mM ( 9 0.6 2 2 0.4 1
Concentration 0.2
0 0 0 50 100 150 200 250 time (min) b
Figure 3.11. Effect of H2O2 dose on: a) aromatic intermediates; b) organic acids. Empty dots: 20% stoichiometric, solid dots: 100% stoichiometric.
In Figure 3.11a, two aromatic intermediate products, 2- chlorobenzoquinone and catechol were identified only in samples containing
20% H2O2. Both compounds reached maximum concentrations at 5 min and remained at constant values after 30 min, coinciding with total H2O2 consumption. These aromatic intermediates were not detected with 100% H2O2, probably because of the higher OH availability that led to a faster aromatic ring cleavage producing aliphatic carboxylic acids. On the other hand, in Figure 3.11b, when 100% H2O2 was used, oxalic and formic acids reached peak concentrations within the first 30-60 min, corresponding with a first stage of fast H2O2 consumption A similar trend was observed for acetic acid, but at lower concentrations, whereas fumaric acid was only detected in trace concentrations. The use of 20% H2O2 led to the formation of the same organic
91 Chapter 3
acids (with the exception of fumaric acid), but with lower maximum concentrations since the oxidation reaction stopped at the formation of the aromatic intermediates because of the total consumption of H2O2.
3.2.2. Effect of temperature on Fenton degradation of 2- chlorophenol
The analysis of 2-CP degradation was performed at two different temperatures, 20ºC and 70ºC, for both H2O2 dose, 20% and 100% of the 2+ stoichiometric dose. H2O2/Fe molar ratios were kept at 225 and 28 for H2O2 at
20% and 100%, respectively. The changes in normalized 2-CP and H2O2 concentrations with time using both temperatures, 20ºC and 70ºC, and H2O2 at 20% and 100% of the stoichiometric amount, are depicted in Figure 3.12.
2-CP H2O2 1 o ] 1 2 O
2 0.8 0.8 0.6 CP, H - 0.6 0.4
]/[2 0.2 2 0.4 O
2 0 0 10 20 30 0.2 CP, H - [2 0 0 50 100 150 200 250 time (min) a 1 o
] 1 2
O 0.8 2 0.8 0.6 CP, H
- 0.6 0.4
]/[2 0.2 2
O 0.4 0 2 0 10 20 30
CP, H 0.2 - [2 0 0 50 100 150 200 250 time (min) b
Figure 3.12. Effect of Temperature on 2-CP and H2O2: a) 20% H2O2; b) 100% H2O2. Empty dots: 70ºC, solid dots: 20ºC.
92 Results and Discussion
According to Figure 3.12, faster degradations of 2-CP were observed at
70ºC than at 20ºC for both H2O2 concentrations, as would be expected due to the increase in the kinetic constant of the oxidation reactions with temperature. Near complete degradation of 2-CP was observed within the first min of reaction for both H2O2 dose. On the other hand, H2O2 was almost consumed within 5 and 15 min for its concentrations at 20% and 100%, respectively. Some authors claim that the thermal instability of H2O2 is a constriction for the use of Fenton oxidation at high temperatures (Gogate & Pandit 2004). Nevertheless, Zazo et al. (2011) found that increasing the temperature in the range of 25 to 130ºC resulted in faster Fe-catalyzed conversions of H2O2 into OH enhancing mineralization rates. On this way, COD and TOC were analyzed and their normalized profiles at 20ºC and 70ºC, with H2O2 at 20% and 100% are shown in Figure 3.13.
TOC COD 1 o
0.8
0.6
0.4
0.2 [TOC, COD]/[TOC, COD]/[TOC, [TOC, COD] 0 0 50 100 150 200 250 time (min) a 1 o 0.8
0.6
0.4
0.2 [TOC, COD]/[TOC, COD]/[TOC, [TOC, COD] 0 0 50 100 150 200 250 time (min) b
Figure 3.13. Effect of Temperature on COD and TOC: a) 20% H2O2; b) 100% H2O2. Empty dots: 70ºC, solid dots: 20ºC.
93 Chapter 3
Results from Figure 3.13 indicate that the overall mineralization was considerably more effective at 70ºC. The reduction in TOC increased from 3% at
20ºC to 27% at 70ºC (H2O2 at 20%, 4h) and from 23% at 20ºC to 69% at 70ºC
(H2O2 at 100%, 4 h).The positive effect of temperature was also depicted for COD measurements. The reduction in COD increased from 15% at 20ºC to 54% at 70ºC (H2O2 at 20%, 4h) and from 72% at 20ºC to 90% at 70ºC (H2O2 at 100%,
4 h). With H2O2 at 20%, the increase of working temperature, favored COD reduction at higher extent than TOC mineralization. On the other hand, at 100%
H2O2 the grade of mineralization depicted the highest enhancement when the temperature was increased.
COD and TOC profiles depicted again the presence of non-degraded organic matter remaining in the reaction medium. As it was previously reported, aromatic intermediates were not detected when the stoichiometric dose of
H2O2 was used. For H2O2 at 20%, the evolution of the detected aromatic intermediates with the treatment time at both temperatures, 20ºC and 70ºC, is shown in Figure 3.14.
Catechol 2-Chlorobenzoquinone
) 6 0.6 mM ) 4 0.4 mM (
2 0.2 Catechol chlorobenzoquinone ( - 2 0 0 0 50 100 150 200 250
time (min)
Figure 3.14. Effect of Temperature on aromatic intermediates with 20% H2O2. Empty dots: 70ºC, solid dots: 20ºC.
As can be observed in Figure 3.14, the concentration of the main aromatic intermediates detected, 2-chlorobenzoquinone and catechol, was lower at 70ºC than at 20ºC. These results were in agreement with the enhancement in COD oxidation and TOC mineralization reported in Figure 3.13 as consequence of the increase in the operating temperature. On the other hand, the changes in the
94 Results and Discussion
concentrations of the aliphatic organic acids with the treatment time at both temperatures, 20ºC and 70ºC, and working with H2O2 at 20% and 100% are displayed in Figure 3.15.
Acetic Formic Maleic Oxalic X Fumaric 0.6
0.4
0.2 Concentration (mM) Concentration
0 0 50 100 150 200 250 time (min) a
16
) 12 mM ( 8 34 time (min) 2
Concentration 1
0 0 50 100 150 200 250 time (min) b
Figure 3.15. Effect of Temperature on organic acids. a) 20% H2O2; b) 100% H2O2. Empty dots: 70ºC, solid dots: 20ºC.
As it can be seen in Figure 3.15a, with 20% H2O2 the temperature increase resulted in high concentrations for some organic acids, such as oxalic and formic acids. When 100% H2O2 was used, Figure 3.15b, formic and acetic acids were observed at similar concentrations with both temperatures, whereas higher amounts of oxalic acid were observed at 70ºC. Under these conditions, acetic, formic and oxalic acids were the only compounds present in the reaction medium accounting for 100% of the measured TOC. Although the increase in the working temperature favored TOC degradation, a complete mineralization
95 Chapter 3
could not be achieved. Carboxylic acids such as acetic, maleic, malonic, and oxalic acids are known to be poorly degradable by the Fenton reagent (Nakagawa & Yamaguchi 2012). In particular, oxalic acid can deactivate iron ion activity by the strong coordination with Fe3+ (Pignatello et al. 2006) and as a result, high relative concentrations of this compound was reported during the Fenton treatment of phenols (Zazo et al. 2005).
3.2.3. Effect of chloride ions on Fenton degradation of 2-chlorophenol
High concentrations of inorganic salts, especially NaCl, have been measured in wastewaters, such as those generated during the manufacture of pesticides, pharmaceuticals, dyes, and from landfill leachates (Primo et al. 2008; Anglada et al. 2009; Bacardit et al. 2007). Besides, since chloride may be one of the final mineralization products of 2-CP, its effect was analyzed using two sets of experiments differently favored in terms of TOC removal (100% H2O2 at 70ºC; -1 - 20% H2O2 at 20ºC), and in the presence of 56.3 mM (2000 mg L ) of Cl . Firstly, the effect of chloride on 2-CP removal, as well as on COD and TOC degradation, with 100% H2O2 at 70ºC was displayed in Figure 3.16. Experimental data are depicted together with error bars obtained after replication of the experiments. According to the results from this Figure, with 100% H2O2 and 70ºC, the presence of 56.3mM of Cl- in the reaction medium did not affect the oxidation of 2-CP, but increased TOC mineralization by 11% and COD degradation, although at lower extent (5%), when compared to samples without Cl-. As was shown by Micó et al. (2013), this fact can be attributed to - the formation of chlorine radical (Cl ) from the reaction between OH and Cl (r17 to r19), which contribute to the degradation of organic compounds.
• − •− OH + Cl ↔ ClOH (r17)
•− + • ClOH + H ↔ Cl +H2O (r18)
• − •− Cl + Cl ↔ Cl2 (r19)
96 Results and Discussion
2-CP H2O2 1 o ] 2
O 0.8 2 CP, H
- 0.6 ]/[2 2 0.4 O 2
CP, H 0.2 - [2 0 0 5 10 15 time (min) a TOC COD 1 o 0.8
0.6
0.4
0.2 [TOC, COD]/[TOC, COD]/[TOC, [TOC, COD] 0 0 50 100 150 200 250 time (min) b
Figure 3.16. Effect of chloride with H2O2 at 100% and at 70ºC in: a) 2-CP and H2O2; b) COD and TOC. Empty dots: with 56.3 mM NaCl, solid dots: without NaCl.
Although Cl formation requires OH scavenging, and Cl are less reactive with organic species than OH, the magnitude of the rate constants for Cl reactions with organic compounds are comparable to those for OH (Micó et al. 2013; Deng et al. 2012). In addition, Pignatello (1992) found that the cited scavenging effect of OH by Cl- was noticeable above 10 mM Cl- and accordingly, chloride concentrations higher than 10 mM have been used in this work. These results were contrary to those found in other studies, where Cl- significantly inhibited the removal of the organic matter as a result of OH scavenging and ferric ion complexation (Micó et al. 2013). However, the formation of ferric chlorocomplexes, which affects the regeneration of Fe2+, can be overcome if Cl-
97 Chapter 3
concentrations are less than 200 mM (Lu et al. 2005), as was also observed in this study. The increase in the degradation of the organic compounds by the presence of Cl- in the reaction medium led to a higher conversion of the organic matter into aliphatic organic acids (Figure 3.17).
Acetic Formic Oxalic 18
15
12
9
6
Concentration (mM) Concentration 3
0 0 50 100 150 200 250
time (min)
Figure 3.17. Effect of chloride ions with H2O2 at 100% and at 70ºC in the formation of organic acids. Empty dots: with NaCl, solid dots: without NaCl.
As can be seen in Figure 3.17, the presence of Cl- favored the conversion of the organic intermediates into oxalic acid, whereas it had no effect on acetic and formic acids. Regarding the experiments carried out with 20% H2O2 and at 20ºC, Figure 3.18, the presence of chloride did not affect the degradation of 2- CP or the overall TOC removal, whereas COD displayed a slight degradation increase by 13%. These results suggest that when low concentrations of the Fenton reagent, H2O2, were used, the lower production of OH led to preferential degradation of the main organic contaminant, 2-CP.
98 Results and Discussion
2-CP H2O2 1 o ]
2 1 O 2 0.8 0.8 0.6 CP, H - 0.6 0.4
]/[2 0.2 2 0.4 O
2 0 0 10 20 30 40 50 60
CP, H 0.2 - [2 0 0 50 100 150 200 250 time (min) a TOC COD 1 o 0.8
0.6
0.4
0.2 [TOC, COD]/[TOC, COD]/[TOC, [TOC, COD]
0 0 50 100 150 200 250 time (min) b
Figure 3.18. Effect of chloride with H2O2 at 20% and at 70ºC in: a) 2-CP and H2O2; b) COD and TOC. Empty dots: with NaCl, solid dots: without NaCl.
3.2.4. Mass balances of total organic carbon and chlorine
Theoretical TOC values after 4 h of treatment were calculated from the detected intermediate products and compared with the experimental TOC measured in the experiments previously described. This comparison together with the unidentified chlorine percentage are displayed in Figure 3.19.
99 Chapter 3
Measured TOC Calculated TOC Unidentified Cl (%) 100
80
60
40
20 TOC (mM); Unidentified Cl (%) Unidentified(mM); Cl TOC
0 20% estq 100% estq 20% estq 100% estq 20% estq 100% estq (20ºC) (20ºC) (70ºC) (70ºC) (20ºC, NaCl) (70ºC, NaCl)
Figure 3.19. Differences between measured and theoretical TOC (mM) and unidentified chlorine content (%).
Data from Figure 3.19 depicted that the difference between measured and theoretical TOC values decreased with increasing H2O2 dose and temperature.
After 4 h the TOC balance was only 100% satisfied with 100% H2O2 at 70ºC. Under the remaining operating conditions, a difference between measured and theoretical TOC was observed and attributed to the presence of unidentified oxidation intermediates. In addition, chlorine measurements were only balanced when the TOC was balanced. It is, therefore, reasonable that the aforementioned condensation byproducts include chlorinated organic compounds.
There are significant gaps in the understanding of the formation and identification of aromatic intermediates when using pronounced substoichiometric H2O2 to CPo molar ratios for Fenton oxidation (Poerschmann et al. 2009). In order to characterize some of these unknown byproducts, samples were extracted with dichloromethane and qualitatively analyzed using GC-MS. The main species identified after 4 h of treatment working with 20%
H2O2 were summarized in Table 3.1.
100 Results and Discussion
Table 3.1. Main compounds identified by GC-MS after 4 h with 20% H2O2.
According to Table 3.1, the main species identified at low retention times were chlorobenzenediols, mainly as 2-chlorohydroquinone, 4-chlorocatechol and 4-chlororesorcinol (which were not detected by HPLC, suggesting low yields). Compounds with higher molecular weights, such as condensation products formed by two-chlorinated aromatic rings, were detected at higher retention times. They were primarily 2,4'-dichloro-5-hydroxydiphenyl ether (m/z: 254, 184) and 4,4'-dichloro[1,1'-biphenyl]-3,3'-diol (m/z: 254, 155). The formation of coupling-reaction aromatic byproducts is also supported by the development of a brownish reagent medium (Sedlak & Andren 1991) and have also been reported during the treatment of CPs under nonstoichiometric Fenton (Poerschmann et al. 2009) and Fenton-like systems (Detomaso et al. 2003; Munoz et al. 2011; Munoz et al. 2013). Nevertheless, these compounds are expected to be present at low concentrations and therefore they could not balance the differences observed between measured and theoretical TOC. Therefore, it is likely that unknown analytes that did not separate and/or were not detected by the analytical methods applied remained in the solution.
101 Chapter 3
Anyway, these unknown analytes are susceptible to be oxidized, as the organic carbon balance was closed by increasing H2O2 dose and temperature. On this way, Sedlak & Andren (1991) identified the formation of colored aromatic polymers that were further oxidized by subsequent OH attack during the Fenton oxidation of chlorobenzenes.
3.2.5. Formation of PCDD/Fs during the Fenton degradation of 2-chlorophenol
The assessment of the potential formation of PCDD/Fs as minor byproducts during the Fenton oxidation of 2-CP was evaluated since this compound is a known precursor for PCDD/Fs formation, and its Fenton oxidized byproducts (i.e., chlorinated biphenyls and hydroxydiphenyl ethers) are key intermediates in the formation of PCDD/Fs from CPs (Weber & Hagenmaier 1999; Liu et al. 2004). The analyses of PCDD/Fs were performed after 4 h of Fenton treatment under the operating conditions of the aforementioned described experiments, where the influence of H2O2 (20% and 100% of the stoichiometric dose), temperature (20ºC and 70ºC) and presence of NaCl (56.3 mM) was investigated. The homologue profiles of total PCDD/Fs in oxidized samples (4 h), together with that found in the untreated 2-CP solutions, are shown in Figure 3.20.
10000 Untreated 2-CP 4h (100% estq, 20ºC))
) 4h (100% estq, 70 ºC) 1 - 1000 4h (100% estq, 70ºC, NaCl) 4h (20% estq, 20ºC) 4h (20% estq, 20ºC, NaCl) 100 Concnetration (pg L (pg Concnetration
10
1 TCDD PeCDD HxCDD HpCDD OCDD TCDF PeCDF HxCDF HpCDF OCDF
Figure 3.20. Comparison of the homologue profiles of total PCDD/Fs in the untreated 2- CP solutions and after 4h of Fenton oxidation.
102 Results and Discussion
According to data reported in Figure 3.20, three groups of homologues (OCDD, HpCDD and TCDF) were detected at very low concentrations (4.6-12.2 −1 pg L ) in the untreated water solutions containing 2-CP. For 100% H2O2, although new groups of homologues appeared in the oxidized samples in comparison to the untreated ones, taking into account the low PCDD/Fs concentration and the experimental error, a noticeable formation of PCDD/Fs could not be observed. For TCDD and TCDF it could be noted a common trend of decreasing in their concentrations after increasing the working temperature up to 70ºC. On the other hand, after 4 h of treatment with 20% H2O2 relatively large quantities of PCDD/Fs were observed, resulting in a total PCDD/Fs concentration (312.8 pg L-1) 11 times higher than in the untreated sample. Results showed a preferential formation of PCDFs over PCDDs indicating that the formation of PCDFs intermediates was favored against the corresponding ones for PCDDs.
Moreover, the total concentration of PCDD/Fs (2092 pg L-1) was 74.4 times higher in the presence of NaCl, highlighting the positive influence of ubiquitous chloride in the reaction medium on the potential formation of PCDD/Fs. As it was mentioned in section 3.2.3, according to Pignatello (1992), the formation of - chlorine radicals from the reaction between OH and Cl (r17 to r19), was noticeable for chloride concentrations above 10 mM. The produced Cl could contribute to the formation of PCDD/Fs through the chlorination of either precursors present in the reaction medium or lower chlorinated PCDD/Fs, leading to the formation of higher chlorinated congeners that were not expected from direct condensation reactions between the reactants.
Among total PCDD/Fs, congeners with chlorine at 2,3,7,8 positions are of particular interest due to their high toxicity and potential effects on human health. Their concentrations in the untreated 2-CP model solutions as well as in the oxidized samples at 4 h are shown in Figure 3.21.
103 Chapter 3
14 Untreated 2-CP 12 4h (100% estq, 20ºC) ) 1 - 10 4h (100% estq, 70ºC) 4h (100% estq, 70ºC, NaCl) 8 4h (20% estq, 20ºC) 6 4h (20% estq, 20ºC, NaCl)
Concnetration (pg L (pg Concnetration 4
2
0
Figure 3.21. Congener profile of 2,3,7,8-PCDD/Fs in the untreated 2-CP solutions and after 4h of Fenton oxidation.
It can be observed in Figure 3.21 that only two 2,3,7,8-chlorinated congeners, 2,3,7,8-TCDF and OCDD, were detected at very low concentrations (1.2−4.6 pg L−1) in the untreated 2-CP solution. Taking into account the low
PCDD/Fs concentrations and the experimental error, only with 20% H2O2 the two congeners found in the untreated 2-CP solution remained in the reaction medium after 4 h of treatment, although at low concentrations (2.6 and 6.9 pg L-1, respectively). Therefore, the formation of the less toxic PCDD/Fs congeners was favored under the operating conditions used in this work. Although PCDD/Fs have been shown to be formed in very small concentrations (i.e. 2092 -1 - pg L for the system with 20% H2O2 in the presence of Cl ), it must be emphasized that these low yields do not eliminate environmental threats, as many compounds, such as PCDD/Fs, are toxic even at very low concentrations. In fact, as it has been previously mentioned, the MCL established by the U.S. EPA is 30 pg L-1 of the equivalent 2,3,7,8-TCDD. In addition, as PCDD/Fs are highly persistent in the environment and their lypophilic nature causes their accumulation through food chains, low and continuous exposure to these compounds can result in adverse effects for biota, including human health.
104 Results and Discussion
The formation of dioxins and related compounds has been previously reported, during the treatment of chlorinated phenols using Fenton-related processes. Fukushima and Tatsumi (2001) observed the formation of OCDD during the photo-Fenton treatment of pentachlorophenol (PCP). Poerschmann et al. (2009) showed the generation of chlorinated biphenyls, diphenyl ethers, benzofurans and related compounds during the Fenton oxidation of 2-CP under substoichiometric conditions. Later studies depicted the formation of similar compounds, including chlorinated dibenzodioxins, as a result of the Fenton-like treatment of several CPs with substoichiometric doses of H2O2 and low quantities of iron (Munoz et al., 2013, 2011). In addition, these results are in agreement with those shown in chapter 3.1 where the formation of PCDD/Fs as a result of the electrochemical oxidation of 2-CP was reported. Therefore, these findings provide further evidence for the importance of selecting proper operating conditions during the treatment of chlorinated pollutants, in particular when PCDD/Fs precursors are present (or may be formed during the treatment) together with the concomitant presence of chloride in the reaction medium.
3.2.6. Proposed reactions pathway for the Fenton oxidation of 2-chlorophenol
Based on the detected products during the Fenton oxidation of 2-CP, a generic scheme of the reaction pathways are proposed in Figure 3.22. The hydroxyl functional group in 2-CP is a highly active ortho/para-directing group and OH are strong electrophilic radicals (Huang & Chu 2012). In addition, although the chlorine atom of the aromatic ring is a deactivating group, it is also an ortho/para-directing group. Therefore, hydroxylation products with predominantly ortho/para substitutions are well expected. On this way, according to the route A of the reaction pathway proposed in Figure 3.22, the nucleophilic addition of OH to 2-CP led to the formation of aromatic byproducts; 2-chlorobenzoquinone was formed by hydroxylation in para position followed by consecutive dehydrogenation, whereas catechol was formed by hydroxylation in ortho position along with the loss of chlorine (Wang & Wang 2009; Boudenne & Cerclier 1999).
105 Chapter 3
OH Cl
(A) Aromaticintermediates 2-CP (B) PCDD/Fs
OH Cl + OH + OH + OH - H - Cl OH O OH OH Cl Cl OH Cl + Cl-phenoxy radical OH OH -H2O Cl O Cl-phenoxy radical DHCD radical 2-CP 2-chlorobenzoquinone catechol Coupling reactions
O OH OH Cl
(C) Ring Cleavage OH Cl Cl-phenoxyphenol Cl-biphenyl O O O -HCl +H· HO -Cl HO OH +Cl OH O O OH H· O Cl Maleic acid Fumaric acid O Cl Cl O PCDDs O O +Cl H H3C HO OH O OH OH O Formic acid Acetic acid Oxalic acid Cl PCDFs Cl
Figure 3.22. Scheme for the formation of intermediate products in the Fenton oxidation of 2-CP.
On the other hand, the route B, which deals with the formation of PCDD/Fs, has been proposed according to results previously reported in the literature and the coupling-reaction aromatic byproducts identified. On this way, the electrophilic addition of OH to CPs leads to the formation of chlorosubstituted dihydroxycyclohexadienyl (DHCD) radicals (Poerschmann et al. 2009). The consequent elimination of water from DHCD radicals gives resonance-stabilized Cl-phenoxy radicals (Duesterberg & Waite 2007). Oxidative coupling reactions of Cl-phenoxy resonance-stabilized radicals with either other radicals or molecules result in the formation of biphenyl- and diphenyl ether- type intermediates, as well as phenoxyphenols (Poerschmann et al. 2009; Altarawneh et al. 2009). According to Altarawneh et al. (2009). PCDF formation is considered to take place exclusively from the condensation of two radicals, whereas PCDD involves additionally molecules/radical and molecules/molecule type coupling.
106 Results and Discussion
Consecutive chlorination of the condensation products contributed to the higher chlorinated congeners that were not expected from direct condensation reactions between the reactants (Briois et al. 2007). Finally, the aromatic ring cleavage, route C, led to the formation of aliphatic carboxylic acids.
107 Chapter 3
3.3. Electrochemical and Fenton Oxidation of Landfill Leachates
Results from chapters 3.1 and 3.2 dealing with the electrochemical and Fenton oxidation of model aqueous solutions containing 2-CP, a known PCDD/Fs precursor, depicted the potential formation of PCDD/Fs as reaction byproducts under certain operating conditions applied. In a further step, this chapter focuses on the assessment of PCDD/Fs when Fenton and electrochemical oxidation were applied to the remediation of high polluted landfill leachates, characterized by high concentration of organic matter and concomitant presence of high amounts of chloride. The leachates used in this study came from a municipal solid waste (MSW) landfill located in the municipality of Meruelo, Cantabria (North of Spain). This landfill site serves a population of over 591,888 inhabitants and generates daily between 500 and 800 m3 of leachate depending on the season. The leachates are accumulated in equalization tanks and then pumped through a 500 µm rotary sieve to a storage tank for their later treatment. Nowadays, leachates are treated in a first stage by a nitrifying-denitrifying biological process and afterwards, in a second stage involving ultrafiltration membranes where the treated leachates are separated from the biomass.
3.3.1. Characterization of landfill leachate samples
The average physico-chemical characteristics of the landfill leachates used in this work, which were collected from June 2010 to November 2011, were summarized in Table 3.2. Regarding general parameters, the pH of the leachate was slightly alkaline, with a mean value 7.8. pH values in the range of 7.5 to 9 are typical of the later stable methanogenic phase of a landfill (Kjeldsen et al. 2002; Ahmed & Lan 2012). The average conductivity was approximately 15 mS cm-1 and the total suspended solids (TSS) were present in low concentration, 115 mg L-1, as a result of the storage of the leachates in collection ponds that allows solids to sediment.
108 Results and Discussion
Table 3.2. Physico-chemical properties of the leachate samples used in this study.
Properties Mean value (± sd) pH 7.8 ± 0.4 Conductivity (mS cm-1) 14.6 ± 1.3 TSS (mg L-1) 115 ± 26.9 TOC (mg L-1) 934.2 ± 137 COD (mg L-1) 1981 ± 263 -1 BOD5 (mg L ) 655
BOD5/COD 0.33 ± 0.04 N-NH4+ (mg L-1) 984.9 ± 56 - -1 NO2 (mg L ) 12.84 - -1 NO3 (mg L ) 11.23 ± 1.83 Cl- (mg L-1) 2205.3 ± 262.3
3- -1 PO4 (mg L ) 48.1 ± 13.3
2- -1 SO4 (mg L ) 98.4 ± 85.7
The organic matter content was described by three parameters, namely
TOC, COD and biological oxygen demand (BOD5). The average TOC content was -1 -1 934 mg L and COD value was around 1981 mg L . In addition, the mean BOD5 -1 of the leachates was 655 mg L being the BOD5/COD ratio equal to 0.33. Low
COD and BOD5/COD ratios are characteristic of the methanogenic phase of landfills, and are explained by low biological activity, which in turn is a result of low concentrations of volatile fatty acids and relatively higher amounts of fulvic and humic-like compounds (Oman & Junestedt 2008; Ahmed & Lan 2012). The + -1 N-ammonium (N-NH4 ) concentration was 985 mg L . Other types of measured nitrogen compounds include nitrate and nitrite, which showed lower concentration values, 11.2 and 12.8 mg L-1 respectively, than ammonium, indicating that this compound represented the larger fraction of the total nitrogen amount. Ammonium is a stable compound under methanogenic conditions (Burton & Watson-Craik 1998) and therefore has been identified as the most significant component of leachates for the long term (Kjeldsen et al. 2002). Besides, ammonium is in equilibrium with ammonia, which is considered the primary cause of toxicity of municipal landfill leachates (Oman & Junestedt 2008). The average concentration of inorganic macrocomponents such as
109 Chapter 3
chloride, phosphate and sulfate were also included in Table 3.2. Chloride ions showed the highest value, approximately 2205 mg L-1, whereas the concentrations of phosphate and sulfate were a great deal lower, with respective values of 48.1 and 98.4 mg L-1. Sulfate concentration is expected to be lower in the methanogenic phase of a landfill as a result of its reduction to sulfide by means of microorganisms activity (Kjeldsen et al. 2002). According to the aforementioned properties, the leachates used in this work can be classified as mature leachates mainly characterized by a COD of 1981 mg L-1, slightly basic + -1 pH (7.8), low BOD5/COD ratio (0.33) and high N-NH4 concentration (985 mg L ). Qualitative screening (non-target analysis) of organic compounds in leachate samples was performed by GC-MS in full scan mode. Confirmation of all structural assignments for the identified compounds was made using the NIST08 spectra library as well as analytical standards for some compounds. Taking as an example one of the leachate samples analyzed, Table 3.3 classifies those detected compounds showing a match percentage with the library higher than 70%. The sum of the peak area contribution of each compound to the total area, included in Table 3.3, represented the 88% of the total detected compounds.
Table 3.3. Organic compounds detected by GC-MS (full scan mode) in the leachates used in this study.
Retentio Matc Contributio Compound n time h (%) n (%) (min) Alkanes
Eicosane 97 23.18 0.57 Hexadecane 95 29.53 0.14 Alcohols
2,4,7,9-tetramethyl-5-decyne-4,7-diol (Surfinol 104 H) 72 15.15 0.07 Ethanol, 2-[2-[4-(1,1,3,3- 70 25.07 0.17 tetramethylbutyl)phenoxy]ethoxy]- Aldehydes and ketons
4-[(4'-Methylphenyl)thio]acetophenone 90 19.75 0.45 1-{5-Methyl-2-[(4- 70 21.08 0.38 methylphenyl)sulfanyl]phenyl}ethanone
110 Results and Discussion
Table 3.3. Organic compounds detected by GC-MS (full scan mode) in the leachates used in this study (Cont.).
Aldehydes and ketons
7,9-di-tert-butyl-1-oxaspiro[4.5]deca-6,9-diene-2,8- 97 22.35 0.34 dione 1-Methyl-oestra-1,3,5(10)-trien-nor-17-ketone 80 25.66 0.26 5-Methyl-2-phenyl-1,5-benzothiazepine-4(5H)-one 90 28.30 0.10 Aliphatic acids and esters
Isocyanic acid, cyclohexyl ester 98 7.52 0.43 Cyclohexanecarboxylic acid 96 10.14 0.46 Hexanoic acid, 3,5,5-trimethyl- 78 10.42 0.19 3-Ethylheptanoic acid 90 11.77 0.17 Acetic acid, [(2-methylpropyl)thio]- 70 33.28 0.24 Aromatic carboxylic acids and esters
Benzeneacetic acid 70 12.47 0.11 Benzenepropanoic acid 92 13.93 0.25 Benzoic acid, p-tert-butyl- 95 16.49 1.22 Propanoic acid, 2-methyl-3-[4-t-butyl]phenyl- 94 19.31 0.49 Benzenepropanoic acid, 3,5-bis(1,1-dimethylethyl)-4- 83 23.31 4.62 hydroxy Polyaromatic and heterocyclic compounds
6-Chloro-1H-Indole 96 17.08 0.29 1-(4-Isopropylphenyl)adamantane 80 26.02 0.06 7H-Dibenzo[c,g]carbazole 70 28.15 0.11 1-Ethoxyanthracene-9,10-dione 86 28.56 0.54 Anthracen-9-yl(trimethyl)silane 90 29.16 0.45 Terpenoids
Beta-eudesmol) 89 21.14 0.20 Sclareolide 70 24.32 0.11 Valencene 90 28 0.06 Phthalic acid esters
Diethyl Phthalate 96 17.93 0.08 Isobutyl phthalate 87 21.66 0.16
111 Chapter 3
Table 3.3. Organic compounds detected by GC-MS (full scan mode) in the leachates used in this study (Cont.).
Phenolic compounds
Phenol 94 7.26 0.12 Phenol, 4-methyl- 96 9.11 3.48 Phenol, 2,4-dichloro- 98 10.92 0.04 Phenol, 4-chloro- 97 11.49 0.17 Phenol, 4-(1,1-dimethylethyl)- 94 13.14 0.15 Phenol, 2,4-dichloro-6-methyl- 93 13.46 0.08 Phenol, 2,4,6-trichloro- 99 14.18 0.17 Biphenol A (4-[2-(4-hydroxyphenyl)propan-2- 95 25.46 14.08 yl]phenol) 3,3'-Dichlorobisphenol A (2-chloro-4-[2-(3-chloro-4- 96 26.19 0.16 hydroxyphenyl)propan-2-yl]phenol) 2-[(E)-2-(4-tert-Butylphenyl)ethenyl]phenol 80 28.61 0.09 Benzene derivatives
Benzene, 1-(1-hydroxyethyl)-4-isobutyl- 91 14.90 0.07 Benzene, 1,1'-(chloromethylene)bis 70 21.48 0.08 Benzene, 1,1'-(1-ethyl-1,2-ethenediyl)bis[4-methoxy- 86 28.67 1.05 Benzothiazoles
2(3H)-Benzothiazolone 94 19.22 3.02 Sulfonamides
N-Butyl benzenesulfonamide 97 20.89 38.70 Pharmaceutical and personal care products (PPCPs)
Methyl ester of Ibuprofen (methyl 2-[4-(2- 98 17.13 0.09 methylpropyl)phenyl]propanoate) Ibuprofen (2-[4-(2-methylpropyl)phenyl]propanoic) 99 18.55 13.31 Propyphenazone 83 22.95 0.05 Pesticides
Diethyltoluamide (DEET) 96 17.75 0.20 Parathion 90 26.84 0.12 Pentachloroaniline 91 27.15 0.12 Total 88 %
112 Results and Discussion
As a summary of the information reported in Table 3.3, the depicted compounds were classified into families of substances in order to have a global qualitative overview of the organic compounds detected in the leachates studied in this work, Figure 3.23.
Pharmaceutical and Aromatic carboxylic acids Aldehydes and ketons personal care products and esters Aliphatic acids and esters 6.7% 13.5% Aromatic carboxylic acids and esters Phenolic compounds Polyaromatic and heterocyclic hydrocarbons 18.5 % Phenolic compounds Benzene derivatives Sulfonamides 38.7 % Benzothiazoles Sulfonamides Pharmaceutical and personal care products Others
Figure 3.23. Dominant groups of organic compounds found in the landfill leachates.
According to Figure 3.23, the dominant groups of compounds listed in Table 3.3 included aldehydes and ketons (1.7%), aliphatic (1.5%) and aromatic acids and esters (6.7%), polyaromatic and heterocyclic compounds (1.4%), phenolic compounds (18.5%), benzene related compounds (1.2%), benzothiazoles (3%), sulfonamides (38.7%) and pharmaceutical and personal care products (PPCPs) (13.5%). Among them, sulfonamides, as is the case of n- butyl benzenesulfonamide (NBBS) accounted for 38.7% of the total detected compounds, followed by phenolic compounds (18.5%), which were dominated by bisphenol A (BPA), and PCPPs (13.5%). NBBS and BPA are one the most common compounds detected in landfill leachates as was shown by Baderna et al. (2011), who found them as the most recurrent chemicals in the leachates form an industrial landfill in 55 monitoring campaigns during 11 years.
Over the past decade there has been growing concern regarding the presence of biologically active contaminants in the aquatic environment, including endocrine disruptors compounds (EDCs), pharmaceuticals, personal care products and other substances. Emerging contaminants (EC) are defined as natural or synthetic substances that are not commonly monitored in the environment, but may induct undesirable effects on human and ecosystems (Stuart et al. 2012). The largest contributing sources of ECs to aquatic
113 Chapter 3
environments are wastewater treatment plants for domestic sewage, wastewater from hospital effluents, chemical manufacturing plants, livestock and agriculture (De la Cruz et al. 2012). ECs comprise different compounds such as disinfectants, industrials, detergents and PPCPs, list that is likely to be expanded with the development of their analytical techniques (Meffe & de Bustamante 2014).
For most EC, occurrence, risk assessment and ecotoxicological data are not available, and it is difficult to predict their fate in the aquatic environment (Petrovic 2003). In the European context, surface water quality standards are regulated under the Directives 2000/60/EC, 2008/105/EC, and 2013/39/EU which modifies the earlier ones with regard to contaminants, defining new priority substances. However due to the lack of information on toxicity and environmental impacts, a large number of contaminants, especially organic compounds, are not included in the list of chemicals to be monitored (Fatta- Kassinos et al. 2011; Boonnorat et al. 2014). Although EC concentrations encountered in the environment are quite low, ranging between ng L−1 and μg L−1, a continuous exposure of the aquatic communities may result in potential harmful effects (Meffe & de Bustamante 2014). Besides, the complete removal of some EC in sewage treatment plants cannot be assured by biological treatment methods, showing bioaccumulation and toxic effects in aquatic and terrestrial ecosystems (De la Cruz et al. 2012).
Leachates from MSW landfills can be contaminated with several EC since a great quantity of them are found in some household and industrial waste. NBBS is an EC that is being evaluated for its potential human and environmental health risks (Eggen et al. 2013). NBBS is commercially used as plasticizer in the polymerization of polyamides and in the synthesis of sulfonyl carbamate herbicides. As a result of its wide use in the industry and household products, and considering its polar properties, it is widespread in the aqueous environment (Eggen et al. 2010). In addition, because of its stability and persistence in the environment, relatively high concentrations have been found in wastewaters (Roell & Baniahmad 2011). With respect to the leachates studied in this work, NBBS depicted an average concentration of 1289.2 ± 9.8 µg L-1.
Phenolic compounds are detected in landfill leachates as a result of their extensive use in industrial products, i.e. as plastic softeners, lacquers, paints and as antioxidants in fuel and oils (Eggen et al. 2010). In particular, BPA was
114 Results and Discussion
the dominant compound within phenolic substances identified in the leachates used here. BPA is an EDC that has the ability to mimic hormones and as a result might interfere or disorder their normal functions (Dorival-García et al. 2012). Although it is well established that long term exposure to EDCs could have adverse health effects at concentrations as low as ng L-1, their quantification remains a major challenge that has yet to be addressed. The European Food Safety Authority (EFSA) established a maximum total daily intake of 50 µg of BPA kg-1 of body weight d-1 (EFSA, 2006).
BPA has been extensively used in making polycarbonate plastic and epoxy resins that are used in a great variety of domestic products (Umar et al. 2013). Numerous studies have confirmed the presence of BPA in packaged foods and its leaching from food containers (Dorival-García et al. 2012). In addition to plastic waste, which is one of the major components of MSW, paper waste is another potential source of BPA since it is extensively used as additive in adhesives and coating composites in paper products (Xu et al. 2011). Such wide use of BPA has made it into a common contaminant in landfill leachates, showing concentrations in the range of 1 µg L-1 to 5 mg L-1 (Urase & Miyashita 2003). Values within this range were found in this work, where BPA showed an average concentration of 1144.4 ± 74 µg L-1.
The occurrence and fate of pharmaceuticals in water have raised especial interest as a result of the unknown environmental impact and possible damages to the flora and fauna of the aquatic systems (Méndez-Arriaga et al. 2010). In landfills, the presence of PPCPs has been largely neglected. Once discarded in MSW landfills, these compounds may undergo degradation, adsorption, or percolation through the landfill entering the leachates (Musson & Townsend 2009). As it was expected, ibuprofen was the main PPCP detected in the leachates from this work since it is amongst the most consumed pharmaceuticals (Paíga et al. 2013). The worldwide occurrence of ibuprofen in the aquatic environment has been reported with concentrations between 10 ng L-1 and 169 µg L-1 (Méndez-Arriaga et al. 2010). In 2010, the European Commission included ibuprofen in a list of 19 possible new priority contaminants (Paíga et al. 2013).
115 Chapter 3
3.3.2. Determination of PCDD/Fs in landfill leachates
Landfills are considered PCDD/Fs reservoirs in the environment (Fiedler 2003; Weber et al. 2008; Weber et al. 2011), and as a result these compounds have been detected in the leachates generated from them (Casanovas et al.
1994; Pujadas et al. 2001). The presence of PCDD/Fs in the leachate samples (S1,
S2 and S3) from a MSW landfill, whose average physico-chemical properties were shown in Table 3.2, was evaluated. The homologue profiles of total PCDD/Fs in the cited leachate samples are depicted in Figure 3.24. Regarding the most toxic congeners, 2,3,7,8-PCDD/Fs, their profiles in the same samples are displayed in Figure 3.25.
3800 S₁ S₂ S₃ 2800
1800 ) 1 -
L 800 800
pg 250 ( 250
200
150 Concentration
100
50
0 TCDD PeCDD HxCDD HpCDD OCDD TCDF PeCDF HxCDF HpCDF OCDF
Figure 3.24. Homologue profile of total PCDD/Fs in three landfill leachate samples.
116 Results and Discussion
4000
3000 S₁ S₂ S₃ 2000 ) 1
- 1000 L pg ( 0 50
40
Concentration 30
20
10
0
Figure 3.25. Congener profile of 2,3,7,8-PCDD/Fs in three landfill leachate samples.
-1 The total concentration of PCDD/Fs ranged from 4479 pg L for leachate S3 -1 to 4981.8 pg L for leachate S1. As can be observed in Figure 3.24, all homologue profiles were dominated by PCDDs, with OCDD and HpCDD as the main groups, which represented between 68% and 75.2%, and 18.6% and 22.6%, respectively, of the total concentration of PCDD/Fs. Regarding PCDFs, samples S2 and S3 showed similar homologue profiles governed by high chlorinated homologue groups, whereas sample S1 depicted higher concentrations of low chlorinated ones.
With respect to the most toxic congeners (Figure 3.25), the total -1 concentration of 2,3,7,8-PCDD/Fs ranged from 3710 pg L for leachate S3 to -1 4244 pg L for leachate S2. The congener profiles were dominated by OCDD, which accounted between 83% (S1) and 86.7% (S2) of the total 2,3,7,8-PCDD/Fs concentration. The second prevalent congener was 1,2,3,4,6,7,8-HpCDD that contributed between 11% for S2 and 13.8% for S1 of the sum of 2,3,7,8- PCDD/Fs. The sum of 2,3,7,8-PCDD/Fs concentrations depicted between 81.7% for S1 and 86.7% for S2 of the total concentration of PCDD/Fs due to the high concentration of congener OCDD. Among PCDDs, toxic congeners represented
117 Chapter 3
between 83.6% (S3) and 88.2% (S2) of the total PCDDs concentration, whereas only 30.4% (S1)-57.9% (S3) of the total PCDFs were 2,3,7,8-chlorinated.
Comparison with PCDD/Fs profiles in leachates reported in the literature
Although scarce, there are some works in the literature that report the concentration of PCDD/Fs in landfill leachates. Casanovas et al. (1994) reported the 2,3,7,8-PCDD/Fs profile of a high polluted landfill leachate characterized by a general trend of increasing in PCDDs concentrations with chlorination degree, whereas PCDFs concentrations did not show a common pattern. The authors found high variability and strong dependence on the origin of the sample. Although the reported PCDD/Fs concentrations were higher than those displayed in the present work, the profile was dominated by the same congeners, 1,2,3,4,6,7,8-HpCDD and OCDD. Pujadas et al. (2001) presented the characterization of a leachate sample from a MSW landfill with a 2,3,7,8- PCDD/Fs profile very close to our findings. In addition, Choi and Lee (2006) and Ham et al. (2008) also reported PCDD/Fs profiles in municipal landfill leachates with clear dominance of 1,2,3,4,6,7,8-HpCDD and OCDD congeners.
The 2,3,7,8-PCDD/Fs profile found in the leachates studied in this work was compared more in depth with those found in the leachates from the aforementioned literature (and summarized in Table 3.4) by principal component analysis (PCA).
Table 3.4. Leachate samples used in the PCA analysis.
Sample name Landfill type Reference
S1, S2, S3 MSW This work CAS Unkonwn Casanovas et al. (1994) PUJ MSW Pujadas et al. (2001) MSW (can also containing A, B, C, D, E Choi & Lee (2006) incineration residues) BC, UJ, CS, BS-1, SA, SG, MSW (can also containing GY, GG, ST, BS-2, GY, incineration and industrial Ham et al. (2008) MC residues)
118 Results and Discussion
PCA analysis was used to derive two new components (principal components, PC) as a linear combination of the original variables (17 2,3,7,8- PCDD/Fs congener concentrations). Each sample was ascribed a score in each component, and the results were represented in a biplot (Figure 3.26) to observe the similarity between leachates, since samples with similar patterns must be located close together. In this work, PCA was used from the logarithms of the congeners concentrations to extract two principal components, PC1 and PC2, which explained 86.3% of the variability of the data, accounting for 70.3% (PC1) and 16% (PC2), respectively.
3 CAS 2.5 1 1 2,3,7,8-TCDD
2,3,4,6,7,8-HxCDF JY 6 2,3,7,8-TCDF 0.5 UJ 2 1,2,3,4,7,8-HxCDF 5 PUJ SA 1,2,3,4,6,7,8-HpCDD ST S1 OCDD 1,2,3,4,6,7,8-HpCDF CS MC SG S2 0 GY BS-1 S3 1,2,3,4,7,8,9-HpCDF GG
PC2 (16 %) (16 PC2 E BS-2 4 BC OCDF -0.5 A
C D 3 -1 B
1,2,3,7,8-PeCDF
-1.5 -1 -0.5 0 0.5 1 1.5
PC1 (70.32 %) Figure 3.26. PCA biplot from the logarithms of 2,3,7,8-PCDD/Fs concentrations found in different landfill leachates samples.
As can be seen in the biplot represented in Figure 3.26, all PCDD/Fs congeners were found to be positively correlated to the first component PC1. Then, PC1 represents the overall concentration of PCDD/Fs. The second component PC2 was positively correlated with most PCDD/Fs, especially with congeners 2,3,7,8-TCDD, 2,3,4,6,7,8-HCDF and 2,3,7,8-TCDF, and negatively
119 Chapter 3
correlated with OCDF and 1,2,3,7,8-PeCDF. Different groups of samples were identified in the biplot from Figure 3.26 and the congener profiles (expressed as contribution (%) to the total 2,3,7,8-PCDD/Fs concentration) of each group of samples were compared in Figure 3.27. According to the information reported in both Figures, next conclusions are discussed. At the upper-right side of the plot, the sample CAS (group 1) showed great positive values of both PC1 and PC2. Therefore, this sample, according to Casanovas et al. (1994), was characterized by the highest PCDD/Fs concentration (12817.2 pg L-1), with relative high contribution of 2,3,7,8-TCDD and 2,3,7,8-TCDF and, with low contribution of OCDF and non-presence of 1,2,3,7,8-PCDF (Figure 3.27a). Group
2, comprised of leachate samples S1, S2, S3 (this work) and PUJ, which came from MSW landfills, depicted positive correlation with both PC1 and PC2 (Figure 3.26). These samples reported high PCDD/Fs concentrations (705.1-4244.3 pg L- 1) and a profile dominated by the congeners 1,2,3,4,6,7,8-HpCDD and OCDD (Figure 3.27b).
Next samples group, namely group 3, was comprised of samples B, C and D that showed positive values of PC1 and negative values of PC2 (Figure 3.26). Accordingly, these samples were described by relative high PCDD/Fs concentration (719.6-1332.4 pg L-1) and high contributions of OCDF and 1,2,3,4,6,7,8-HpCDF, as well as of 1,2,3,4,6,7,8-HpCDD and OCDD, to the total concentration (Figure 3.27c). Besides, the presence of PeCDFs and HxCFs is worth to mention. Sample A, displayed also in Figure 3.27c, showed a similar profile but with great lower concentration (173.2 pg L-1). Although these samples came from landfills mainly composed of MSW, they could also contain incineration residues, principally bottom ash (Choi & Lee 2006). This fact could explain the high contribution of some PCDFs congeners to the total concentration observed in Figure 3.27c, since the abundance of PCDFs is commonly linked to thermal processes such as those observed in air emissions from MSW incinerators (Rappe 1994).
120 Results and Discussion
100 100 CAS S₁ S₂ S₃ PUJ 80 a 80 b 60 60 40 40 Contribution (%) Contribution 20 (%) Contribution 20 0 0 100 100 c B C D A d E BC BS-1 GG BS-2 80 80 60 60 40 40 Contribution (%) Contribution
20 (%) Contribution 20 0 0 100 100 UJ CS SA MC SG GY ST e f JY 80 80 60 60 40 40 Contribution (%) Contribution Contribution (%) Contribution 20 20 0 0
Figure 3.27. Congener profiles of 2,3,7,8-PCDD/Fs in the landfill leachate samples displayed in the biplot from Figure 3.26.
121
Results and Discussion
Group 4 included samples E, BC, BS-1, GG and BS-2 that were negatively correlated with both PC1 and PC2, with the exception of sample BC, which depicted positive values for PC1. Therefore, sample BC had the highest concentration of the group (640.1 pg L-1), whereas the concentration of the remaining samples was in the range of 279.3 to 555.3 pg L-1. Congener profiles of samples from group 4 were dominated by 1,2,3,4,6,7,8-HpCDD, OCDD and also by OCDF (Figure 3.27d). Although most of the samples included in this group came from landfills composed mainly by municipal waste, they also contained incineration and industrial waste (Ham et al. 2008).
The next group of samples, namely group 5, encompassed samples UJ, CS, SA, SG, GY, ST and MC, which were negatively correlated with PC1 and positively with PC2. As a result, this group of samples was described by low concentration of PCDD/Fs (25.8-322.2 pg L-1), with the exception of sample UJ that possessed a relative high PCDD/Fs concentration (959.3 pg L-1). Samples from group 5 were characterized by a profile dominated by 1,2,3,4,6,7,8- HpCDD, OCDD, 1,2,3,4,6,7,8-HpCDF and OCDF, and additionally with certain presence of low chlorinated PCDFs (Figure 3.27e). These samples, as in the case of those from group 3, came from MSW landfills with presence of incineration waste (Ham et al. 2008), that could explained the relative high contribution of OCDF to the total PCDD/Fs concentration. Finally, sample JY, which was the most negatively correlated with PC1, displayed consequently the lowest concentration, 16.2 pg L-1.
Toxic Equivalents (TEQ) of the detected 2,3,7,8-PCDD/Fs
The concentrations of 2,3,7,8-PCDD/Fs reported in the leachate samples analyzed in this work (Figure 3.25) were converted into one value of TEQ using both the I-TEFs and the WHO-TEFs. The contribution of each congener to the total TEQ, I-TEQ and WHO-TEQ, is depicted in Figure 3.28 and Figure 3.29, respectively.
122
Results and Discussion
100 OCDF 1,2,3,4,7,8,9-HpCDF 1,2,3,4,6,7,8-HpCDF 80 2,3,4,6,7,8-HxCDF TEQ (%) - 28.1 23.2 1,2,3,6,7,8-HxCDF 60 19.7 1,2,3,4,7,8-HxCDF 2,3,4,7,8-PeCDF 1,2,3,7,8-PeCDF 40 37.0 32.8 35.5 2,3,7,8-TCDF OCDD 20 1,2,3,4,6,7,8-HpCDD
Congener contribution to I to Congener contribution 1,2,3,7,8,9-HxCDD 0 1,2,3,6,7,8-HxCDD S₁ S₂ S₃ 1,2,3,4,7,8-HxCDD
Figure 3.28. Contribution of individual 2,3,7,8-PCDD/Fs congeners to the total I-TEQ.
100 OCDF 1,2,3,4,7,8,9-HpCDF 1,2,3,4,6,7,8-HpCDF 80 11.0 8.7 2,3,4,6,7,8-HxCDF TEQ (%) - 7.4 1,2,3,6,7,8-HxCDF 60 1,2,3,4,7,8-HxCDF 2,3,4,7,8-PeCDF 46.2 46.5 41.3 1,2,3,7,8-PeCDF 40 2,3,7,8-TCDF OCDD 20 1,2,3,4,6,7,8-HpCDD 9.3 10.1 10.4 1,2,3,7,8,9-HxCDD 9.5 10.1 10.4 Congener contribution to WHO to Congener contribution 0 1,2,3,6,7,8-HxCDD S₁ S₂ S₃ 1,2,3,4,7,8-HxCDD
Figure 3.29. Contribution of individual 2,3,7,8-PCDD/Fs congeners to the total WHO-TEQ.
123 Chapter 3
When expressing PCDD/Fs concentration in TEQ values using the I-TEFs (Figure 3.28), the TEQ of the leachate samples ranged from 13.1 pg I-TEQ L-1 for -1 S2 to 17.2 pg-ITEQ L for S1. The congener 1,2,3,4,6,7,8-HpCDD contributed with the higher percentage (32.8-37%) to the TEQ, followed by OCDD which accounted for 19.7%-28.1% of the total TEQ. Although OCDD was the most abundant congener, it contributed to the TEQ value in lower extent than 1,2,3,4,6,7,8-HpCDD since its I-TEF is 10 times lower. On the other hand, when the WHO-TEFs were applied (Figure 3.29) instead of the international ones, the TEQ of the leachates decreased to 10-13.7 pg WHO-TEQ L-1, since the WHO-TEF corresponding to OCDD was 3.3 times lower than the corresponding I-TEF. Therefore, as can be seen in Figure 3.29, the contribution of OCDD to the total TEQ decreased to 7.4%-11%. In this case, the TEQ value was dominated mainly by 1,2,3,4,6,7,8-HpCDD (41.3%-46.5%) followed by 1,2,3,4,7,8-HxCDD and 1,2,3,6,7,8-HxCDD, which contributed to the total TEQ between 9.3% and 10.4%.
Considering the two dominant congeners, OCDD and 1,2,3,4,6,7,8-HpCDD, a correlation was found between their concentration in landfill leachates samples and their corresponding TEQ. The expressions were obtained by non- linear regression with XLSTAT for Microsoft Excel using the results from this work and including some of the PCDD/Fs concentrations reported in the leachates from the aforementioned literature (Pujadas et al. 2001; Choi & Lee 2006; Ham et al. 2008), in order to have a wide range of data. The following equations were found for both I-TEQ (eq 7) and WHO-TEQ (eq 8) with regression coefficients (r2) of 0.94 and 0.91, respectively.
I-TEQ: [pg I-TEQ L-1] = 3.71 [pg 1,2,3,4,6,7,8-HpCDD L-1]1.38 [pg OCDD L-1]-0.89 (eq 7)
WHO-TEQ: [pg WHO-TEQ L-1] = 3.96 [pg 1,2,3,4,6,7,8-HpCDD L-1]1.34 [pg OCDD L-1]-0.9 (eq 8)
The TEQ values predicted with both expressions (eq 7 and eq 8) were plotted versus the experimental data in Figures Figure 3.30 and Figure 3.31, respectively.
124 Results and Discussion
18 Pujadas et al. (2001) 16 Choi et al. (2006) 14 Ham et al. (2008) This work 12
experimental 10 ] 1 - 8 TEQ L
- 6
[pg I 4 2 0 0 2 4 6 8 10 12 14 16 18 [pg I-TEQ L-1] predicted Figure 3.30. Calculated values of I-TEQ obtained from 1,2,3,4,6,7,8-HpCDD and OCDD content vs. experimental values for the leachates analyzed in this work and those reported in the literature.
14 Pujadas et al. (2001) 12 Choi et al. (2006) Ham et al. (2008) 10 This work experimental
] 8 1 - 6 TEQ L - 4
[pg WHO 2
0 0 2 4 6 8 10 12 14 [pg WHO-TEQ L-1] predicted Figure 3.31. Calculated values of WHO-TEQ obtained from 1,2,3,4,6,7,8-HpCDD and OCDD content vs. experimental values for the leachates analyzed in this work and those reported in the literature.
125 Chapter 3
As it is displayed in Figures 3.30 and 3.31, values were located near the diagonal indicating that they satisfactorily comply with the proposed equations. The suggested expressions, eq 7 and eq 8, are useful to determine the TEQ content of landfill leachates with a similar congener profile to that found in this work, from the concentrations of the most abundant congeners, 1,2,3,4,6,7,8- HpCDD and OCDD, in the range from 0.6 to 17.2 pg I-TEQ L-1 (0.5-13.7 pg WHO- TEQ L-1).
Identification of potential sources of PCDD/Fs in the leachates
Landfill leachates are known to be PCDD/Fs reservoirs in the environment (Fiedler 2003; Weber et al. 2008). A large part of POPs, including PCDD/Fs, has been deposited in landfills along the history and many of them, especially those contained within consumer goods are still widely deposited in landfills in industrial countries (Weber et al. 2011). It is estimated that municipal waste may contain 34% of all PCDDs/Fs released to the wastes (Gworek et al. 2013). Municipal solid waste may also be the source of PCDDs/Fs due to presence of organic and inorganic chlorine molecules, which are forerunners of these compounds (Gworek et al. 2013). For a more detailed assessment, a common strategy applied to the identification of potential sources of PCDD/Fs in the environment is the comparison of homologue/congener profiles in environmental samples with those of known sources (Clarke et al. 2008). On this way, the comparison between the congener profile found in leachate samples from this work and similar profiles reported in different samples from the literature is depicted in Figure 3.32.
126 Results and Discussion
100 a. Sewage Sludge Hagenmaier et al. (1994) 80 Stevens et al. (2001) Eljarrat et al. (2003) 60 Fuentes et al. (2007) De la Torre et al. (2011) 40 This work Contribution (%) Contribution 20
0 100 b. PCP formulations PCP (Hagenmaier and Brunner, 1987) 80 PCP-Na (Hagenmaier and Bunner, 1987) PCP-Na (Palmer et al., 1988) PCP-Na (Santl et al., 1994) 60 PCP-Na (Holt et al., 2008) PCP (Holt et al., 2008) 40 This work Contribution (%) Contribution 20
0 100 c. Air and atmospheric Air- Urban (Tysklind et al., 1993) 80 deposition Air- Rural (Cleverly et al., 2007) Air- Remote (Cleverly et al., 2007) 60 Dry deposition (Correa et al., 2006) This work 40 Contribution (%) Contribution 20
0 100 d. Air Air- Urban + high traffic 80 Air- Rural close to MWI Air- MWI zone 60 Air- MWI zone + traffic + industry Air- Industrial close to industrial IWI 40 Air- High Industrial Activity
Contribution (%) Contribution 20
0
Figure 3.32. Comparison of congener profiles for different environmental samples.
127 Chapter 3
The characteristic congener profile shown in the leachates studied in this work, dominated by OCDD and 1,2,3,4,6,7,8-HpCDD, was similar to that frequently reported in the sewage sludge from municipal wastewater treatment plants, Figure 3.32a (Hagenmaier et al. 1994; Stevens et al. 2001; Eljarrat et al. 2003; Fuentes et al. 2007; De la Torre et al. 2011). Sewage sludge from municipal wastewater treatment plants may have some common contaminants to those found in MSW landfills and therefore in their leachates. Moreover, it must be highlighted that the deposit in landfill was the only disposal route of sewage sludge in Cantabria until 2008, when it started to be considered for agricultural application (PFR, 2014). Therefore, it is likely that sewage sludge was an important source of PCDD/Fs in landfill leachates. In addition to the dominance of OCDD and 1,2,3,4,6,7,8-HpCDD, sewage sludge samples are characterized by a PCDDs TEQ/PCDFs TEQ ratio > 1 (Eljarrat et al. 1999), in agreement with the ratios found in the leachates used in this work, which were in the range of 2.4 to 4.4.
The congener profile observed in these leachates also matched the typical PCDD/Fs contamination profile found in PCP formulations, Figure 3.32b (Palmer et al. 1988; Santl et al. 1994). Although the use of PCP has been regulated since 2004 under the Rotterdam Convention (Spain derogated the banned of its use in textiles and treated wood until December 31st 2008 (De la Torre et al. 2011)), large quantities of PCP were used in the past, especially in the timber industry as wood preservative and in the manufacture of pesticides, and this may mean that there is a considerable reservoir of PCDD/Fs from this source (Dyke et al. 1997). Hagenmaier et al. (1986) found that the main source of PCDD/Fs contamination in sewage sludge came from the industrial use of PCP and sodium-pentachlorophenate (PCP-Na) (Clarke et al., 2008). In addition, Horstmann et al. (1993) identified household wastewater as a significant contributor to dioxin-like compounds in sewage sludge, due to clothing contamination as a result of the use of PCP within the textile industry.
Apart from the prevalence of highly chlorinated PCDD/Fs, the leachates from this study showed a typical domination of congener 1,2,3,4,6,7,8-HpCDF over 1,2,3,4,7,8,9-HpCDF, in agreement with PCP formulations from Figure 3.32b. On the other hand, OCDD/OCDF ratios in the range of 6.5 to 10, and HpCDD/HpCDF ratios of 2 were described by Hutzinger and Fiedler (1993) for contaminated PCP formulations. Another indicator that PCDD/Fs come from a PCP source is the ratio of 1,2,3,6,7,8-HxCDD to 1,2,3,4,7,8-HxCDD, which is
128 Results and Discussion
generally > 100 in a PCP mixture (Eljarrat et al. 1999; Stevens et al. 2001). As a summary, Table 3.5 shows different congener ratios commonly used as indicators of PCP as source of PCDD/Fs contamination. Such indicators were calculated for the leachates samples studied in this work and included in Table 3.5, together with those reported in the works depicted in Figure 3.32b. According to the results from Table 3.5, PCP could not be considered a major source of PCDD/Fs contamination in leachates. However, it must be noticed that even within different PCP formulations there is a considerable variability in the congener ratios values depicted in Table 3.5.
Table 3.5. Congener ratios commonly used as indicators of PCP as source of PCDD/Fs contamination.
PCP PCP PCP-Na PCP-Na This work PCP-Na PCP-Na (b) (c) (a) (d) 1,2,3,4,7,8,9-HpCDF/ 0.05 0.10 1,2,3,4,6,7,8-HpCDF > 1 0.08-0.12 0.02 0.22 0.97 0.13 (Stevens et al. 2001) 1,2,3,6,7,8-HxCDD/ 1,2,3,4,7,8-HxCDD > 100 - 0.73 0.97-1 - 45.94 (Eljarrat et al. 1999; 71.87 108.75 Stevens et al. 2001) 1,2,3,4,6,7,8-HpCDD/ 1,2,3,4,6,7,9-HpCDD > 1 0.96-1.13 (Koch et al. 2001) OCDD/OCDF: 6.5-10 5.77 22.22 74.26-87.79 2.25 7.67 (Hutzinger & Fiedler 1993) 1.12 123.4 HpCDD/HpCDF: 2 15 - 31.68 (Hutzinger & Fiedler 1993)
(a) (Hagenmaier & Brunner 1987) (b) Palmer et al. (1988) (c) Santl e al. (1994) (d) (Holt et al. 2008)
The presence of PCDD/Fs in landfill leachates can be also attributed to the atmospheric deposition (Dyke et al. 1997). Once emitted to the atmosphere, PCDD/Fs enter other environmental compartments through wet and dry deposition (Correa et al. 2006). Generally, as shown in Figure 3.32c, HpCDDs
129 Chapter 3
and OCDD have been shown to be the dominant congeners in air and atmospheric deposition in several geographical areas (Lohmann & Jones 1998; Correa et al. 2006; Cleverly et al. 2007). According to Baker and Hites (2000), this general PCDD/Fs atmospheric profile has been proposed to derive from the PCP present in the atmosphere. As an alternative explanation, Tysklind et al. (1993) and Castro-Jiménez et al. (2010) established that since HpCDDs and OCDD are the congeners most pronounced bound to particles, a high concentration of these congeners in air indicates that the pollution is aged and has been transported in air for some time. PCDD/Fs profile in atmospheric deposition can be distinguished from that one of PCP since higher fractions of 1,2,3,4,6,8,9-HpCDF and OCDF are present in PCP formulations (Sundqvist et al. 2009).
Nevertheless, besides the general dominance of HpCDDs and OCDD in atmospheric deposition, samples can show seasonal and geographical variations (Sundqvist et al. 2010). As it was reported by Abad et al. (1997), Figure 3.32d, air samples from zones with industrial activity, high traffic and close to municipal waste incinerators are characterized by higher contributions of PCDFs, especially OCDF. According to Rappe (1994), high abundances of PCDFs are linked to thermal processes such as those observed in air emissions from municipal solid waste incinerators.
On the other hand, Rappe et al. (1999) explained the homologue pattern of PCDD/Fs observed in sewage sludge (dominated by high chlorinated PCDDs) as a result of their natural formation, as it was found for sediments. Moreover, according to (Öberg & Rappe 1992), OCDD and HpCDD were formed in activated sludge in the presence of precursors such as PCP. Besides, (Klimm et al. 1998) depicted the formation of OCDD and HpCDD during semi-anaerobic stabilization of sewage sludge. The congener profile dominated by OCDD was also found in storm sewer water (urban runoff) and in urban street runoff and sediments from storm water (Dai et al. 2007).
As summary from the above comparison, it could be concluded that the disposal of sewage sludge in landfills, which was a common practice in Cantabria until 2008, could be an important source of PCDD/Fs in the studied leachates. Besides, the dominant profile shown in both, leachates and sewage sludge, could be result of the widespread use of PCP in the past within the timber and textile industries and the manufacture of pesticides, among others. Moreover, biochemical formation of PCDD/Fs in activated sludge from
130 Results and Discussion
precursors such as PCP has been reported. Finally, PCDD/Fs in landfills can also result from atmospheric deposition.
3.3.3. Electrochemical oxidation of landfill leachates
The viability of the electrochemical oxidation using BDD electrodes to treat landfill leachates has been assessed in earlier works of the research group where this thesis has been carried out (Cabeza et al. 2007; Anglada et al. 2009; Anglada et al. 2010; Anglada et al. 2011; Alonso 2013). In the base of the satisfactory results previously obtained in terms of COD and ammonium removal, a current density of 900 A m-2 was applied in the electrooxidation experiments carried out with the leachate sample S2. The changes in normalized concentration of the major parameters, namely COD, TOC and ammonium, shown in Figure 3.33, were selected to evaluate the effectiveness of the degradation process.
+ COD TOC NH4 1.0
0.8
0.6
0.4
0.2 Normalized concentration
0.0 0 100 200 300 400
time (min) Figure 3.33. Changes in normalized concentration of COD, TOC and ammonium with
time during the electrochemical oxidation of leachate S2.
As was shown in Figure 3.33, after 6 h of electrochemical treatment, nearly 95% of COD and 90% of TOC were removed, whereas ammonium was completely degraded. At this time, COD was therefore reduced to below disposal limits (160 mg L-1, RD 849/1986). COD and TOC concentration profiles displayed similar trends with time and their oxidation took place from the beginning of the treatment. On the contrary, ammonium depicted a
131 Chapter 3
concentration profile with an initial delay, in agreement with an oxidation mechanism in which the oxidation of organic compounds takes place at higher extent during the initial stages. Afterwards, in a subsequent stage, chlorine evolution at the anode gains importance and therefore the indirect oxidation of ammonium by active chlorine does as well (Cossu et al. 1998; Pérez et al. 2012). Due to the higher selectivity of active chlorine towards ammonium than towards organic compounds, chlorine evolution on the anode surface led to lower efficiencies of the OH mediated oxidation of organics at the BDD surface (Anglada 2010).
In addition to the major parameters described above, the concentration of the two main organic contaminants identified in the raw leachates as a result of their qualitative non-target analysis, namely n-butyl benzenesulfonamide (NBBS) and bisphenol A (BPA), was checked during the electrochemical treatment (Figure 3.34).
NBBS BPA 1400
1200 ) 1 - 1000
800
600
400
Concentration (µg L (µg Concentration 200
0 0 50 100 150 200 time (min)
Figure 3.34. Changes in NBBS and BPA concentrations during the electrochemical
treatment of leachate S2.
According to Figure 3.34, although both compounds showed similar trends with time, BPA was almost completely removed (95%) within the first 30 min of reaction, whereas only 52% of NBBS was degraded at that time. Although NBBS profile depicted a fast degradation until 30 min, which is consistent with the aforementioned higher extent in the oxidation of the organic matter during the initial stages, its concentration stagnated after 30 min. At this time, the oxidation of other compounds than NBBS appeared to be favored since after 3
132 Results and Discussion
h of treatment, and when nearly 92% of COD was removed, 47.6% of NBBS (613.4 µg L-1) remained in the solution.
Despite the high efficiency of the electrochemical oxidation in the degradation of major contaminants shown in this work, as well as in former literature (Anglada et al. 2009; Angela Anglada et al. 2010; Anglada et al. 2011; Alonso 2013) up to our knowledge data dealing with changes in PCDD/Fs concentration during the electrochemical oxidation of landfill leachates has not been previously reported. In fact, not many works have been found regarding the monitoring of PCDD/Fs during the advanced oxidation treatment of landfill leachates. On this way, Vollmuth et al. (1994) and Vollmuth & Niessner (1995) observed an increase in the concentration of some PCDD/Fs congeners during the treatment of landfill seepage waters by means of UV/O3. Next, results dealing with the monitoring of PCDD/Fs during the electrocoxidation of landfill leachates are introduced. PCDD/Fs were analyzed at different times: 5 and 30 min, corresponding with a fast initial COD degradation, and 3 h, when the removal of COD started to stagnate. The homologue profiles of total PCDD/Fs in raw leachate S2 and in its oxidized samples are shown in Figure 3.35.
4000 S₂ 3000 EOX 5 min EOX 30 min 2000 EOX 180 min ) 1 -
L 1000 pg ( 0 100
80 Concentration
60
40
20
0 TCDD PeCDD HxCDDHpCDD OCDD TCDF PeCDF HxCDF HpCDF OCDF
Figure 3.35. Homologue profile of total PCDD/Fs in raw leachate S2 and in its electrochemically oxidized samples.
133 Chapter 3
The homologue profiles displayed in Figure 3.35 reported a general decreasing trend in PCDD/Fs concentration as a result of the electrochemical treatment. After 3h of treatment, the total concentration of PCDD/Fs was reduced by 71.5%. At this time, the homologue profile was still dominated by PCDDs, which represented the 93% of the total PCDD/Fs concentration due to the high amount of OCDD. Regarding PCDFs, whose concentration explained the remaining 7% of the total PCDD/Fs, it could be observed an increase in the concentration of lower chlorinated homologues (TCDF, PeCDF and HxCDF) in the course of the treatment, possibly as a result of the dechlorination of the higher chlorinated ones during the process (Holt et al. 2012). On the other hand, the congener profiles of 2,3,7,8-PCDD/Fs, found in the raw leachate sample S2 as well as in its electrochemically treated samples, are shown in Figure 3.36.
4000 S₂ 3000 EOX 5 min EOX 30 min 2000 EOX 180 min ) 1 - 1000 L pg ( 0 40
30 Concentration
20
10
0
Figure 3.36. Congener profile of 2,3,7,8-PCDD/Fs in raw leachate S2 and in its electrochemically oxidized samples.
134 Results and Discussion
Concerning the more toxic congeners, i.e. 2,3,7,8-PCDD/Fs (Figure 3.36), it could be observed a concurrent decrease in their concentration over the treatment period. After 3 h of electrochemical oxidation, the total concentration of 2,3,7,8-PCDD/Fs decreased by 71%. At this time, the congener profile remained dominated by 2,3,7,8-PCDDs, which justified the 97% of the 2,3,7,8-PCDD/Fs total concentration due to the high quantity of the congener OCDD. In the case of 2,3,7,8-PCDFs, the most abundant congeners, 1,2,3,4,6,7,8-PCDF and OCDF, showed the same trend of decreasing in their concentration observed for 2,3,7,8-PCDDs. The remaining PCDD/Fs congeners were present at great lower concentrations and did not depict a clear trend with time. In the case of 2,3,7,8-TCDF, as it was depicted by total PCDFs, its concentration depicted an increase during the process, possibly due to the dechlorination of the higher chlorinated congeners. When the concentration of 2,3,7,8-PCDD/Fs was expressed as TEQ, after 3 h of treatment, the TEQ of the raw leachate sample declined by 58% (from 13.1 to 5.5 pg I-TEQ L-1) using the I- TEFs, and by 56.7% (from 10 to 4.4 pg WHO-TEQ L-1) when the WHO-TEFs were applied.
As a summary, results showed that after 3 h of treatment and when 92% of COD was depleted, and therefore was under the discharge limits, the total concentration of PCDD/Fs in the leachate sample declined by 71.5% and the I- TEQ index decreased by 58% (56.7% for WHO-TEQ). Several AOPs have been considered in the development of an efficient and economically feasible approach for destroying PCDD/Fs such as photolysis (Dung & O’Keefe 1994), photocatalysis (Choi et al. 2000) and ozonolysis (Vollmuth & Niessner 1997). In particular, photolysis, which is probably the major degrading mechanism for the removal of chlorinated dioxins from the total environment (Kieatiwong et al. 1990), has been one of the most applied AOPs for PCDD/Fs remediation (Friesen et al. 1990; Friesen et al. 1996). Accordingly, this work reported for the first time the effectiveness of the electrochemical oxidation to degrade PCDD/Fs when it was applied to the treatment of landfill leachates.
3.3.4. Fenton oxidation of landfill leachates
2+ Fenton oxidation of leachate sample S1 was carried out using a H2O2/Fe 2+/ mass ratio equal to 5.86 ((H2O2/COD = 1.7)/(Fe COD = 0.29)), which was optimized in previous works of the research group where this thesis has been performed, in order to remove COD and color from landfill leachates (Primo et
135 Chapter 3
al. 2008; Primo & Rueda 2008). The concentrations of the major parameters,
COD and TOC, along with the residual H2O2, which are shown in Figure 3.37, were selected to assess the effectiveness of the degradation process.
COD TOC H2O2 1.0
0.8
0.6
0.4
0.2 Normalized concentration
0.0 0 50 100 150
time (min)
Figure 3.37. Changes in normalized concentration of COD, TOC and H2O2 during the
Fenton oxidation of leachate S1.
The COD and TOC concentration profiles from Figure 3.37 depicted three different stages in the course of the treatment. In a first stage up to 15 min, COD and TOC concentrations displayed faster degradation rates, consistent with a rapid H2O2 consumption and therefore with an initial fast OH formation. After 15 min, in a second step, COD and TOC started to degrade moderately coinciding with a slower decay of H2O2 and consequent decrease in OH production. With the progress of the Fenton reaction, the amount of available 2+ Fe in the reaction media decreases and limits the consumption of H2O2 to 2+ 3+ produce OH , since Fe regeneration from Fe (r6) is slower than the reaction of H2O2 to produce OH (r5) (Bautista et al. 2008). Besides, carboxylic acids, frequently found as intermediates during the Fenton oxidation of organic compounds, scavenge iron from the reaction medium through the formation of complexes (Nakagawa & Yamaguchi 2012). After 1 h of treatment and coinciding with the complete depletion of H2O2, the concentration of COD and TOC reached a plateau, showing maximum degradation percentages of 75% and 70%, respectively. At this time, the COD (almost 600 mg L-1 ) remaining in solution was above the discharge limits (160 mg L-1). However, an increase in the biodegradability of the leachate samples, BOD5/COD, up to 0.6 was
136 Results and Discussion
reported by Alonso (2013) during the Fenton treatment of similar leachates to those studied in this work and under the same operating conditions. According to Lopez et al. (2004), BOD5/COD ratios greater than 0.4 are considered compatible with the biological treatment of the sample. Therefore, biological degradation can be integrated with the Fenton oxidation as a post-treatment stage after the enhancement of the biodegradability by this process. As in the case of the electrochemical oxidation, section 3.3.3, the concentration of the two main organic contaminants identified in the raw leachate, NBBS and BPA, was checked during the Fenton oxidation of leachate
S1 (Figure 3.38).
NBBS BPA 1500 ) 1 -
L 1000 100 80 60 40 Concentration (µg (µg Concentration 20 0 0 50 100 150 200 time (min) Figure 3.38. Changes in NBBS and BPA concentration during the Fenton oxidation of
leachate S1.
According to Figure 3.38, NBBS and BPA were almost completely degraded within the first 5 min of treatment, showing removal percentages of 99% and 96%, respectively. Afterwards, despite the complete degradation of NBBS and BPA was not achieved within 3 h of treatment, these compounds remained in the solution at very low concentrations, i.e. 8.5 µg L-1 for NBBS and 42 µg L-1 for BPA. As noted for the electrochemical oxidation, the changes in PCDD/Fs concentration during the Fenton oxidation of landfill leachates have not been reported before. Next, PCDD/Fs were analyzed at different Fenton oxidation times: 5 min and 30 min, corresponding with a fast initial COD degradation, and 3 h, when the removal of COD remained invariable. The profiles of total
137 Chapter 3
PCDD/Fs in the untreated leachate sample S1 and in its Fenton oxidized samples are compared in Figure 3.39.
5000 S₁ 4000 Fenton 5 min Fenton 30 min 3000 Fenton 180 min
2000 ) 1 -
L 1000 pg ( 0 250
200 Concentration 150
100
50
0 TCDD PeCDD HxCDDHpCDD OCDD TCDF PeCDF HxCDF HpCDF OCDF
Figure 3.39 Homologue profile of total PCDD/Fs in raw leachate S1 and in its Fenton oxidized samples.
As can be observed in Figure 3.39, several homologue groups depicted an increase in their concentration during the Fenton oxidation. Accordingly, after 3 h of treatment the total concentration of PCDD/Fs increased by 29% and the homologue profile continued being dominated by PCDDs, which represented the 97% of the total PCDD/Fs concentration as a result of the high concentration of the congener OCDD. The higher concentrations of HpCDD and OCDD suggest that their precursors were also present in higher amounts in the reaction medium and from the very beginning, as depicted the faster increase of PCDD/Fs concentration during the first minutes of treatment. With regard to PCDFs, which accounted for the remaining 3% of total PCDD/Fs, a clear decreasing in the concentration of lower chlorinated homologues was reflected. One possible explanation would be as the result of their chlorination to lead to the higher chlorinated congeners, whose concentration increased with time.
138 Results and Discussion
With respect to the most toxic congeners, 2,3,7,8-PCDD/Fs, their concentrations in the raw leachate sample S1 as well as in its Fenton oxidized samples are shown in Figure 3.40.
5000
S₂S1 4000 Fenton 5 min Fenton 30 min 3000 Fenton 180 min 2000 ) 1 - L 1000 pg ( 0 75
60 Concentration 45
30
15
0
Figure 3.40. Congener profile of 2,3,7,8-PCDD/Fs in raw leachate S1 and in its Fenton oxidized samples.
Data from Figure 3.40 showed a clear trend of increasing concentration for the dominant congeners 1,2,3,4,6,7,8-HpCDD and OCDD, and also for 1,2,3,4,6,7,8-HpCDF and OCDF. For the remaining congeners, which are present at great lower concentrations, a clear pattern could not be observed. The total concentration of 2,3,7,8-PCDD/Fs increased by 37% after 3 h of treatment. At this time, the congener profile was dominated by the family of 2,3,7,8-PCDDs due to the high concentration of high chlorinated congeners, especially OCDD. Besides, after 3 h of Fenton oxidation the TEQ of the raw leachate increased in 12.6% (from 17.2 to 19.37 pg I-TEQ L-1) using the I-TEFs, and by 9.5% (from 13.7 to 15 pg WHO-TEQ L-1) when the WHO-TEFs were applied, since the WHO-TEF for the main congener, OCDD, was lower than its corresponding I-TEF.
139 Chapter 3
In order to confirm these results, an additional Fenton experiment was carried out with a different leachate sample, S3, and after 3 h of treatment the concentrations of PCDD/Fs were determined. The obtained values in the oxidized sample were compared with those resulting from their analysis in the raw leachate in Figure 3.41 (total PCDD/Fs) and Figure 3.42 (2,3,7,8-PCDD/Fs).
10000 S₃ 8000 Fenton 180 min 6000
4000 ) 1 -
L 2000 pg ( 0 100
80 Concentration 60
40
20
0 TCDD PeCDDHxCDDHpCDD OCDD TCDF PeCDF HxCDF HpCDF OCDF
Figure 3.41. Homologue profile of total PCDD/Fs in raw leachate S3 and in its Fenton oxidized sample.
140 Results and Discussion
10000
8000 S₃ Fenton 180 min 6000
4000 ) 1 -
L 2000 pg ( 0 80
60 Concentration 40
20
0
Figure 3.42. Congener profile of 2,3,7,8-PCDD/Fs in raw leachate S3 and in its Fenton
oxidized sample.
A common increasing trend in the concentration of total PCDD/Fs as a result of the Fenton oxidation was depicted in Figure 3.41. The total PCDD/Fs concentration increased by 146% after 3 h. At the same time, the total concentration of 2,3,7,8-PCDD/Fs (Figure 3.42) increased in 147.8% with respect to the untreated sample. As consequence, the TEQ of the sample raised from 13.4 to 30.6 pg I-TEQ L-1 (128.4%) with the I-TEFs, and from 10.7 to 24.1 pg WHO-TEQ L-1 (125.2%) when the WHO-TEFs were used.
On the basis of these results, at times when maximum COD and TOC reductions of 75% and 70% were respectively achieved, the Fenton oxidation of landfill leachate resulted in the formation of PCDD/Fs and in the increase of the TEQ index. After 3 h of Fenton treatment, the total PCDD/Fs concentration in two leachate samples increased by 29-146% and the TEQ was incremented by 12.6-128.4%. These findings were opposite to those depicted in the electrochemical oxidation of landfill leachates, where after 3 h of treatment and when 92% of COD was depleted, the total concentration of PCDD/Fs in the leachate sample and the TEQ index declined by 71.5% and 58%, respectively.
141 Chapter 3
Since leachate samples had the same origin, the main difference between the results may reside in the difference in the operating conditions. Electrochemical oxidation generates OH continuously on the anode surface by the oxidative decomposition of water, whereas in the Fenton process, these radicals are 2+ produced by the mixture of H2O2 with Fe as catalyst under acidic pH. Therefore, in the Fenton oxidation the amount of formed OH is limited by the dose of H2O2 applied. As it was depicted for COD and TOC, a plateau in their concentration profiles was reached when H2O2 was completely consumed. However, the electrochemical oxidation showed a higher capacity to remove the organic content, with nearly 95% of COD and 90% of TOC removal. Accordingly, this technology depicted high effectiveness in the degradation of PCDD/Fs when it was applied to the treatment of landfill leachates under the operating conditions used in this work.
On the other hand, the increase in the concentration of PCDD/Fs in leachates as a result of their Fenton oxidation was in agreement with the findings shown in chapter 3.2 regarding the Fenton treatment of 2-CP. Those results reported the formation of some PCDD/Fs congeners working with low doses of H2O2 (20% of the substoichiometric amount). Under that conditions, although 2-CP was almost completely removed, the percentage of TOC removal achieved was negligible leading to the formation of intermediate products that remained in the reaction medium. Although TOC balance was far from being closed, and therefore a significant amount of reaction products were not identified, the presence of coupling-reaction aromatic byproducts, including PCDD/Fs, was reported. The formation of this type of compounds during the Fenton oxidation of different CPs was also described in previous research works found in the literature (Fukushima & Tatsumi 2001; Poerschmann & Trommler 2009; Munoz et al. 2011; Munoz et al. 2013), as was described in more detail in section 3.2.
Furthermore, as it was briefly summarized in Table 1.5 from chapter 1.3, the formation of PCDD/Fs has also been reported during the treatment of aqueous solutions containing compounds that can act as their precursors, mainly CPs, using other AOPs. On this way, Vollmuth et al. (1994) found that the photolytic treatment of PCP in water led to the formation of several PCDD/Fs congeners. Hong et al. (2000) observed the presence of 1,2,3,4,6,7,8-HpCDD and OCDD as minor transformation products during the photolysis of PCP. Moreover, photolytic and photocatalytic treatments of triclosan in water
142 Results and Discussion
solutions resulted in the formation of some PCDD congeners (Aranami & Readman 2007; Buth et al. 2011; Sankoda et al. 2011). On the other hand, Holt et al. (2012) found that the concentration of 93 PCDD/F congeners in two pesticide formulations increased by 5600% and 3000% after their exposure to natural light.
The presence of several CPs in the leachates studied in this work, although at lower concentrations, ≈ µg L-1, was reported in Table 3.3 from section 3.3.1. In addition, other potential PCDD/Fs precursors such as chlorinated diphenylethers (PClDEs), PCBs, etc, are likely to be present in landfill leachates. As a result of the complexity of the matrix, these kind of compounds were not detected in the qualitative screening (non-target analysis) of organic compounds carried out in the leachate samples. Nevertheless, their presence in leachates from MSW landfills at levels has been reported in literature works where specific analytical methods were applied for their target monitoring. Therefore, the presence of PCDD/Fs precursors in the leachates, although expected in low concentrations, may lead to the formation of PCDD/Fs under the Fenton operating conditions applied in this study. On this way, Vollmuth and Niessner (1995) investigated the degradation of some persistent compounds, among them PCDD/Fs, in landfill seepage waters by UV, O3, and
UV/O3 treatments. Similar to our observations, the degradation of PCDD/Fs could not be detected, whereas on the contrary, an increase in their concentrations was observed. However, the authors concluded that higher O3 and UV doses were needed for the degradation of PCDD/Fs because of the large concentration of organic compounds contained in the seepage waters. As consequence, maybe more stringent Fenton conditions in the treatment of leachates could lead to further destruction of PCDD/Fs, as it was shown during their electrochemical oxidation, where the operating conditions applied depicted high effectiveness in the degradation of PCDD/Fs.
On the contrary, results from chapter 3.1 regarding the electrooxidation of 2-CP reported the formation of PCDD/Fs under similar operating conditions that those applied for the leachate treatment (J = 900 A m-2; Cl- = 2000 mg L-1). Although a high TOC mineralization percentage was achieved after 4 h of treatment (95%), the high concentrations of 2-CP and the presence of concomitant chloride in the reaction medium led to the formation of high amounts of PCDD/Fs. It seems then, that the concentration of PCDD/Fs precursors was determining in their potential formation in the electrochemical
143 Chapter 3
process. However, in the case of the Fenton oxidation, the type of precursor displayed more importance than their concentration, since the Fenton oxidation of high amounts of 2-CP led to PCDD/Fs formation in lower extent that it was depicted for landfill leachates.
144 Results and Discussion
3.4. References
Abad, E., Caiaach, J. & Rivers, J., 1997. PCDD/PCDF from emission sources and ambient air in Northeast Spain. Chemosphere, 35(3), pp.453–463.
Ahmed, F.N. & Lan, C.Q., 2012. Treatment of landfill leachate using membrane bioreactors: A review. Desalination, 287, pp.41–54.
Aissam, H., Penninckx, M.J. & Benlemlih, M., 2007. Reduction of phenolics content and COD in olive oil mill wastewaters by indigenous yeasts and fungi. World Journal of Microbiology and Biotechnology, 23(9), pp.1203–1208.
Alonso, E., 2013. Contribución al tratamiento de lixiviados de vertedero de residuos sólidos urbanos mediante procesos de oxidación avanzada. Universidad de Cantabria.
Altarawneh, M. et al., 2009. Mechanisms for formation, chlorination, dechlorination and destruction of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs). Progress in Energy and Combustion Science, 35(3), pp.245–274.
Anglada, A. et al., 2011. Boron-doped diamond anodic treatment of landfill leachate: evaluation of operating variables and formation of oxidation by- products. Water research, 45(2), pp.828–38.
Anglada, Á., 2010. Electro-oxidation on Boron-doped Diamond Anodes of Ammonia and Organic Pollutants in Landfill Leachate. Universidad de Cantabria.
Anglada, A., Urtiaga, A. & Ortiz, I., 2009. Pilot Scale Performance of the Electro-Oxidation of Landfill Leachate at Boron-Doped Diamond Anodes. Environmental Science & Technology, 43(6), pp.2035–2040.
Anglada, A., Urtiaga, A.M. & Ortiz, I., 2010. Laboratory and pilot plant scale study on the electrochemical oxidation of landfill leachate. Journal of hazardous materials, 181(1-3), pp.729–35.
Aranami, K. & Readman, J.W., 2007. Photolytic degradation of triclosan in freshwater and seawater. Chemosphere, 66(6), pp.1052–6.
Azizi, O. et al., 2011. Mechanism of perchlorate formation on boron-doped diamond film anodes. Environmental science & technology, 45(24), pp.10582– 90.
145 Chapter 3
Bacardit, J. et al., 2007. Effect of Salinity on the Photo-Fenton Process. Industrial & Engineering Chemistry Research, 46(23), pp.7615–7619.
Baderna, D. et al., 2011. A combined approach to investigate the toxicity of an industrial landfill’s leachate: chemical analyses, risk assessment and in vitro assays. Environmental research, 111(4), pp.603–13.
Baker, J.I. & Hites, R.A., 2000. Critical Review Is Combustion the Major Source of Polychlorinated Dibenzo- p -dioxins and Dibenzofurans to the Environment ? A Mass Balance Investigation. , pp.2879–2886.
Bautista, P. et al., 2008. An overview of the application of Fenton oxidation to industrial wastewaters treatment. , 1338(April), pp.1323–1338.
Benitez, F.J. et al., 2000. Contribution of free radicals to chlorophenols decomposition by several advanced oxidation processes. Chemosphere, 41(8), pp.1271–1277.
Bergmann, M.E.H., 2010. Drinking Water Disinfection by In-line Electrolysis: Product and Inorganic By-product Formation. In C. Comninellis & G. Chen, eds. Electrochemistry for the Environment. Springer New York Dordrecht Heidelberg London, pp. 163–204.
Boonnorat, J. et al., 2014. Removals of phenolic compounds and phthalic acid esters in landfill leachate by microbial sludge of two-stage membrane bioreactor. Journal of hazardous materials.
Boudenne, J.-L. & Cerclier, O., 1999. Performance of carbon black-slurry electrodes for 4-chlorophenol oxidation. Water Research, 33(2), pp.494–504.
Briois, C. et al., 2007. Formation of Polychlorinated Dibenzo- p -dioxins and Dibenzofurans from a Mixture of Chlorophenols over Fly Ash: Influence of Water Vapor. Environmental Science & Technology, 41(3), pp.850–856.
Burton, S.A.Q. & Watson-Craik, I.A., 1998. Ammonia and nitrogen fluxes in landfill sites: applicability to sustainable landfilling. Waste Management & Research, 16(1), pp.41–53.
Buth, J.M. et al., 2011. Removal and formation of chlorinated triclosan derivatives in wastewater treatment plants using chlorine and UV disinfection. Chemosphere, 84(9), pp.1238–43.
146 Results and Discussion
Cabeza, A., Urtiaga, A.M. & Ortiz, I., 2007. Electrochemical Treatment of Landfill Leachates Using a Boron-Doped Diamond Anode. Industrial & Engineering Chemistry Research, 46(5), pp.1439–1446.
Cañizares, P. et al., 2005. Electrochemical oxidation of phenolic wastes with boron-doped diamond anodes. Water research, 39(12), pp.2687–703.
Casanovas, J. et al., 1994. PCDF and PCDD levels in different types of environmental samples in spain. Fresenius’ Journal of Analytical Chemistry, 348(1-2), pp.167–170.
Castro-Jiménez, J. et al., 2010. Atmospheric occurrence and deposition of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) in the open Mediterranean Sea. Environmental science & technology, 44(14), pp.5456–63.
Chamarro, E., 2001. Use of fenton reagent to improve organic chemical biodegradability. Water Research, 35(4), pp.1047–1051.
Chiang, L.-C., Chang, J.-E. & Wen, T.-C., 1995. Indirect oxidation effect in electrochemical oxidation treatment of landfill leachate. Water Research, 29(2), pp.671–678. [Accessed May 10, 2014].
Choi, K.-I. & Lee, D.-H., 2006. PCDD/DF in leachates from Korean MSW landfills. Chemosphere, 63(8), pp.1353–60.
Choi, W. et al., 2000. Photocatalytic Degradation of Polychlorinated Dibenzo- p -dioxins on TiO 2 Film under UV or Solar Light Irradiation. Environmental Science & Technology, 34(22), pp.4810–4815. A
Clarke, B. et al., 2008. Dioxin-like compounds in Australian sewage sludge-- review and national survey. Chemosphere, 72(8), pp.1215–28.
Cleverly, D. et al., 2007. A general indication of the Contemporary background levels of PCDDs, PCDFs, and coplanar PCBs in the ambient air over rural and remote areas of the United States. Environmental science & technology, 41(5), pp.1537–44.
Comninellis, C. & Nerini, A., 1995. Anodic oxidation of phenol in the presence of NaCl for wastewater treatment. Journal of Applied Electrochemistry, 25(1), pp.23–28.
147 Chapter 3
Correa, O. et al., 2006. Depositional flux of polychlorinated dibenzo-p- dioxins and polychlorinated dibenzofurans in an urban setting. Chemosphere, 64(9), pp.1550–61.
Cossu, R. et al., 1998. Electrochemical Treatment of Landfill Leachate: Oxidation at Ti/PbO 2 and Ti/SnO 2 Anodes. Environmental Science & Technology, 32(22), pp.3570–3573.
Dai, J. et al., 2007. PCDD/F, PAH and heavy metals in the sewage sludge from six wastewater treatment plants in Beijing, China. Chemosphere, 66(2), pp.353–61. [Accessed May 7, 2014].
Deng, Y., Rosario-Muniz, E. & Ma, X., 2012. Effects of inorganic anions on Fenton oxidation of organic species in landfill leachate. Waste management & research : the journal of the International Solid Wastes and Public Cleansing Association, ISWA, 30(1), pp.12–9.
Detomaso, A. et al., 2003. Practical Applications of the Fenton Reaction to the Removal of Chlorinated Aromatic Pollutants: Oxidative Degradation of 2,4- Dichlorophenol. In Environmental Science and Pollution Research. pp. 379–384.
Dirany, A. et al., 2012. Electrochemical treatment of the antibiotic sulfachloropyridazine: kinetics, reaction pathways, and toxicity evolution. Environmental science & technology, 46(7), pp.4074–82.
Dorival-García, N. et al., 2012. Improved sample treatment for the determination of bisphenol A and its chlorinated derivatives in sewage sludge samples by pressurized liquid extraction and liquid chromatography-tandem mass spectrometry. Talanta, 101, pp.1–10.
Duesterberg, C.K. & Waite, T.D., 2007. Kinetic Modeling of the Oxidation of p -Hydroxybenzoic Acid by Fenton’s Reagent: Implications of the Role of Quinones in the Redox Cycling of Iron. Environmental Science & Technology, 41(11), pp.4103–4110.
Duesterberg, C.K. & Waite, T.D., 2006. Process Optimization of Fenton Oxidation Using Kinetic Modeling. Environmental Science & Technology, 40(13), pp.4189–4195.
Dung, M.H. & O’Keefe, P.W., 1994. Comparative rates of photolysis of polychlorinated dibenzofurans in organic solvents and in aqueous solutions. Environmental science & technology, 28(4), pp.549–54.
148 Results and Discussion
Dyke, P.H. et al., 1997. A review of dioxin releases to land and water in the UK., Available at: http://www.ncbi.nlm.nih.gov/pubmed/9447741.
Eggen, T. et al., 2013. Uptake and translocation of organophosphates and other emerging contaminants in food and forage crops. Environmental science and pollution research international, 20(7), pp.4520–31.
Eggen, T., Moeder, M. & Arukwe, A., 2010. Municipal landfill leachates: a significant source for new and emerging pollutants. The Science of the total environment, 408(21), pp.5147–57.
El-Ashtoukhy, E.S.Z. et al., 2013. Treatment of petrochemical wastewater containing phenolic compounds by electrocoagulation using a fixed bed electrochemical reactor. International Journal of Electrochemical Science, 8(1), pp.1534–1550.
Eljarrat, E., Caixach, J. & Rivera, J., 2003. A comparison of TEQ contributions from PCDDs, PCDFs and dioxin-like PCBs in sewage sludges from Catalonia, Spain. Chemosphere, 51(7), pp.595–601.
Eljarrat, E., Caixach, J. & Rivera, J., 1999. Decline in PCDD and PCDF Levels in Sewage Sludges from Catalonia (Spain). Environmental Science & Technology, 33(15), pp.2493–2498.
Fatta-Kassinos, D. et al., 2011. The risks associated with wastewater reuse and xenobiotics in the agroecological environment. The Science of the total environment, 409(19), pp.3555–63.
Fiedler, H., 2003. Dioxins and Furans (PCDD/PCDF). ChemInform, 34(26), pp.11–13.
Friesen, K.J., Foga, M.M. & Loewen, M.D., 1996. Aquatic Photodegradation of Polychlorinated Dibenzofurans: Rates and Photoproduct Analysis. Environmental Science & Technology, 30(8), pp.2504–2510.
Friesen, K.J., Muir, D.C.G. & Webster, G.R.B., 1990. Evidence of sensitized photolysis of polychlorinated dibenzo-p-dioxins in natural waters under sunlight conditions. Environmental Science & Technology, 24(11), pp.1739–1744.
Fuentes, M.J. et al., 2007. Multivariant statistical analysis of PCDD/FS in sewage sludges from different areas of the Valencian Community (Spain). Chemosphere, 67(7), pp.1423–33.
149 Chapter 3
Fukushima, M. & Tatsumi, K., 2001. Degradation Pathways of Pentachlorophenol by Photo-Fenton Systems in the Presence of Iron(III), Humic Acid, and Hydrogen Peroxide. Environmental Science & Technology, 35(9), pp.1771–1778.
Gogate, P.R. & Pandit, A.B., 2004. A review of imperative technologies for wastewater treatment I: oxidation technologies at ambient conditions. Advances in Environmental Research, 8(3-4), pp.501–551.
Govindaraj, M. et al., 2013. Electrochemical oxidation of bisphenol-A from aqueous solution using graphite electrodes. Environmental Technology, 34(4), pp.503–511.
Gworek, B. et al., 2013. Influence of a municipal waste landfill on the spatial distribution of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDDs/Fs) in the natural environment. Chemosphere, 92(7), pp.753–9.
Hagenmaier, H. & Brunner, H., 1987. Isomerspecific analysis of pentachlorophenol and sodium pentachlorophenate for 2,3,7,8-substituted PCDD and PCDF at sub-ppb levels. Chemosphere, 16(8-9), pp.1759–1764.
Hagenmaier, H., Lindig, C. & She, J., 1994. Correlation of environmental occurrence of polychlorinated dibenzo-p-dioxins and dibenzofurans with possible sources. Chemosphere, 29(9-11), pp.2163–2174.
Ham, S.-Y., Kim, Y.-J. & Lee, D.-H., 2008. Leaching characteristics of PCDDs/DFs and dioxin-like PCBs from landfills containing municipal solid waste and incineration residues. Chemosphere, 70(9), pp.1685–93.
Holt, E. et al., 2008. Assessing dioxin precursors in pesticide formulations and environmental samples as a source of octachlorodibenzo-p-dioxin in soil and sediment. Environmental science & technology, 42(5), pp.1472–8.
Holt, E. et al., 2012. Formation of dioxins during exposure of pesticide formulations to sunlight. Chemosphere, 88(3), pp.364–70.
Hong, J. et al., 2003. Identification of electrical degradation products of 4- chlorophenol in water. Analytical sciences : the international journal of the Japan Society for Analytical Chemistry, 19(4), pp.537–42.
Hong, J. et al., 2000. Identification of photolytical transformation products of pentachlorophenol in water. Analytical Sciences, 16(6), pp.621–626.
150 Results and Discussion
Horstmann, M., McLachlan, M.S. & Reissinger, M., 1993. Investigations of the origin of PCDD/F in municipal sewage sludge. Chemosphere, 27(1-3), pp.113–120.
Huang, C.P. & Chu, C., 2012. Indirect Electrochemical Oxidation of Chlorophenols in Dilute Aqueous Solutions. Journal of Environmental Engineering, 138(3), pp.375–385.
Hutzinger, O. & Fiedler, H., 1993. From source to exposure: Some open questions. Chemosphere, 27(1-3), pp.121–129.
Iniesta, J., 2001. Electrochemical oxidation of phenol at boron-doped diamond electrode. Electrochimica Acta, 46(23), pp.3573–3578.
Kapałka, A. et al., 2010. Direct and mediated electrochemical oxidation of ammonia on boron-doped diamond electrode. Electrochemistry Communications, 12(12), pp.1714–1717. [Accessed April 28, 2014].
Kieatiwong, S. et al., 1990. Photolysis of chlorinated dioxins in organic solvents and on soils. Environmental Science & Technology, 24(10), pp.1575– 1580.
Kjeldsen, P. et al., 2002. Present and Long-Term Composition of MSW Landfill Leachate: A Review. Critical Reviews in Environmental Science and Technology, 32(4), pp.297–336.
Klimm, C. et al., 1998. Formation of Octa- and heptachlorodibenzo-p- dioxins during semi anaerobic digestion of sewage sludge. Chemosphere, 37(9- 12), pp.2003–2011.
Koch, M., Knoth, W. & Rotard, W., 2001. Source identification of PCDD/Fs in a sewage treatment plant of a German village. Chemosphere, 43(4-7), pp.737– 741.
De la Cruz, N. et al., 2012. Degradation of 32 emergent contaminants by UV and neutral photo-fenton in domestic wastewater effluent previously treated by activated sludge. Water research, 46(6), pp.1947–57.
De la Torre, A. et al., 2011. Sources and behaviour of polybrominated diphenyl ethers (PBDEs), polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) in Spanish sewage sludge. Waste management (New York, N.Y.), 31(6), pp.1277–84.
151 Chapter 3
Liu, W. et al., 2004. Formation of PCDD/Fs and PCBs in the process of production of 1,4-dichlorobenzene. Chemosphere, 57(10), pp.1317–23.
Lohmann, R. & Jones, K.C., 1998. Dioxins and furans in air and deposition: a review of levels, behaviour and processes. The Science of the total environment, 219(1), pp.53–81.
Lopez, A. et al., 2004. Fenton’s pre-treatment of mature landfill leachate. Chemosphere, 54(7), pp.1005–1010.
Lopez, A. et al., 2005. Temperature activated degradation (mineralization) of 4-chloro-3-methyl phenol by Fenton’s reagent. Chemosphere, 59(3), pp.397– 403.
Lu, M.-C. et al., 2005. Effect of chloride ions on the oxidation of aniline by Fenton’s reagent. Journal of environmental management, 75(2), pp.177–82.
Mascia, M. et al., 2010. Electrochemical treatment of phenolic waters in presence of chloride with boron-doped diamond (BDD) anodes: experimental study and mathematical model. Journal of hazardous materials, 174(1-3), pp.314–22.
Meffe, R. & de Bustamante, I., 2014. Emerging organic contaminants in surface water and groundwater: A first overview of the situation in Italy. The Science of the total environment, 481C, pp.280–295.
Méndez-Arriaga, F., Esplugas, S. & Giménez, J., 2010. Degradation of the emerging contaminant ibuprofen in water by photo-Fenton. Water research, 44(2), pp.589–95.
Micó, M.M. et al., 2013. Enhancement of pesticide photo-Fenton oxidation at high salinities. Applied Catalysis B: Environmental, 132-133, pp.162–169.
Morrison, R.T. & Boyd, R.N., 1998. Química Orgánica fifth., Pearson Educación.
Munoz, M. et al., 2011. Assessment of the generation of chlorinated byproducts upon Fenton-like oxidation of chlorophenols at different conditions. Journal of hazardous materials, 190(1-3), pp.993–1000.
Munoz, M. et al., 2013. Chlorophenols breakdown by a sequential hydrodechlorination-oxidation treatment with a magnetic Pd-Fe/γ-Al2O3 catalyst. Water research, 47(9), pp.3070–80.
152 Results and Discussion
Munter, R. et al., 2006. Cost Effectiveness of Ozonation and AOPs for Aromatic Compound Removal from Water: A Preliminary Study. Ozone: Science & Engineering, 28(5), pp.287–293.
Musson, S.E. & Townsend, T.G., 2009. Pharmaceutical compound content of municipal solid waste. Journal of hazardous materials, 162(2-3), pp.730–5.
Nakagawa, H. & Yamaguchi, E., 2012. Influence of oxalic acid formed on the degradation of phenol by Fenton reagent. Chemosphere, 88(2), pp.183–7.
Neyens, E. & Baeyens, J., 2003. A review of classic Fenton’s peroxidation as an advanced oxidation technique. Journal of Hazardous Materials, 98(1-3), pp.33–50.
Öberg, L.G. & Rappe, C., 1992. Biochemical formation of PCDD/Fs from chlorophenols. Chemosphere, 25(1-2), pp.49–52.
Oman, C.B. & Junestedt, C., 2008. Chemical characterization of landfill leachates--400 parameters and compounds. Waste management (New York, N.Y.), 28(10), pp.1876–91.
Paíga, P. et al., 2013. Pilot monitoring study of ibuprofen in surface waters of north of Portugal. Environmental science and pollution research international, 20(4), pp.2410–20.
Pan, W. et al., 2013. New insight into the formation mechanism of PCDD/Fs from 2-chlorophenol precursor. Environmental science & technology, 47(15), pp.8489–98.
Panizza, M. et al., 2001. Anodic oxidation of 2-naphthol at boron-doped diamond electrodes. Journal of Electroanalytical Chemistry, 507(1-2), pp.206– 214.
Panizza, M. & Cerisola, G., 2005. Application of diamond electrodes to electrochemical processes. Electrochimica Acta, 51(2), pp.191–199.
Pera-Titus, M. et al., 2004. Degradation of chlorophenols by means of advanced oxidation processes: a general review. Applied Catalysis B: Environmental, 47(4), pp.219–256.
Pérez, G. et al., 2012. Kinetic study of the simultaneous electrochemical removal of aqueous nitrogen compounds using BDD electrodes. Chemical Engineering Journal, 197, pp.475–482..
153 Chapter 3
Petrovic, M., 2003. Analysis and removal of emerging contaminants in wastewater and drinking water. TrAC Trends in Analytical Chemistry, 22(10), pp.685–696.
Pignatello, J.J., 1992. Dark and photoassisted iron(3+)-catalyzed degradation of chlorophenoxy herbicides by hydrogen peroxide. Environmental Science & Technology, 26(5), pp.944–951.
Pignatello, J.J., Oliveros, E. & MacKay, A., 2006. Advanced Oxidation Processes for Organic Contaminant Destruction Based on the Fenton Reaction and Related Chemistry. Critical Reviews in Environmental Science and Technology, 36(1), pp.1–84.
Poerschmann, J. et al., 2009. Formation of chlorinated biphenyls, diphenyl ethers and benzofurans as a result of Fenton-driven oxidation of 2- chlorophenol. Chemosphere, 75(6), pp.772–80.
Poerschmann, J. & Trommler, U., 2009. Pathways of advanced oxidation of phenol by Fenton’s reagent--identification of oxidative coupling intermediates by extractive acetylation. Journal of chromatography. A, 1216(29), pp.5570–9.
Primo, O., Rivero, M.J. & Ortiz, I., 2008. Photo-Fenton process as an efficient alternative to the treatment of landfill leachates. Journal of hazardous materials, 153(1-2), pp.834–42.
Primo, O. & Rueda, A., 2008. An Integrated Process , Fenton Reaction - Ultrafiltration , for the Treatment of Landfill Leachate : Pilot Plant Operation and Analysis. , (3), pp.946–952.
Pujadas, E. et al., 2001. Application of the new C18 speedisksTM to the analysis of polychlorinated dibenzo-p-dioxins and dibenzofurans in water and effluent samples. Chemosphere, 43(4-7), pp.449–454.
Rappe, C., 1994. Dioxin, patterns and source identification. Fresenius’ Journal of Analytical Chemistry, 348(1-2), pp.63–75.
Roell, D. & Baniahmad, A., 2011. The natural compounds atraric acid and N- butylbenzene-sulfonamide as antagonists of the human androgen receptor and inhibitors of prostate cancer cell growth. Molecular and cellular endocrinology, 332(1-2), pp.1–8.
Ryu, J.-Y. et al., 2005. CuCl2-catalyzed PCDD/F formation and congener patterns from phenols. Chemosphere, 61(9), pp.1312–26.
154 Results and Discussion
Sankoda, K. et al., 2011. Identification of triclosan intermediates produced by oxidative degradation using TiO2 in pure water and their endocrine disrupting activities. Bulletin of environmental contamination and toxicology, 86(5), pp.470–5.
Sedlak, D.L. & Andren, A.W., 1991. Oxidation of chlorobenzene with Fenton’s reagent. Environmental Science & Technology, 25(4), pp.777–782.
Sirés, I. et al., 2014. Electrochemical advanced oxidation processes: today and tomorrow. A review. Environmental science and pollution research international.
Stevens, J., Green, N.J.. & Jones, K.C., 2001. Survey of PCDD/Fs and non- ortho PCBs in UK sewage sludges. Chemosphere, 44(6), pp.1455–1462.
Stuart, M. et al., 2012. Review of risk from potential emerging contaminants in UK groundwater. The Science of the total environment, 416, pp.1–21.
Sundqvist, K.L. et al., 2009. Congener fingerprints of tetra- through octa- chlorinated dibenzo-p-dioxins and dibenzofurans in Baltic surface sediments and their relations to potential sources. Chemosphere, 77(5), pp.612–20.
Sundqvist, K.L. et al., 2010. PCDD/F source apportionment in the Baltic Sea using positive matrix factorization. Environmental science & technology, 44(5), pp.1690–7.
Terashima, C. et al., 2002. Electrochemical Oxidation of Chlorophenols at a Boron-Doped Diamond Electrode and Their Determination by High-Performance Liquid Chromatography with Amperometric Detection. Analytical Chemistry, 74(4), pp.895–902.
Tysklind, M. et al., 1993. Atmospheric Transport and Transformation of Polychlorinated Dibenzo-pdioxins and Dibenzofurans. , (12), pp.2190–2197.
Umar, M. et al., 2013. Application of ozone for the removal of bisphenol A from water and wastewater--a review. Chemosphere, 90(8), pp.2197–207.
Urase, T. & Miyashita, K., 2003. Factors affecting the concentration of bisphenol A in leachates from solid waste disposal sites and its fate in treatment processes. Journal of Material Cycles and Waste Management, 5(1), pp.77–82.
155 Chapter 3
Vollmuth, S. & Niessner, R., 1995. Degradation of PCDD, PCDF, PAH, PCB and chlorinated phenols during the destruction-treatment of landfill seepage water in laboratory model reactor (UV, Ozone, and UV/Ozone). Chemosphere, 30(12), pp.2317–2331.
Vollmuth, S. & Niessner, R., 1997. Degradation of polychlorinated dibenzo‐p‐dioxins and polychlorinated dibenzofurans during the UV/ozone treatment of pentachlorophenol‐containing water. Toxicological & Environmental Chemistry, 61(1-4), pp.27–41.
Vollmuth, S., Zajc, a & Niessner, R., 1994. Formation of Polychlorinated Dibenzo-p-dioxins and Polychlorinated Dibenzofurans during the Photolysis of Pentachlorophenol-Containing Water. Environmental science & technology, 28(6), pp.1145–9.
Wang, H. & Wang, J., 2009. Comparative study on electrochemical degradation of 2,4-dichlorophenol by different Pd/C gas-diffusion cathodes. Applied Catalysis B: Environmental, 89(1-2), pp.111–117.
Weber, R. et al., 2008. Dioxin- and POP-contaminated sites--contemporary and future relevance and challenges: overview on background, aims and scope of the series. Environmental science and pollution research international, 15(5), pp.363–93.
Weber, R., 2007. Relevance of PCDD/PCDF formation for the evaluation of POPs destruction technologies--review on current status and assessment gaps. Chemosphere, 67(9), pp.S109–17.
Weber, R. et al., 2011. Review Article: Persistent organic pollutants and landfills - a review of past experiences and future challenges. Waste management & research : the journal of the International Solid Wastes and Public Cleansing Association, ISWA, 29(1), pp.107–21.
Weber, R. & Hagenmaier, H., 1999. Mechanism of the formation of polychlorinated dibenzo-p-dioxins and dibenzofurans from chlorophenols in gas phase reactions. Chemosphere, 38(3), pp.529–549.
Xu, S. et al., 2011. Leaching behaviour of bisphenol A from municipal solid waste under landfill environment. Environmental Technology, 32(11), pp.1269– 1277.
156 Results and Discussion
Xu, X.H. et al., 2003. 2-chlorophenol oxidation kinetic by photo-assisted Fenton process. Journal of Environmental Sciences, 15(4), pp.475–481.
Zazo, J. a. et al., 2011. Intensification of the Fenton Process by Increasing the Temperature. Industrial & Engineering Chemistry Research, 50(2), pp.866– 870.
Zazo, J.A. et al., 2005. Chemical Pathway and Kinetics of Phenol Oxidation by Fenton’s Reagent. Environmental Science & Technology, 39(23), pp.9295– 9302.
157
Conclusions
This thesis focuses on the study of intermediate products, paying special attention to the potential formation of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) during the application of advanced oxidation processes (AOPs) to the treatment of aqueous samples containing chlorinated organic compounds. Particularly, the electrochemical oxidation and Fenton treatment of both model solutions containing 2-chlorophenol (2-CP), a known precursor of the formation of PCDD/Fs, and leachates from a municipal solid waste (MSW) landfill have been assessed. The main conclusions derived from this work are reported below.
On the electrochemical oxidation of 2-chlorophenol
In chapter 3.1, the electrochemical degradation of 2-CP solutions (15.56 mM) on BDD anodes was carried out under galvanostatic conditions (J = 900 A m-2) studying the influence of two widely used electrolytes, NaCl (56.3 mM) and
Na2SO4 (21.1 mM). 2-CP was completely removed with both electrolytes, but faster rates were observed in NaCl pointing out to the positive contribution of indirect oxidation by electrogenerated active chlorine at the anode surface. No influence of the electrolyte type on the degradation of either TOC or COD was observed. TOC mineralization required higher times than did the reduction of both COD and 2- CP, indicating the formation of oxidation intermediate products that need longer time to be mineralized.
The electrochemical degradation of 2-CP led to the formation of hydroxylation and chlorination products with predominantly ortho/para substitutions. Chlorophenols (CPs) and related compounds were detected only with NaCl, whereas hydroxylated aromatic intermediates such as hydroquinone, catechol and benzoquinone were formed using both electrolytes. Finally, the cleavage of the aromatic ring led to the formation of several aliphatic carboxylic acids, mainly formic and oxalic acids.
After 4 h of treatment, the TOC balance neared 100% when Na2SO4 was employed as electrolyte, whereas a difference between measured TOC (5 mM) and theoretical TOC (1 mM) was reported with NaCl. This difference, together with the unbalanced chlorine measurements, could be explained by the presence of minor chlorinated organic compounds.
161 Conclusions
At times when almost complete mineralization was achieved (4 h), the use of NaCl resulted in a total PCDD/Fs concentration in the oxidized sample of 391.05 ng L-1, that is 2.68 × 104 times higher than in the untreated 2-CP sample. Besides, although the total 2,3,7,8-PCDD/Fs concentration only accounted for 1.6% of the total PCDD/Fs, the TEQ level increased from values close to zero for the untreated 2-CP solution to 220 pg I-TEQ L-1 (243 pg WHO-TEQ L-1), values several times higher than the maximum contaminant level, 30 pg L-1 of 2,3,7,8- TCDD, established by the U.S. EPA for water ingestion. Regarding the use of
Na2SO4 as supporting electrolyte, the increase in total PCDD/Fs concentration was 134 times lower than with NaCl and there was not significant 2,3,7,8- PCDD/Fs formation.
The obtained results highlight the importance of considering the concomitant presence of chloride and organic matter in the reaction medium, especially when PCDD/Fs precursors are initially present or may be formed during the treatment. NaCl, which exerted a positive influence in the removal of 2-CP, showed a strong propensity to form polychlorinated derivatives, including PCDD/Fs. On this way, these findings emphasize the importance of the assessment of transformation products beyond the analysis of conventional physicochemical parameters in extending the understanding of the overall efficiency of AOPs.
On the Fenton oxidation of 2-chlorophenol
Fenton oxidation of 2-CP solutions (15.56 mM) was assessed at different operating conditions in chapter 3.2. H2O2 dose corresponding to 100% (202.28 mM) and 20% (40.44 mM) of the theoretical stoichiometric amount to oxidize 2-CP was used. Two different temperatures, 20ºC and 70ºC, were evaluated, and the presence of chloride (56.3 mM) in the reaction medium was tested.
Although the oxidation of 2-CP was effective for both concentrations (20% and 100% of the stoichiometric H2O2 dose), the experiments with 100% H2O2 resulted in faster degradation of 2-CP and greater COD and TOC removals were achieved due to the higher OH availability. Nevertheless, mineralization was far from complete at both H2O2 doses due to the low concentration of H2O2 remaining in solution and to the presence of organic products (such as oxalic acid) that could deactivate iron through the formation of complexes. When using H2O2 at 20%, the main reaction product formed was 2-
162
Conclusions
chlorobenzoquinone along with final oxidation organic acids such as oxalic, formic and acetic acids, whereas with 100% H2O2, the detected byproducts were mainly aliphatic organic acids (principally oxalic and formic acids).
The increase of working temperature from 20ºC to 70ºC resulted in faster conversions of H2O2 into OH , and therefore enhanced the percentage of mineralization, which was more pronounced for stoichiometric amounts of
H2O2. However, with 20% H2O2, the effect of temperature aimed to favor the kinetics of oxidation reactions for the aromatic derivatives, resulting in their higher conversion into organic acids more than increasing the degree of mineralization.
The effect of chloride ions (56.3 mM ) was analyzed using two sets of experiments differently favored in terms of TOC removal: 100% H2O2 at 70ºC and 20% H2O2 at 20ºC. The presence of chloride in the reaction medium with
20% H2O2 at 20ºC did not affect either 2-CP degradation or organic content removal. In the case of 100% H2O2 and 70ºC, chloride did not affect the oxidation of 2-CP but increased TOC mineralization (11%) and COD degradation (5%). This fact can be attributed to the formation of chlorine radicals (Cl) from the reaction between OH and chloride, which contribute to the degradation of the organic matter.
In general terms, H2O2 dose, temperature and the presence of chloride in the medium showed significant and positive contributions to 2-CP oxidation and mineralization rates. However, after 4 h of treatment, TOC and chlorine measurements were only balanced working with 100% H2O2 and at 70ºC. Under the remaining operating conditions, the gaps in TOC and chlorine balances, especially stressed with substoichiometric H2O2 dose (20%), suggested the presence of unidentified (chlorinated)-organic compounds in the reaction medium.
At 20% H2O2 and after 4 h of treatment, condensation products formed by two-chlorinated aromatic rings including chlorinated diphenylethers and biphenyls, which are known to be intermediates in the formation of PCDD/Fs, were detected. At this time, total PCDD/Fs concentration increased by 12 times in relation to the untreated sample (i.e. no NaCl) and by 80.5 times when NaCl was present in the reaction medium. These results provide evidence for the necessity of using proper operating conditions during the Fenton treatment of chlorinated pollutants, in particular when PCDD/Fs precursors are present (or
163
Conclusions
may be formed in the treated samples) together with the concurrent presence of chloride in the reaction medium.
On the electrochemical and Fenton oxidation of landfill leachates
In chapter 3.3, the electrochemical and Fenton oxidation of leachate samples from a MSW landfill were reported. They were mature leachates characterized mainly by a COD content of 1981 mg L-1, slightly basic pH (7.8), + -1 low BOD5/COD ratio (0.33) and high N-NH4 concentration (985 mg L ). Besides, their non-target full scan analysis depicted the presence of lots of compounds, some of which are considered emerging contaminants. Among them sulfonamides, as is the case of n-butyl benzenesulfonamide (NBBS) accounted for 38.7% of the total detected compounds, followed by phenolic compounds (18.5%), mainly as bisphenol A (BPA) and minor presence of chlorinated phenols, and pharmaceutical and personal care products (PCPPs) (13.5%).
On the other hand, PCDD/Fs were detected in the leachate samples with total concentrations within the range 4479-4981.8 pg L-1. Among total PCDD/Fs, the concentration of 2,3,7,8-PCDD/Fs ranged from 3710 pg L-1 to 4244 pg L-1, which represents the 81.7-86.7% of the total PCDD/Fs. The TEQ of the samples ranged from 13.1 pg I-TEQ L-1 to 17.2 pg I-TEQ L-1 (10-13.7 pg WHO-TEQ L-1). Congener profiles, dominated by OCDD (84.5%) and 1,2,3,4,6,7,8-HpCDD (12.7%), were compared with those found in different landfill leachates reported in the literature by principal components analysis. The next correlation between the concentration of the two dominant congeners and the TEQ of the sample (within the range 0.6-17.2 pg I-TEQ L-1) was obtained (including data from both this work and the aforementioned literature): [pg I-TEQ L-1] = 3.71 [pg 1,2,3,4,6,7,8-HpCDD L-1]1.38 [pg OCDD L-1]-0.89. The similarity between the congener profile found in the studied leachates and those reported in different literature works, led to hypothesize as possible sources of PCDD/Fs in the leachates the following: sewage sludge, PCP formulations, atmospheric deposition and natural formation form precursors.
164
Conclusions
Electrochemical oxidation of landfill leachates
The electrochemical degradation of landfill leachates from a MSW was carried out on BDD anodes under galvanostatic conditions (J = 900 A m-2). Results demonstrated that after 6 h of treatment nearly 95% of COD (< discharge limit, 160 mg L-1) and 90% of TOC were removed, whereas ammonium was completely degraded. Regarding the emerging contaminants NBBS and BPA, after 3 h of treatment (92% COD removal) 47.6% of NBBS remained in the solution whereas 97% of BPA was removed.
As a result of the electrochemical treatment PCDD/Fs reported a general decreasing trend with time. After 3 h of treatment and when the removal of COD started to stagnate (92%), the total concentration of PCDD/Fs was reduced by 71.5%. Regarding the most toxic congeners, 2,3,7,8-PCDD/Fs, their concentration decreased by 71% and the sample TEQ was reduced by nearly 58% (5.5 pg I-TEQ L-1) After the treatment, total PCDD/Fs were still dominated by 2,3,7,8-congeners due to the high concentrations of OCDD and 1,2,3,4,7,8- HpCDD. Fenton oxidation of landfill leachates
Fenton oxidation of landfill leachate samples was carried out using the 2+/ mass ratios H2O2/COD = 1.7 and Fe COD = 0.29. Under these conditions, after
1 h of treatment and coinciding with the complete depletion of H2O2, the concentration of COD and TOC reached a plateau, showing maximum degradation percentages of 75% and 70%, respectively. Although at this time COD remained above the discharge limits, previous studies reported the enhancement of the sample biodegradability. NBBS and BPA were almost completely degraded within the first 5 min of treatment, showing removal percentages of 99% and 96%, respectively.
During the Fenton oxidation of landfill leachates, PCDD/Fs concentrations reported a general increasing trend with time, in contrast to the results reported during their electrochemical treatment. After 3 h of treatment (75% of COD removal), the total concentration of PCDD/Fs increased by 29-146%. Regarding 2,3,7,8-PCDD/Fs, their concentrations increased by 37-148% and therefore sample TEQ increased within the range 12.6% (19.37 pg I-TEQ L-1) and 128.4% (30.6 pg I-TEQ L-1). After the treatment, total PCDD/Fs were still dominated by 2,3,7,8-congeners due to the high concentrations of OCDD and 1,2,3,4,7,8-HpCDD.
165
Conclusions
Conclusiones
La Tesis Doctoral aquí presentada tiene como objetivo la evaluación de la potencial formación de dibenzo-p-dioxinas y dibenzofuranos policlorados (PCDD/Fs) durante la aplicación de procesos de oxidación avanzada (AOPs) en el tratamiento de aguas contaminadas con compuestos orgánicos clorados. En particular, dos tecnologías de oxidación avanzada ampliamente estudiadas como son, la oxidación electroquímica y el proceso Fenton, han sido aplicadas en el tratamiento de disoluciones acuosas de 2-clorofenol (2-CP), precursor de la formación de PCDD/Fs, así como de lixiviados de vertedero caracterizados por elevadas concentraciones de materia orgánica y cloruros.
Sobre la oxidación electroquímica del 2-clorofenol
Como se recoge en el capítulo 3.1, en primer lugar se llevó a cabo la oxidación electroquímica de disoluciones acuosas de 2-clorofenol (15.56 mM) bajo condiciones galvanostáticas (J = 900 A m-2) utilizando ánodos de diamante dopado con boro (BDD) y estudiando la influencia de dos electrolitos comúnmente utilizados como son el NaCl (56.3 mM) y el Na2SO4 (21.1 mM). Aunque con ambos electrolitos se consiguió la completa degradación del 2- CP, ésta fue más rápida en el caso del NaCl, indicando la contribución de la oxidación indirecta por medio de especies de cloro activas generadas a partir de la oxidación del cloruro en la superficie del ánodo. Con respecto a la demanda química de oxigeno (COD) y al carbono orgánico total (TOC), ambos parámetros no se vieron influenciados por el tipo de electrolito. Sin embargo por otro lado se observó que la mineralización del TOC requirió mayores tiempos de reacción que el 2-CP y la COD para un mismo porcentaje de degradación, indicando la formación de productos intermedios de reacción más difíciles de mineralizar.
Como resultado de la electrooxidación del 2-CP se formaron intermedios aromáticos fundamentalmente con sustituciones del anillo aromático en las posiciones orto y para. En el caso de utilizar NaCl se formaron tanto intermedios aromáticos clorados (clorofenoles y compuestos relacionados) como hidroxilados aromáticos (hidroquinona, benzoquinona y catechol principalmente), mientras que con Na2SO4 sólo se detectaron intermedios aromáticos hidroxilados. Finalmente, la rotura del anillo aromático dio lugar a la
166
Conclusions
formación de ácidos orgánicos alifáticos, mayoritariamente ácidos fórmico y oxálico.
Después de 4 horas de tratamiento, el balance al TOC se cerró por completo utilizando Na2SO4. Sin embargo, cuando se empleó NaCl se encontró una diferencia entre el TOC medido experimentalmente (5 mM) y el calculado a partir de las especies identificadas (1 mM). Además, la medida del cloro tampoco fue balanceada indicando así la presencia de especies minoritarias, potencialmente organocloradas, en el medio de reacción. En este sentido, para el mismo tiempo de reacción el uso del NaCl dio lugar a una formación notable de PCDD/Fs con una concentración total de (391, 05 ng L-1), es decir 2,68 x 104 veces superior a la determinada en la muestra de 2-CP sin tratar. Con respecto a los congéneres 2,3,7,8-clorosustituídos, aunque sólo contribuyeron en un 1,6% a la concentración total de los PCDD/Fs, cuya toxicidad expresada como Equivalente Tóxico (TEQ) fue de 220 pg I-TEQ L-1(243 pg WHO-TEQ L-1), valor muy por encima del límite establecido por la EPA en aguas de bebida (30 pg L-1 del equivalente 2,3,7,8-TCDD). Por otro lado, cuando se utilizó Na2SO4 como electrolito, aunque se observó un incremento en la concentración de algunos grupos de homólogos , el aumento de la concentración total de PCDD/Fs fue 134 veces inferior al valor medido en el caso del electrolito NaCl, y no se advirtió la formación de los congéneres más tóxicos, los 2,3,7,8-PCDD/Fs.
Los resultados obtenidos enfatizan la importancia de considerar la coexistencia de cloruros y materia orgánica en el medio de reacción, especialmente cuando potenciales precursores de la formación de dioxinas están presentes o pueden formarse durante el tratamiento. En este caso, aunque el cloruro mostró una influencia positiva en la degradación del 2-CP, favoreció la formación de intermedios de reacción clorados, incluyendo los PCDD/Fs.
Sobre la oxidación Fenton del 2-clorofenol
En el capítulo 3.2 se evaluó la influencia de diferentes variables de operación en el tratamiento Fenton de disoluciones de acuosas de 2-CP (15.56 mM). Para ello, se utilizaron diferentes dosis de H2O2 correspondientes al 20% y 100% de la cantidad estequiométrica necesaria para oxidar el 2-CP. Además, se trabajó a dos temperaturas diferentes, 20ºC y 70 ºC, y se estudió el efecto de la presencia de cloruros en el medio de reacción (56.3 mM).
167
Conclusions
Aunque la oxidación del 2-CP fue efectiva para ambas concentraciones de
H2O2, el uso de la cantidad estequiométrica resultó en una degradación del 2-CP más rápida, así como en mayores porcentajes de degradación para el TOC y la COD, como consecuencia de la mayor disponibilidad de radicales hidroxilo (OH) resultado del aumento de la concentración de H2O2. Sin embargo, la completa mineralización de la materia orgánica estuvo lejos de conseguirse con ambas dosis de H2O2 debido, además de a su consumo, a la posible formación de productos de reacción, como el ácido oxálico, que forman complejos con el hierro desactivándolo. En este sentido, cuando se trabajó con dosis subestequiométricas de H2O2 (20%) el principal producto de reacción identificado fue la 2-clorobenzoquinona, junto con ácidos orgánicos alifáticos
(ácidos acético, fórmico y oxálico). Por otro lado, trabajando con el H2O2 al 100% de la cantidad estequiométrica, los únicos productos de reacción detectados fueron ácidos orgánicos, principalmente oxálico y fórmico.
El incremento de la temperatura de trabajo de 20ºC a 70ºC dio lugar a conversiones más rápidas de H2O2 en OH aumentando el porcentaje de mineralización que fue más acusado para condiciones estequiométricas del
H2O2. Para el H2O2 al 20%, el aumento de la temperatura de trabajo favoreció la conversión de los intermedios aromáticos en ácidos orgánicos más que en el aumento del porcentaje de mineralización.
El efecto de la presencia de cloruro en el medio de reacción se evaluó para dos condiciones de trabajo diferentemente favorecidas en términos de degradación del TOC: 20% H2O2 a 20ºC y 100% H2O2 a 70ºC. En el primero de los sistemas, la presencia de cloruro no afectó ni a la degradación del 2-CP ni al
TOC. Sin embargo, para el 100% de H2O2 a 70ºC, aunque no se observó un efecto en la oxidación del 2-CP, el porcentaje de mineralización aumentó en un 11%. Esta observación podría explicarse por la formación de radicales cloro (Cl) a partir de la reacción entre el cloruro y los OH que contribuyen a la degradación de la materia orgánica.
Las tres variables de operación estudiadas, dosis de H2O2, temperatura y presencia de cloruro en el medio, mostraron en términos generales efectos positivos en la oxidación del 2-CP y en la degradación de la materia orgánica. Sin embargo, tras 4 h de tratamiento, los balances al TOC y al cloro sólo se cerraron en el caso de trabajar con H2O2 al 100% de la cantidad estequiométrica y a 70ºC. Para el resto de condiciones de operación, especialmente en el caso de las condiciones subestequiométricas de H2O2, elevados porcentajes del TOC
168
Conclusions
permanecieron sin identificar sugiriendo la presencia de compuestos organo(clorados) en el medio de reacción.
Trabajando en condiciones subestequiométricas de H2O2, después de 4 horas de tratamiento se detectaron compuestos de condensación, fundamentalmente difenil éteres y bifenilos policlorados, los cuales son conocidos intermedios en la formación de PCDD/Fs. Para el mismo tiempo de reacción, la concentración total de PCDD/Fs se incrementó 12 veces con respecto a la concentración presente en la muestra de 2-CP sin tratar. Este aumento fue incluso mayor, 80.5 veces, en el caso de añadir cloruro al medio de reacción.
Sobre la oxidación electroquímica y el tratamiento Fenton de lixiviados de vertedero
El capítulo 3.3 de la presente Tesis tuvo como objetivo el análisis de la evolución de la concentración de PCDD/Fs en muestras de lixiviados de vertedero de residuos sólidos urbanos, los cuales fueron tratados mediante oxidación electroquímica y Fenton. Los lixiviados se caracterizaron por una relativa alta carga orgánica, COD = 1981 mg L-1, pH básico (7.8), baja biodegradabilidad expresada como BOD5/COD (0.33) y alta concentración de + -1 amonio (N-NH4 = 985 mg L ). Además, el análisis de los lixiviados mediante GC- MS en modo scan permitió la identificación de numerosos compuestos orgánicos, algunos de los cuales se consideran contaminantes emergentes. Entre ellos, la n-butil benceno sulfonamida (NBBS) fue el compuesto mayoritario, contribuyendo en un 38.7% al total de los compuestos orgánicos identificados, seguido de la familia de los fenoles (18.5%), en la que destacó el bisfenol (BPA), y de los productos farmacéuticos y de higiene personal (PCPPs) (13.5%).
Por otro lado, del análisis de PCDD/Fs en las muestras de lixiviados se obtuvieron concentraciones totales en el rango de 4479-4981.8 pg L-1. Los congéneres 2,3,7,8-clorosustituidos contribuyeron en un 81.7-86.7% a la concentración total de los PCDD/Fs, lo que supone concentraciones comprendidas entre 3710 y 4244 pg L-1 y toxicidades equivalentes de 13.1-17.2 pg I-TEQ L-1 (10-13.7 pg WHO-TEQ L-1). De este análisis se obtuvieron perfiles de concentración dominados por la OCDD (84.5%) y la 1,2,3,4,6,7,8-HpCDD (12.7%). Estos perfiles fueron similares a los mostrados en la literatura en el
169
Conclusions
análisis en muestras de lixiviado de vertedero, con los que fueron comparados mediante el análisis de componentes principales. Como complemento a esta comparación se obtuvo la siguiente correlación mediante regresión no lineal entre la toxicidad equivalente de la muestra y la concentración de los dos congéneres mayoritarios (en el rango de 6-17.2 pg I-TEQ L-1): [pg I-TEQ L-1] = 3.71 [pg 1,2,3,4,6,7,8-HpCDD L-1]1.38 [pg OCDD L-1]-0.89. Por otro lado, la similitud encontrada entre el perfil de los congéneres 2,3,7,8-PCDD/Fs obtenido en este trabajo con el mostrado en la literatura para diferentes matrices llevó a suponer como posibles fuentes responsables de la presencia de los PCDD/Fs en los lixiviados las siguientes: lodos de depuradora, formulaciones químicas de pentaclorofenol, deposición atmosférica y posible formación natural a partir de precursores.
Oxidación electroquímica de lixiviados de vertedero
El tratamiento electroquímico de los lixiviados de vertedero de residuos sólidos urbanos se llevó a cabo utilizando ánodos de diamante dopado con boro bajo condiciones galvanostáticas (J = 900 A m-2). Como resultado del tratamiento electroquímico, tras 6 horas de operación la COD y el TOC fueron degradados en un 95% (por debajo del límite de vertido, 160 mg L-1) y 90% respectivamente, mientras que el amonio se eliminó por completo. Con respecta a los contaminantes emergentes NBBS y BPA, después de 3 horas de tratamiento el BPA fue prácticamente eliminado, mientras que un 47.6% de NBBs permaneció en el medio de reacción. En el caso de los PCDD/Fs la concentración total se redujo en un 71.5% tras 3 horas de tratamiento (92% de degradación de la COD). En el caso de los congéneres 2,3,7,8-PCDD/Fs, su concentración disminuyó en un 71% y la Toxicidad Equivalente se redujo en un 58% (5.5 pg I-TEQ L-1).
Oxidación Fenton de lixiviados de vertedero
El tratamiento de los lixiviados mediante el proceso Fenton se llevó a cabo 2+/ utilizando las relaciones másicas: O2/COD = 1.7 and Fe COD = 0.29. Tras 1 hora de tratamiento, y coincidiendo con el consumo total del H2O2, la COD y el TOC se estabilizaron alcanzando máximos porcentajes de degradación del 75% y 70% respectivamente. Aunque la COD se mantuvo por encima del límite de vertido, estudios previos del grupo de investigación mostraron un aumento de la biodegradabilidad de la muestra. Con relación a los compuestos NBBS y al BPA, fueron prácticamente degradados durante los primeros 5 minutos de reacción,
170
Conclusions
con porcentajes de eliminación del 99% y 96% respectivamente. Por otro lado, contrario a los resultados observados en la electrooxidación, la concentración total de los PCDD/Fs aumento en el rango 29-146%. Considerando los congéneres más tóxicos, el aumento de su concentración (en el rango de 37- 148%) se tradujo en un incremento de la toxicidad equivalente del 12.6-128.4% (equivalente a una concentración total de los congéneres 2,3,7,8-PCDD/Fs de 19.37-30.6 pg I-TEQ L-1).
171
Annexes
Annexes
ANNEX I. PCDD/Fs NOMENCLATURE
Table I.1. Name and abbreviations used for 17 2,3,7,8-PCDD/Fs.
Name Abbreviation 2,3,7,8-Tetrachlorodibenzo-p-furan 2,3,7,8-TCDF 2,3,7,8-Tetrachlorodibenzo-p-dioxin 2,3,7,8-TCDD 1,2,3,7,8-Pentachlorodibenzo-p-furan 1,2,3,7,8-PeCDF 2,3,4,7,8-Pentachlorodibenzo-p-furan 2,3,4,7,8-PeCDF 1,2,3,7,8-Pentachlorodibenzo-p-dioxin 1,2,3,7,8-PeCDD 1,2,3,4,7,8-Hexachlorodibenzo-p-furan 1,2,3,4,7,8-HxCDF 1,2,3,6,7,8-Hexachlorodibenzo-p-furan 1,2,3,6,7,8-HxCDF 2,3,4,6,7,8-Hexachlorodibenzo-p-furan 2,3,4,6,7,8-HxCDF 1,2,3,7,8,9-Hexachlorodibenzo-p-furan 1,2,3,7,8,9-HxCDF 1,2,3,4,7,8-Hexachlorodibenzo-p dioxin 1,2,3,4,7,8-HxCDD 1,2,3,6,7,8-Hexachlorodibenzo-p-dioxin 1,2,3,6,7,8-HxCDD 1,2,3,7,8,9-Hexachlorodibenzo-p-dioxin 1,2,3,7,8,9-HxCDD 1,2,3,4,6,7,8-Heptachlorodibenzo-p-furan 1,2,3,4,6,7,8-HpCDF 1,2,3,4,7,8,9-Heptachlorodibenzo-p-furan 1,2,3,4,7,8,9-HpCDF 1,2,3,4,6,7,8-Heptachlorodibenzo-p-dioxin 1,2,3,4,6,7,8-HpCDD Octaclorodibenzo-p-furan OCDF Octaclorodibenzo-p-dioxin OCDD
163
Annexes
Table I.2. Total PCDD/Fs abbreviations.
Name Abbreviation
Total natives-TCDD TCDD
Total natives-PeCDD PeCDD
Total natives -HxCDD HxCDD
Total natives -HpCDD HpCDD
Total natives -TCDF TCDF
Total natives -PeCDF PeCDF
Total natives -HxCDF HxCDF
Total nativos -HpCDF HpCDF
164 Annexes
ANNEX II.GENERAL NOMENCLATURE
AOPs Advanced oxidation processes BDD Boron-doped diamond BPA Bisphenol A CBz Chlorobenzene CDEs chlorinated diphenylethers COD Chemical oxygen demand CWA Clean Water act CBzs CPs Chlorophenols D Diffusivity EC Emerging contaminant EDC Endocrine disruptors compound 2,4-DCP 2,4-Dichlorophenol 2,6-DCP 2,6-Dichlorophenol
dh Equivalent diameter DHCD Dihydroxycyclohexadienyl radical EQS Environmental quality standards F Faraday’s constant I-TEF International toxicity equivalency factors J Current density
Jlim Limiting current density
Km Mass transport coefficient
LD50 Lethal dose for 50% of the population MCL Maximum contaminant level MCLG Maximum contaminant level goal MSW Municipal solid waste PCP-Na Sodium-pentachlorophenate NBBS n-Butyl benzenesulfonamide OH Hydroxyl radicals
165 Annexes
PC Principal component PCBs Polychlorinated biphenyls PCA Principal component analysis PCP Pentachlorophenol PCPPs Polychlorinated phenoxy phenols PPCPs Pharmaceutical and personal care products POP Persistent organic pollutants Re Reynolds number SDWA Safe Drinking Water Act Sc Schmidt number Sh Sherwood number TEQ Toxic equivalents TOC Total organic carbon TEF Toxicity equivalency factor 2,3,6-TCP 2,3,6-Trichlorophenol TSS Total suspended solids u Fluid velocity WFD Water Framework Directive WHO-TEF World Health Organization toxicity equivalency factors ρ Density µ Viscosity
166 Annexes
ANNEX III. RESULTS FROM PCDD/Fs ANALYSES DISCUSSED IN CHAPTER 3
A. PCDD/Fs analyses in the electrochemical oxidation of 2-chlorophenol
Table III.1. Total PCDD/Fs concentrations in the electrochemical oxidized 2-CP solutions with NaCl (t= 4 h) (n=2).
Homologue Group Mean (pg L-1) RSD (%) TCDD 13811.22 1.00 PeCDD 21066.36 58.97 HxCDD 9099.20 24.72 HpCDD 1564.62 18.86 OCDD 1116.64 1.92 TCDF 22362.51 27.80 PeCDF 49004.09 89.26 HxCDF 265657.44 28.12 HpCDF 6619.64 3.31 OCDF 751.32 30.61 Total 391053.02
167 Annexes
Table III.2. 2,3,7,8- PCDD/Fs concentrations in the electrochemical oxidized 2- CP solutions with NaCl (t= 4 h) (n=2).
Mean DL Congener -1 RSD (%) -1 (pg L ) (pg L ) 2,3,7,8-TCDD < DL - - 1,2,3,7,8-PeCDD 74.00 35.09 1.56 1,2,3,4,7,8-HxCDD 50.76 20.24 1.62 1,2,3,6,7,8-HxCDD 683.57 25.60 1.64 1,2,3,7,8,9-HxCDD 68.69 44.45 0.5 1,2,3,4,6,7,8-HpCDD 1070.81 17.89 0.3 OCDD 1116.64 1.92 0.7 2,3,7,8-TCDF < DL - 2.5 1,2,3,7,8-PeCDF < DL - 2.56 2,3,4,7,8-PeCDF 63.39 6.01 2.08 1,2,3,4,7,8-HxCDF 21.02 48.69 2.08 1,2,3,6,7,8-HxCDF < DL - 2.22 1,2,3,7,8,9-HxCDF < DL - 2.06 2,3,4,6,7,8-HxCDF 351.57 107.28 0.64 1,2,3,4,6,7,8-HpCDF 2052.77 34.51 0.76 1,2,3,4,7,8,9-HpCDF 6.55 33.10 0.28 OCDF 751.32 30.61 3.08 Total 6311.0925
168 Annexes
Table III.3. Total PCDD/Fs concentrations in the electrochemical oxidized 2-CP
solutions with Na2SO4 (t= 4 h) (n=2).
Homologue Group Mean (pg L-1) RSD (%) TCDD 1505.93 32 PeCDD - - HxCDD 70.71 27.3 HpCDD - - OCDD - - TCDF 1160.7 29.1 PeCDF - - HxCDF - - HpCDF - - OCDF - - Total 2737.34
169 Annexes
Table III.4. 2,3,7,8- PCDD/Fs concentrations in the electrochemical oxidized 2-
CP solutions with Na2SO4 (t= 4 h) (n=2).
Mean DL Congener -1 RSD (%) -1 (pg L ) (pg L ) 2,3,7,8-TCDD < DL - 0.58 1,2,3,7,8-PeCDD < DL - 1.14 1,2,3,4,7,8-HxCDD < DL - 0.46 1,2,3,6,7,8-HxCDD < DL - 0.48 1,2,3,7,8,9-HxCDD < DL - 0.5 1,2,3,4,6,7,8-HpCDD < DL - 0.28 OCDD < DL - 0.36 2,3,7,8-TCDF < DL - 0.38 1,2,3,7,8-PeCDF 0.68 58.23 0.28 2,3,4,7,8-PeCDF 1.27 21.16 0.66 1,2,3,4,7,8-HxCDF < DL - 0.5 1,2,3,6,7,8-HxCDF < DL - 0.5 1,2,3,7,8,9-HxCDF < DL - 0.58 2,3,4,6,7,8-HxCDF 1.34 42.23 0.52 1,2,3,4,6,7,8-HpCDF < DL - 0.1 1,2,3,4,7,8,9-HpCDF < DL - 0.16 OCDF < DL - 0.1 Total 3.29
170 Annexes
B. PCDD/Fs analyses in the Fenton oxidation of 2-chlorophenol
Table III.5. Total PCDD/Fs concentrations in the Fenton oxidized 2-CP solutions
with 100% H2O2 (t= 4 h) (n=2).
20ºC 70ºC 70ºC, NaCl Homologue Group Mean RSD Mean RSD Mean RSD (pg L-1) (%) (pg L-1) (%) (pg L-1) (%) TCDD 8.66 84.00 2.95 0.64 0.81 20.21 PeCDD 1.73 10.18 0.96 0.64 0.57 141.42 HxCDD 0.15 173.21 0.32 0.64 0.15 141.42 HpCDD 0.61 48.58 0.09 0.64 0.46 58.10 OCDD 0.75 58.08 0.81 7.93 0.80 21.40 TCDF 4.25 17.19 2.68 0.64 1.40 33.22 PeCDF 0.53 87.86 0.36 0.64 0.23 55.39 HxCDF 1.92 27.22 0.49 0.64 0.55 3.17 HpCDF 0.04 173.21 < DL - 0.32 36.85 OCDF 0.49 88.49 0.70 0.64 0.08 141.42 Total 19.13 9.34 5.38
171 Annexes
Table III.6. 2,3,7,8-PCDD/Fs concentrations in the Fenton oxidized 2-CP
solutions with 100% H2O2 (t= 4 h) (n=2).
20ºC 70ºC 70ºC, NaCl Congener Mean Mean RSD Mean RSD RSD (%) (pg L-1) (pg L-1) (%) (pg L-1) (%) 2,3,7,8-TCDD < DL - < DL - < DL - 1,2,3,7,8-PeCDD 0.03 173.21 < DL - < DL - 1,2,3,4,7,8-HxCDD < DL - < DL - < DL - 1,2,3,6,7,8-HxCDD < DL - < DL - < DL - 1,2,3,7,8,9-HxCDD < DL - < DL - < DL - 1,2,3,4,6,7,8- 0.47 87.11 < DL - 0.20 141.42 HpCDD OCDD 0.75 58.08 < DL 7.93 0.80 21.40 2,3,7,8-TCDF 0.01 173.21 0.81 < DL < DL - 1,2,3,7,8-PeCDF 0.19 70.50 < DL < DL < DL - 2,3,4,7,8-PeCDF 0.43 28.49 < DL 34.54 0.19 36.46 1,2,3,4,7,8-HxCDF 0.16 86.90 0.31 50.74 < DL - 1,2,3,6,7,8-HxCDF < DL < DL 0.19 44.09 < DL - 1,2,3,7,8,9-HxCDF 0.24 173.21 0.16 0.21 26.25 2,3,4,6,7,8-HxCDF 0.10 173.21 < DL 141.42 0.05 141.42 1,2,3,4,6,7,8- < DL < DL 0.09 < DL 0.23 23.24 HpCDF 1,2,3,4,7,8,9- < DL < DL < DL 141.42 0.15 6.27 HpCDF OCDF 0.49 88.49 0.20 0.64 0.08 141.42 Total 2.87 0.70 1.90
172 Annexes
Table III.7. Total PCDD/Fs concentrations in the Fenton oxidized 2-CP solutions
with 20% H2O2 (t= 4 h) (n=2).
20ºC 20ºC, NaCl Homologue Group Mean RSD Mean RSD (pg L-1) (%) (pg L-1) (%) TCDD 2.33 19.16 348.74 51.03 PeCDD 0.37 141.42 50.06 53.66 HxCDD < DL - 3.70 2.55 HpCDD 0.83 141.42 1.38 59.64 OCDD 2.33 64.94 1.83 23.14 TCDF 303.64 3.55 677.14 40.70 PeCDF 1.22 61.11 968.56 1.45 HxCDF 1.43 11.41 37.34 38.21 HpCDF 0.39 12.86 2.23 44.79 OCDF 0.34 141.42 0.97 73.15 Total 312.85 2091.950
173 Annexes
Table III.8. 2,3,7,8-PCDD/Fs concentrations in the Fenton oxidized 2-CP
solutions with 20% H2O2 (t= 4 h) (n=2).
20ºC 20ºC, NaCl Congener Mean RSD Mean RSD (pg L-1) (%) (pg L-1) (%) 2,3,7,8-TCDD < DL - < DL - 1,2,3,7,8-PeCDD < DL - < DL - 1,2,3,4,7,8-HxCDD 0.41 173.21 0.52 - 1,2,3,6,7,8-HxCDD 0.38 173.21 1.14 - 1,2,3,7,8,9-HxCDD 0.42 93.11 0.82 - 1,2,3,4,6,7,8-HpCDD 1.00 51.97 1.18 - OCDD 2.58 27.63 1.83 7.93 2,3,7,8-TCDF 6.89 55.45 8.43 - 1,2,3,7,8-PeCDF 0.19 173.21 0.53 - 2,3,4,7,8-PeCDF 0.61 89.34 0.78 34.54 1,2,3,4,7,8-HxCDF 0.04 173.21 < DL 50.74 1,2,3,6,7,8-HxCDF < DL - < DL 44.09 1,2,3,7,8,9-HxCDF 0.20 173.21 0.45 141.4 2,3,4,6,7,8-HxCDF < DL - 0.03 2 1,2,3,4,6,7,8-HpCDF 0.12 173.21 0.38 - 141.4 1,2,3,4,7,8,9-HpCDF 0.27 173.21 0.39 2 OCDF 1.32 55.39 0.97 0.64 Total 14.42 17.44
174 Annexes
C. PCDD/Fs analyses in landfill leachates samples
Table III.9. Total PCDD/Fs concentrations in raw leachate samples S1, S2 and S3.
-1 -1 -1 Homologue Group S1 (pg L ) S2 (pg L ) S3 (pg L ) TCDD 5.53 3.16 < DL PeCDD 11.87 6.94 20.52 HxCDD 205.60 124.48 202.78 HpCDD 1065.80 910.08 1013.24 OCDD 3391.00 3680.2 3111.8 TCDF 114.33 18.62 13.26 PeCDF 55.33 11.3 19.44 HxCDF 38.17 36.42 26.64 HpCDF 48.50 60.68 31.98 OCDF 45.67 41.92 39.38 Total 4981.8 4893.8 4479.04
Table III.10. Total PCDD/Fs concentrations in the electrochemically oxidized leachate samples.
Homologue EOX 5 EOX 30 EOX 180 Group (pg L-1) (pg L-1) (pg L-1) TCDD 0.98 1.86 < DL PeCDD 4.52 8.98 1.52 HxCDD 83.04 69.9 26.54 HpCDD 578.46 495.46 216.84 OCDD 2367.54 2093.5 1059.14 TCDF 18.42 177.7 37.86 PeCDF 16.12 53.06 23.86 HxCDF 20.08 67.4 3.66 HpCDF 31.94 31.38 13.66 OCDF 33.78 22.98 13.02 Total 3154.88 3022.22 1396.1
175 Annexes
Table III.11. Total PCDD/Fs concentrations in the Fenton oxidized leachate samples.
Homologue Sample S1 Sample S3 Group Fenton 5 Fenton 30 Fenton 180 Fenton 180 (pg L-1) (pg L-1) (pg L-1) (pg L-1) TCDD 3.90 7.70 0.33 0 PeCDD 35.20 20.73 18.80 54.54 HxCDD 228.40 100.93 212.90 428.7 HpCDD 1242.17 1362.33 1348.60 2437.04 OCDD 3960.00 4599.67 4662.40 7793.4 TCDF 45.50 49.20 28.70 33.42 PeCDF 39.83 37.87 21.90 41.24 HxCDF 42.20 56.30 38.47 65.92 HpCDF 61.17 70.87 60.44 75.58 OCDF 54.83 72.00 67.84 92.94 Total 5713.20 6377.60 6460.37 11022.78
176 Annexes
Table III.12. Concentrations of 2,3,7,8-PCDD/Fs in raw leachate samples S1, S2
and S3.
S (n=2) S S 1 2 3 -1 RSD DL -1 DL -1 DL Congener Pg L Pg L Pg L (%) (pg L-1) (pg L-1) (pg L-1) 2,3,7,8-TCDD < DL 0.0 0.57 < DL 0.12 < DL 0.1 1,2,3,7,8-PeCDD < DL 0.0 1.17 < DL 0.44 < DL 0.46 1,2,3,4,7,8-HxCDD 13.00 71.9 2.53 10.20 1.18 11.14 0.86 1,2,3,6,7,8-HxCDD 12.67 34.8 2.50 10.12 1.2 11.12 0.86 1,2,3,7,8,9-HxCDD 5.67 41.4 2.60 3.32 1.2 3.76 0.9 1,2,3,4,6,7,8-HpCDD 565.00 5.0 4.87 465.10 1.38 496.60 1.04 3391.0 OCDD 1.0 8.67 3680.20 2.12 3111.80 3.26 0 2,3,7,8-TCDF 9.53 51.6 0.80 1.76 0.22 2.36 0.24 1,2,3,7,8-PeCDF < DL 0.0 1.27 < DL 0.34 2.10 0.32 2,3,4,7,8-PeCDF 5.67 41.4 1.10 2.16 0.34 2.46 0.3 1,2,3,4,7,8-HxCDF 3.67 9.6 0.87 3.06 0.56 2.84 0.46 1,2,3,6,7,8-HxCDF 3.00 14.1 0.83 3.34 0.52 3.26 0.46 1,2,3,7,8,9-HxCDF < DL 0.0 1.10 0.00 0.58 < DL 0.52 2,3,4,6,7,8-HxCDF 3.33 13.4 0.73 2.60 0.5 3.18 0.5 1,2,3,4,6,7,8-HpCDF 21.00 13.0 1.27 18.34 0.6 18.52 0.64 1,2,3,4,7,8,9-HpCDF < DL 0.0 1.43 2.22 0.64 1.52 0.7 OCDF 45.67 4.4 1.60 41.92 0.5 39.38 0.46
Total 4079.2 4244.34 3710.04
177 Annexes
Table III.13. Concentrations of 2,3,7,8-PCDD/Fs in the electrochemical oxidized
S2 samples.
EOX 5 min EOX 30 min EOX 180 min Congener -1 DL -1 DL -1 DL Pg L Pg L Pg L (pg L-1) (pg L-1) (pg L-1) 2,3,7,8-TCDD < DL 0.2 0.96 0.46 < DL 7.4 1,2,3,7,8-PeCDD < DL 0.36 < DL 0.34 < DL 0.5 1,2,3,4,7,8-HxCDD 5.5 0.88 5.12 0.72 2.8 0.54 1,2,3,6,7,8-HxCDD 5.6 0.9 5.1 0.72 2.68 0.5 1,2,3,7,8,9-HxCDD 2 0.94 < DL 0.72 < DL 0.52 1,2,3,4,6,7,8-HpCDD 320.3 1.5 265.8 1.02 123.82 0.94 OCDD 2367.54 3.16 2093.5 1.06 1059.14 1.2 2,3,7,8-TCDF 2.36 0.24 10.64 0.74 12.52 0.32 1,2,3,7,8-PeCDF 0.84 0.32 2.36 0.68 0.98 0.38 2,3,4,7,8-PeCDF 1.48 0.3 5.02 0.64 1.72 0.36 1,2,3,4,7,8-HxCDF 1.72 0.38 2.46 0.48 1.18 0.22 1,2,3,6,7,8-HxCDF 1.84 0.38 1.72 0.44 0.86 0.22 1,2,3,7,8,9-HxCDF < DL 0.46 1.56 0.54 < DL 0.28 2,3,4,6,7,8-HxCDF 1.42 0.36 2.18 0.46 1.82 0.24 1,2,3,4,6,7,8-HpCDF 12.18 0.52 13.52 0.42 5.5 0.32 1,2,3,4,7,8,9-HpCDF 1.32 0.6 1.32 0.48 0.56 0.36 OCDF 33.78 0.46 22.98 0.52 13.02 0.3
Total 2757.88 2434.24 1226.6
178 Annexes
Table III.14. Concentrations of 2,3,7,8-PCDD/Fs in the Fenton oxidized S1 samples.
Fenton 5 min Fenton 30 min Fenton 180 min
-1 DL -1 DL -1 DL Congener Pg L Pg L Pg L (pg L-1) (pg L-1) (pg L-1) 2,3,7,8-TCDD < DL 0.77 < DL 0.53 < DL 0.33 1,2,3,7,8-PeCDD < DL 1.63 < DL 1.77 < DL 1.07 1,2,3,4,7,8-HxCDD 14.33 3.67 < DL 3.97 12.57 3.37 1,2,3,6,7,8-HxCDD 14.17 3.67 11.67 3.67 11.63 3.07 1,2,3,7,8,9-HxCDD 7.30 4.00 8.00 4.67 7.87 3.13 1,2,3,4,6,7,8-HpCDD 640.00 7.33 746.33 7.07 777.53 7.90 OCDD 3960.00 12.33 4599.67 13.27 4662.40 12.73 2,3,7,8-TCDF 13.33 0.97 6.33 1.00 3.10 0.77 1,2,3,7,8-PeCDF < DL 1.63 4.00 1.57 2.50 1.47 2,3,4,7,8-PeCDF 0.00 1.57 4.33 1.37 5.00 1.07 1,2,3,4,7,8-HxCDF 7.73 1.63 < DL 1.97 4.67 1.50 1,2,3,6,7,8-HxCDF 0.00 1.67 4.70 1.60 < DL 1.47 1,2,3,7,8,9-HxCDF < DL 2.10 < DL 2.30 < DL 1.83 2,3,4,6,7,8-HxCDF 5.17 1.57 5.00 1.70 < DL 1.30 1,2,3,4,6,7,8-HpCDF 27.33 1.83 29.33 2.43 23.40 1.83 1,2,3,4,7,8,9-HpCDF 5.33 2.03 3.33 2.73 2.40 2.07 OCDF 54.83 1.90 72.00 2.17 67.83 1.63
Total 4754.27 5494.70 5580.90
179 Annexes
Table III.15. Concentrations of 2,3,7,8-PCDD/Fs in the Fenton oxidized S3 sample.
Fenton 180 min Congener DL Pg L-1 (pg L-1) 2,3,7,8-TCDD < DL 0.14 1,2,3,7,8-PeCDD < DL 0.7 1,2,3,4,7,8-HxCDD 25.2 1.66 1,2,3,6,7,8-HxCDD 25.9 1.7 1,2,3,7,8,9-HxCDD 7.74 1.78 1,2,3,4,6,7,8-HpCDD 1169.42 2.16 OCDD 7793.4 3.9 2,3,7,8-TCDF 4.16 0.34 1,2,3,7,8-PeCDF 2.68 0.54 2,3,4,7,8-PeCDF 4.44 0.52 1,2,3,4,7,8-HxCDF 6.68 0.82 1,2,3,6,7,8-HxCDF 5.96 0.86 1,2,3,7,8,9-HxCDF 0 0.96 2,3,4,6,7,8-HxCDF 5.96 0.88 1,2,3,4,6,7,8-HpCDF 43.46 0.98 1,2,3,4,7,8,9-HpCDF 4.98 1.06 OCDF 92.94 0.5 Total 9192.92
180 List of Scientific Contributions
Publications in International Journals
Vallejo, M., San Román, M. F., Irabien, A., Ortiz, I., 2013. Comparative study of the destruction of polychlorinated dibenzo-p-dioxins and dibenzofurans during Fenton and electrochemical oxidation of landfill leachates. Chemosphere 90, 132-138. IF: 3.137.
Vallejo, M., San Román, M. F., Ortiz, I., 2013. Quantitative assessment of the formation of polychlorinated derivatives, PCDD/Fs, in the electrochemical oxidation of 2-chlorophenol as function of the electrolyte type. Environmental Science and Technology 47, 12400-12408. IF: 5.257.
Vallejo, M., San Román, M. F., Ortiz, I., Irabien, A., 2014. Overview of the PCDD/Fs degradation potential and formation risk in the application of advanced oxidation processes (AOPs) to wastewater treatment. Chemosphere, accepted. IF: 3.137.
Vallejo, M., San Román, M. F., Ortiz, I., Irabien, A., 2014. The critical role of the operating conditions on the Fenton oxidation of 2-chlorophenol: Assessment of PCDD/Fs formation. Sent to Journal of Hazardous Materials, under revision.
Contributions to International Conferences
San Román, M.F., Bringas, E., Vallejo, M., Irabien, A., Ortiz, I. Solvent extraction of zinc in HCl media using trybutylphosphate (TBP) and liquid membranes. 19th International Solvent Extraction Conference (ISEC2011), Santiago (Chile) (Octubre, 2011). Conferencia Internacional.
Vallejo, M., San Román, M.F., Ortiz, I. Determination of PCDD/Fs in the leachates of a municipal wastes landfill”. III Reunión Nacional de Dioxinas, Furanos y Compuestos Orgánicos Persistentes Relacionados. Universidad de Cantabria, Santander (España) (Junio-Julio, 2011). Conferencia Internacional
Vallejo, M., San Román, M.F., Ortiz, I. Advanced oxidation treatment of landfill leachates: study of PCDD/Fs evolution. Congreso Internacional de
Química de la ANQUE “Innovando para el futuro”. Sevilla, España. (Junio, 2012). Conferencia Internacional.
Vallejo, M., San Román, M.F., Ortiz, I. Changes of PCDD/Fs concentration during the advanced oxidation of landfill leachates. 20th International Congress of Chemical and Process Engineering CHISA 2012. Prague (Czech Republic) (Agosto 2012). Conferencia Internacional.
Vallejo, M., San Román, M.F., Ortiz, I. Changes in PCDD/Fs concentration during the advanced oxidation treatment of landfill leachates. IV reunión nacional de dioxinas, furanos y compuestos orgánicos relacionados. Alicante, España (Junio 2013). Conferencia Internacional.
Vallejo, M., San Román, M.F., Ortiz, I. Study of PCDD/Fs formation during the advanced oxidation of 2-chlorophenol. XXXIV Reunión Bienal de la Real Sociedad Española de Química, Santander, España (Septiembre, 2013). Conferencia Internacional.