SISTEMAS CATALÍTICOS BASADOS EN DISTINTOS
Transcript of SISTEMAS CATALÍTICOS BASADOS EN DISTINTOS
Departamento de Ingeniería Química y Química Física
TESIS DOCTORAL
SISTEMAS CATALÍTICOS BASADOS EN DISTINTOS
SEMICONDUCTORES PARA LA ELIMINACIÓN
DE CONTAMINANTES EN AGUA MEDIANTE
OZONIZACIÓN FOTOCATALÍTICA SOLAR
ESTEFANÍA MENA RUBIO
Conformidad de los Directores de Tesis:
Fdo.: Ana Rey Barroso Fdo.: Eva M. Rodríguez Franco
Badajoz, 2017
A mis padres
La realización de esta Tesis Doctoral ha sido posible gracias a la concesión de
una beca de Formación de Personal Investigador (Ref. PD12059) financiada por
la Consejería de Empleo, Empresa e Innovación (Junta de Extremadura) y el
Fondo Social Europeo, así como al apoyo económico prestado a través de los
proyectos de investigación de la Junta de Extremadura (GRU10012 y GR15033) y
a través de los proyectos del Programa Nacional de Investigación
CTQ2009/13459/C05/05, titulado “Integración de procesos de fotocatálisis solar en la
depuración biológica de aguas residuales para la eliminación de contaminantes
emergentes”; CTQ2012/35789/C02/01, titulado “Preparación de catalizadores y su
aplicación en la eliminación de contaminantes refractarios de aguas residuales mediante
ozonización fotocatalítica”; y CTQ2015‐64944‐R titulado “Led y fotocatalizadores
polifuncionales basados en grafeno y estructuras metal‐orgánicas para el tratamiento de
aguas por ozonización fotocatalítica” del Ministerio de Economía y Competitividad
(España) y del Fondo Europeo de Desarrollo Regional (FEDER).
Llegado este momento, no puedo dar por finalizado mi trabajo sin antes
agradecer a todas las personas que me han acompañado en esta etapa.
Gran parte de este trabajo es vuestro. A todos, muchas gracias.
ÍNDICE
RESUMEN ………………………………………………………………………...... 1
CAPÍTULO 1: INTRODUCCIÓN Y OBJETIVOS …………………………….. 9
1.1. ANTECEDENTES Y JUSTIFICACIÓN DE LA TESIS ………………... 11
1.1.1. Contextualización del trabajo ...…………........................................ 11
1.1.2. El agua en el mundo actual …..…………........................................ 11
1.1.2.1. Marco legal del agua …..………………...................................... 16
1.1.2.2. Conservación de los recursos hídricos …..………………............ 21
1.1.3. Procesos avanzados de oxidación …..………................................. 25
1.1.3.1. Fotocatálisis heterogénea ………………...................................... 26
1.1.3.2. Combinación de ozono y fotocatálisis heterogénea:
ozonización fotocatalítica …..………………............................................ 30
1.1.4. Optimización de los procesos fotocatalíticos: empleo de
radiación solar y fotocatalizadores apropiados …................................... 38
1.2. OBJETIVOS Y ALCANCE DEL TRABAJO …………………..………... 44
1.3. ORGANIZACIÓN DE LA MEMORIA ………...……………..………... 46
BIBLIOGRAFÍA ………...……………..……………………………….……... 47
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II
CAPÍTULO 2: MATERIALES Y MÉTODOS EXPERIMENTALES ………..... 61
2.1. INSTALACIONES Y PROCEDIMIENTOS EXPERIMENTALES ..….. 63
2.1.1. Tratamientos de eliminación de contaminantes en agua ...…….. 63
2.1.1.1. Ensayos realizados empleando lámparas de luz UVA
(luz negra) como fuente de radiación …..…………….............................. 63
2.1.1.2. Ensayos realizados empleando lámparas de luz solar artificial
como fuente de radiación …..……………................................................. 66
2.1.2. Preparación de los catalizadores empleados …..………............... 68
2.1.2.1. Agitador ………………............................................................... 68
2.1.2.2. Autoclave ………………............................................................. 69
2.1.2.3. Centrífuga …….………............................................................... 69
2.1.2.4. Estufa ……..……………............................................................. 69
2.1.2.5. Horno mufla ……..……………................................................... 69
2.2. EQUIPOS Y MÉTODOS DE ANÁLISIS ……………….……..………... 70
2.2.1. Seguimiento de la eficacia de los tratamientos de eliminación
de contaminantes en agua …..………........................................................ 70
2.2.1.1. Determinación de la concentración de contaminantes modelo
en agua ……………….............................................................................. 72
2.2.1.2. Identificación de productos intermedios de degradación .............. 73
2.2.1.3. Determinación de la concentración de ácidos carboxílicos de
bajo peso molecular y aniones inorgánicos en disolución ......................... 74
2.2.1.4. Determinación de la concentración de carbono orgánico total
en disolución (COT) .................................................................................. 75
2.2.1.5. Determinación de la concentración de ozono en disolución
acuosa …………………………………………………………………… 76
2.2.1.6. Determinación de la concentración de ozono en fase gas ………. 77
2.2.1.7. Determinación de la concentración de peróxido de hidrógeno
en disolución ............................................................................................. 77
2.2.1.8. Determinación de la concentración de hierro total en
disolución .................................................................................................. 78
2.2.1.9. Determinación de la concentración de hierro (II) en disolución .. 79
2.2.1.10. Determinación de la intensidad de radiación ………………..... 80
2.2.1.11. Determinación de la concentración de formaldehído en
disolución ………………………………………………………………. 82
Índice
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2.2.1.12. Determinación de la demanda química de oxígeno (DQO) ....... 83
2.2.1.13. Determinación de la biodegradabilidad: demanda biológica
de oxígeno (DBO) ..................................................................................... 83
2.2.1.14. Determinación de la absorbancia a 254 nm: Aromaticidad ....... 84
2.2.1.15. Determinación de la concentración de fosfatos en disolución .... 85
2.2.1.16. Determinación de la turbidez ………………………………..... 85
2.2.1.17. Determinación de pH, conductividad y temperatura ……........ 86
2.2.2. Caracterización de los catalizadores empleados …..………......... 86
2.2.2.1. Espectroscopía de emisión atómica de plasma por acoplamiento
inductivo (ICP‐OES) ………………........................................................ 88
2.2.2.2. Difracción de Rayos X (XRD) ..................................................... 88
2.2.2.3. Espectroscopía Raman ………..................................................... 90
2.2.2.4. Análisis termogravimétrico y térmico diferencial (TGA‐DTA) .. 91
2.2.2.5. Microscopía electrónica de transmisión (TEM) ………..………. 92
2.2.2.6. Microscopía electrónica de barrido (SEM) ………….....………. 92
2.2.2.7. Isoterma de adsorción‐desorción de nitrógeno …………………. 93
2.2.2.8. Determinación del pH de potencial de carga cero (pHPZC) por
valoración másica ……………….............................................................. 95
2.2.2.9. Espectroscopía fotoelectrónica de Rayos X …………..………..... 95
2.2.2.10. Espectroscopía ultravioleta‐visible de reflectancia difusa
(DR‐UV‐Vis) ………………..................................................................... 97
2.3. REACTIVOS EMPLEADOS ………...……………..………………..…... 98
BIBLIOGRAFÍA ………...……………..……………………………….……... 103
CAPÍTULO 3 (CHAPTER 3). PAPER 1: ON OZONE‐PHOTOCATALYSIS
SYNERGISM IN BLACK‐LIGHT INDUCED REACTIONS: OXIDIZING
SPECIES PRODUCTION IN PHOTOCATALYTIC OZONATION
VERSUS HETEROGENEOUS PHOTOCATALYSIS ........................................ 107
3.1. INTRODUCTION …………………………………………………….….. 109
3.2. EXPERIMENTAL SECTION ……………..………………………….….. 113
3.2.1. Experimental set‐up and oxidation/ozonation procedure ..…..... 113
3.2.2. Photon fluxes determination ………………………………...…..... 114
3.3. RESULTS AND DISCUSSION ….………..………………………….….. 115
3.3.1. Absorbed photon flux ..…………………………………………..... 115
Índice
IV
3.3.2. Photocatalytic oxidation versus photocatalytic ozonation …...... 121
3.3.2.1. Influence of ozone concentration ………...................................... 125
3.3.2.2. Influence of pH …………………………………………………. 126
3.3.2.3. Synergistic effect and quantum yield of photocatalytic induced
reactions ………………………………………........................................ 129
3.3.2.4. Simplified economic considerations ……………………...…..… 132
3.4. CONCLUSIONS ………………....………..………………………….….. 135
AKNOWLEDGEMENTS ………...………...………………………….……... 136
REFERENCES ……...…...……………..……………………………….……... 136
CAPÍTULO 4 (CHAPTER 4). PAPER 2: INFLUENCE OF STRUCTURAL
PROPERTIES ON THE ACTIVITY OF WO3 CATALYSTS FOR VISIBLE
LIGHT PHOTOCATALYTIC OZONATION …................................................. 141
4.1. INTRODUCTION …………………………………………………….….. 143
4.2. EXPERIMENTAL SECTION ……………..………………………….….. 144
4.2.1. Catalysts preparation ………………………………………...…..... 144
4.2.2. Characterization ………………….…………………………...…..... 145
4.2.3. Catalytic activity measurements ………………..…………...…..... 146
4.3. RESULTS AND DISCUSSION ….………..………………………….….. 148
4.3.1. Characterization of the photocatalysts ..………………………..... 148
4.3.2. Comparison of processes ………………………………………...... 154
4.3.2.1. Simplified mechanistic approach of IBP photocatalytic
ozonation ……………...………………………........................................ 160
4.3.3. Catalysts screening for photocatalytic ozonation under
visible‐light radiation …………………………………………………...... 164
4.3.4. Simulated solar light photocatalytic ozonation of ECs in
MWW ………………………...…………………………………………...... 168
4.4. CONCLUSIONS ………………....………..………………………….….. 172
AKNOWLEDGEMENTS ………...………...………………………….……... 172
REFERENCES ……………..……...………...………………………….……... 173
SUPPLEMENTARY INFORMATION OF CHAPTER 4 ………..….……... 177
Índice
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CAPÍTULO 5 (CHAPTER 5). PAPER 3: VISIBLE LIGHT PHOTOCATALYTIC
OZONATION OF DEET IN THE PRESENCE OF DIFFERENT FORMS OF
WO3 ………................................................................................................................. 183
5.1. INTRODUCTION …………………………………………………….….. 185
5.2. EXPERIMENTAL SECTION ……………..………………………….….. 186
5.2.1. Catalysts preparation ………………………………………...…..... 186
5.2.2. Characterization ………………….…………………………...…..... 188
5.2.3. Catalytic activity measurements ………………..…………...…..... 188
5.3. RESULTS AND DISCUSSION ….………..………………………….….. 189
5.3.1. Photocatalysts characterization ………..………………………..... 189
5.3.2. Catalytic activity ……………..…………………………………...... 195
5.4. CONCLUSIONS ………………....………..………………………….….. 202
AKNOWLEDGEMENTS ………...………...………………………….……... 202
REFERENCES ……………..……...………...………………………….……... 202
SUPPLEMENTARY INFORMATION OF CHAPTER 5 ………..….……... 205
CAPÍTULO 6 (CHAPTER 6). PAPER 4: NANOSTRUCTURED CeO2 AS
CATALYSTS FOR DIFFERENT AOPs BASED IN THE APPLICATION OF
OZONE AND SIMULATED SOLAR RADIATION.......................................... 213
6.1. INTRODUCTION …………………………………………………….….. 215
6.2. EXPERIMENTAL SECTION ……………..………………………….….. 216
6.2.1. Catalysts preparation ………………………………………...…..... 216
6.2.2. Characterization ………………….…………………………...…..... 217
6.2.3. Catalytic activity measurements ………………..…………...…..... 217
6.3. RESULTS AND DISCUSSION ….………..………………………….….. 219
6.3.1. Photocatalysts characterization ………..………………………..... 219
6.3.2. Catalytic activity ……………..…………………………………...... 223
6.3.3. Considerations on the reaction mechanism of DEET
photocatalytic ozonation with CeO2 catalysts and solar radiation …... 227
6.4. CONCLUSIONS ………………....………..………………………….….. 230
AKNOWLEDGEMENTS ………...………...………………………….……... 230
REFERENCES ………………………………………………..……..….……... 231
Índice
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CAPÍTULO 7 (CHAPTER 7). PAPER 5: WO3‐TiO2 BASED CATALYSTS FOR
THE SIMULATED SOLAR RADIATION ASSISTED PHOTOCATALYTIC
OZONATION OF EMERGING CONTAMINANTS IN A MUNICIPAL
WASTEWATER TREATMENT PLANT EFFLUENT ….................................... 235
7.1. INTRODUCTION …………………………………………………….….. 237
7.2. EXPERIMENTAL SECTION ……………..………………………….….. 238
7.2.1. Catalysts preparation ………………………………………...…..... 238
7.2.2. Characterization ………………….…………………………...…..... 239
7.2.3. Catalytic activity measurements ………………..…………...…..... 240
7.3. RESULTS AND DISCUSSION ….………..………………………….….. 243
7.3.1. Characterization of the photocatalysts ………..………………..... 243
7.3.2. Visible light response of the photocatalysts …………………...... 251
7.3.3. Photocatalytic degradation of ECs in MWW ………………......... 253
7.4. CONCLUSIONS ………………....………..………………………….….. 263
AKNOWLEDGEMENTS ………...………...………………………….……... 263
REFERENCES ………………………………………………..……..….……... 264
CAPÍTULO 8 (CHAPTER 8). PAPER 6: REACTION MECHANISM AND
KINETICS OF DEET VISIBLE LIGHT ASSISTED PHOTOCATALYTIC
OZONATION WITH WO3 CATALYST ……….................................................. 269
8.1. INTRODUCTION …………………………………………………….….. 271
8.2. EXPERIMENTAL SECTION ……………..………………………….….. 272
8.2.1. Experiments …………………………………………………...…..... 272
8.2.2. Analytical methods ………...…….…………………………...…..... 274
8.3. RESULTS AND DISCUSSION ….………..………………………….….. 276
8.3.1. Comparison of processes ………..……………………...………..... 276
8.3.2. Determination of the main species responsible for DEET
degradation and mineralization ………………………..……………...... 283
8.3.3. Identification of TPs ……………………………………………....... 287
8.3.4. Proposed reaction mechanism ……………………..…………....... 294
8.3.5. Kinetic study ……………………………………………………....... 299
8.4. CONCLUSIONS ………………....………..………………………….….. 309
AKNOWLEDGEMENTS ………...………...………………………….……... 310
Índice
VII
REFERENCES ………………………………………………..……..….……... 310
CAPÍTULO 9: CONCLUSIONES …………………………………………….…. 315
NOMENCLATURA …………………………………………...……………….….. 321
NOMENCLATURE …………………………………………………………….….. 325
RESUMEN
La Tesis Doctoral que se presenta en esta memoria se encuadra dentro del
trabajo desarrollado por el grupo de investigación TRATAGUAS, perteneciente
al Área de Ingeniería Química de la Universidad de Extremadura. En los últimos
años, la actividad del grupo se ha centrado en la eliminación de contaminantes
emergentes presentes en las aguas mediante procesos avanzados de oxidación y,
más recientemente, en la preparación de catalizadores para su aplicación en
dichos procesos. En concreto, en esta Tesis se han abordado fundamentalmente
algunos de los objetivos del proyecto de investigación titulado “Preparación de
catalizadores y su aplicación en la eliminación de contaminantes refractarios de aguas
residuales mediante ozonización fotocatalítica” (referencia CTQ2012/35789/C02/01),
financiado por el Ministerio de Economía, Industria y Competitividad y el Fondo
Europeo de Desarrollo Regional (FEDER).
En el escenario actual donde la disponibilidad de agua de calidad se está
viendo disminuida, la aparición de los denominados “contaminantes
emergentes” y sus posibles riesgos tanto para la vida acuática como la humana
ha despertado un especial interés. A este grupo de contaminantes pertenecen
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compuestos como fármacos, productos de higiene y cuidado personal, drogas de
abuso, pesticidas, surfactantes o aditivos industriales, entre otros. Todos ellos se
producen a escala mundial, tienen multitud de aplicaciones y han llegado a ser
imprescindibles para la sociedad moderna. Así, su introducción en el medio es
continua, principalmente a través de la descarga de las estaciones depuradoras
de aguas residuales (EDAR), dado que muchos de estos contaminantes son
refractarios a los procesos convencionales existentes. Ante esta situación, se hace
necesaria la aplicación de un tratamiento terciario adicional que permita
degradar dichos compuestos antes del vertido del efluente a cauce público o de
su reutilización.
En este trabajo se han seleccionado varios contaminantes emergentes,
detectados con frecuencia en efluentes secundarios de EDAR, para estudiar la
aplicabilidad de procesos avanzados de oxidación (PAO) como tratamiento
terciario. Estos procesos se basan en la generación de radicales libres,
especialmente radicales hidroxilo ( •HO ), que son especies transitorias de vida
media muy corta y alto poder oxidante, capaces de reaccionar de forma no
selectiva con un amplio grupo de contaminantes resistentes a los tratamientos
de agua convencionales. Dentro de los PAO, la presente investigación se centra
en la aplicación de la fotocatálisis heterogénea combinada con ozono
(ozonización fotocatalítica) empleando distintos catalizadores y fuentes de
radiación. Según algunos estudios previos, se espera que en este tratamiento
combinado aumente el rendimiento de formación de especies oxidantes respecto
de los procesos individuales de fotocatálisis y ozonización, siendo por tanto más
eficaz en la eliminación de contaminantes en agua.
El punto de partida de este trabajo de Tesis ha sido justificar la elección del
proceso de ozonización fotocatalítica como tratamiento a aplicar. Para ello, se ha
seleccionado el metanol como compuesto modelo, empleando radiación
ultravioleta (UVA) y TiO2 comercial (P25) como catalizador. La baja reactividad
del metanol frente al ozono molecular y su alta afinidad por las especies
oxidantes generadas en los procesos fotocatalíticos permiten determinar la
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existencia o no de sinergia al combinar los procesos individuales de fotocatálisis
y ozonización, a partir del cálculo del rendimiento cuántico de los tratamientos
fotocatalíticos.
Para determinar dicha sinergia se siguió la evolución de la concentración de
formaldehído, principal producto de oxidación del metanol, en ensayos de
fotocatálisis y ozonización simple y fotocatalítica, evaluando el efecto de la
concentración de ozono y del pH. Para calcular el rendimiento cuántico de las
reacciones fotocatalíticas se determinó la intensidad de radiación absorbida por
el catalizador y se tuvieron en cuenta las distintas reacciones que pueden tener
lugar dependiendo del proceso y las condiciones del medio: reacción directa
ozono‐metanol, reacciones indirectas por descomposición de ozono en especies
radicales y reacciones fotocatalíticas. Esta última contribución es la que permite
calcular el rendimiento cuántico de foto‐generación de especies oxidantes. A pH
= 3 (pH al cual se inhiben las reacciones indirectas del ozono) la presencia de
ozono ejerció un efecto positivo en la velocidad de formación de especies
oxidantes debido a las reacciones inducidas por la luz, incrementándose el valor
del rendimiento cuántico con respecto al proceso de fotocatálisis. Este parámetro
aumentó aún más a pH = 7 al verse favorecidas las reacciones indirectas del
ozono. El efecto positivo de la presencia de ozono en las reacciones
fotocatalíticas se ha atribuido a su papel como aceptor de electrones, dando
lugar a una mayor concentración de radicales hidroxilo en el medio y
reduciendo en cierta medida el proceso de recombinación en la superficie del
catalizador.
Comprobada la mayor eficiencia del proceso fotocatalítico cuando se
combina con ozono, el resto del trabajo de esta Tesis Doctoral se ha centrado en
el empleo de fotocatalizadores que permitan un mayor aprovechamiento de la
radiación solar con respecto al fotocatalizador utilizado por excelencia (TiO2), y
la posterior aplicación de estos materiales en la eliminación de distintos
contaminantes emergentes mediante ozonización fotocatalítica solar. De este
modo, se intenta paliar uno de los principales inconvenientes de esta tecnología
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derivado del uso de fuentes artificiales de radiación. Todos los ensayos se han
realizado a escala de laboratorio, empleando luz solar simulada y distintos
materiales previamente sintetizados y/o comerciales.
Se han sintetizado catalizadores de WO3 con diferentes propiedades
estructurales, estudiando su actividad en el proceso de ozonización fotocatalítica
solar aplicado a la eliminación de: i) ibuprofeno (IBP) en agua ultrapura
empleando radiación solar restringida a la zona del visible; y ii) una mezcla de
10 contaminantes emergentes añadidos al efluente de una EDAR urbana,
empleando todo el espectro de la radiación solar. Estos catalizadores se
sintetizaron mediante descomposición térmica de un precursor de wolframio
(tungstita, H2O∙WO3) a 300 y 450 °C y diferentes tiempos de calcinación. El
catalizador sometido a descomposición térmica a 450 °C durante 5 minutos
presentó la estructura monoclínica del WO3 y una concentración mayor de
vacantes de oxígeno que el resto de catalizadores, lo que condujo a una mayor
actividad en el proceso de ozonización fotocatalítica. Bajo luz solar visible, la
desaparición de ibuprofeno (CIBP,0 = 10 mg L‐1) fue completa en 20 minutos y la
mineralización superior al 85 % tras dos horas de reacción en las condiciones de
operación empleadas (pH0 = 6,5, T = 20 ‐ 40 °C, V = 0,5 L, CWO3 = 0,25 g L‐1,
CO3,g = 10 mg L‐1, Qg = 20 L h‐1). Bajo radiación solar global, en las condiciones
anteriores este catalizador permitió la degradación completa de la mezcla de
contaminantes (C0 = 0,5 mg L‐1 cada uno) en el efluente de EDAR urbana
(CCOT,0 = 17 mg L‐1) en 60 minutos y una mineralización del 40 % en 2 horas.
Por otro lado, se ha evaluado la influencia de la morfología y de nuevo la
estructura del WO3 en el proceso de ozonización fotocatalítica bajo luz solar
visible, empleando N,N‐dietil‐meta‐toluamida (DEET) como compuesto modelo
en agua ultrapura. Los catalizadores de WO3 se obtuvieron mediante síntesis
sol‐gel e hidrotermal, procesos que favorecieron el desarrollo de diferentes
formas y estructuras cristalinas.
Los resultados indican que el desarrollo de las estructuras cristalinas
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monoclínica y ortorrómbica del WO3 incrementa la actividad fotocatalítica del
material en mayor medida que un área superficial más desarrollada. Asimismo,
en los procesos que emplean ozono la presencia de estados reducidos de W
condujo a una mayor actividad fotocatalítica. En las condiciones de operación
estudiadas (CDEET,0 = 5 mg L‐1, pH0 = 6, T = 20 ‐ 40 °C, V = 0,5 L, CWO3 = 0,25 g L‐1,
CO3,g = 10 mg L‐1, Qg = 15 L h‐1), los mejores catalizadores condujeron a la
degradación total de DEET en menos de 20 minutos y una mineralización
superior al 70 % tras 2 horas de tratamiento.
Otra alternativa al TiO2 fueron los catalizadores de CeO2. Se ha estudiado la
influencia de la morfología del material en su actividad evaluando la misma a
través de ensayos de degradación de DEET en agua ultrapura mediante
ozonización fotocatalítica bajo luz solar visible y radiación solar global. Se
prepararon dos catalizadores nanoestructurados de CeO2 (nanocubos y
nanovarillas) mediante síntesis hidrotermal. En las condiciones de operación
ensayadas (CDEET,0 = 5 mg L‐1, pH0 = 6, T = 35 ‐ 40 °C, V = 0,5 L, CWO3 = 0,25 g L‐1,
CO3,g = 10 mg L‐1, Qg = 15 L h‐1), ambos catalizadores resultaron activos en la
eliminación de DEET mediante ozonización fotocatalítica bajo radiación tanto
visible como solar. Por su menor energía de salto de banda fueron las
nanovarillas las que presentaron una mayor actividad bajo luz visible, logrando
la eliminación total de DEET en 45 minutos y una mineralización del 68 % tras 2
horas de tratamiento. Por el contrario, bajo radiación solar global los nanocubos
tuvieron un mejor rendimiento (degradación total de DEET en 30 minutos y 80
% de mineralización tras 2 horas de reacción), presentando esta morfología,
además, una mayor velocidad específica de mineralización debido a la mayor
actividad de las caras cristalinas expuestas.
Finalmente, se han preparado catalizadores compuestos de WO3‐TiO2 con un
contenido en WO3 de aproximadamente un 4 % en peso empleando dos soportes
de TiO2 diferentes: TiO2 comercial (P25) y nanotubos de TiO2 sintetizados
mediante tratamiento hidrotermal del catalizador comercial. Una vez
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preparados, los materiales han sido empleados en la ozonización fotocatalítica
del efluente de una EDAR urbana dopado con una mezcla de contaminantes
bajo radiación solar global.
El catalizador basado en nanotubos presentó una porosidad más
desarrollada y una mayor eficiencia en el proceso de ozonización fotocatalítica.
Así, en las condiciones de operación estudiadas (C0 = 2 mg L‐1 de cada
contaminante emergente, CCOT,0 = 35 mg L‐1, CCI,0 = 42 mg L‐1, pH0 = 8.3, T = 20 ‐ 40
°C, V = 0,5 L, CNT‐WO3 = 0,5 g L‐1, CO3,g = 10 mg L‐1, Qg = 20 L h‐1), este material
condujo a la total desaparición de los contaminantes emergentes en menos de 40
minutos y a un grado de mineralización superior al 64 % tras 2 horas. La mayor
actividad fotocatalítica del material compuesto en comparación con el
catalizador de TiO2 puro se debe a varios efectos combinados tales como: el
mayor aprovechamiento de la radiación visible; una mayor capacidad de
adsorción de compuestos orgánicos; y una eficiente descomposición del ozono
gracias a la presencia de WO3, el cual resulta activo en la descomposición del
ozono en la oscuridad (ozonización catalítica).
Tras la síntesis, caracterización y aplicación de los distintos materiales, en
base a los resultados obtenidos bajo luz visible se ha seleccionado, por su mayor
actividad, el WO3 con estructura monoclínica y morfología de microesferas, y se
ha aplicado en la ozonización fotocatalítica con luz visible solar de DEET en
disolución acuosa, al objeto de proponer un modelo cinético que describa el
comportamiento de este sistema con vistas a su posible desarrollo y aplicación a
mayor escala.
Mediante el empleo de inhibidores (ter‐butanol y ácido oxálico), se
analizaron las principales especies involucradas en el proceso resultando ser los
radicales •HO los principales responsables de la degradación y mineralización
del DEET. Se identificaron diferentes productos intermedios de degradación de
DEET y se estableció la evolución de sus abundancias relativas a lo largo del
tiempo de reacción. La eficacia del tratamiento de ozonización fotocatalítica se
Resumen
7
manifiesta no solo en la velocidad de eliminación del DEET, sino también en la
de desaparición de los intermedios de degradación y de mineralización. En las
condiciones de operación ensayadas (CDEET,0 = 15 mg L‐1, pH0 = 6, T = 20 ‐ 40 °C,
V = 0,5 L, CWO3 = 0,25 g L‐1, CO3,g = 10 mg L‐1, Qg = 15 L h‐1), los intermedios
formados al inicio, de estructura similar a la del compuesto de partida, fueron
eliminados por completo tras 60 minutos de operación, permaneciendo en
disolución únicamente ácidos orgánicos de cadena corta de muy baja toxicidad y
a concentraciones en concordancia con el grado de mineralización alcanzado (60
% en 2 horas). En base a los resultados obtenidos se ha propuesto para la
ozonización fotocatalítica bajo luz visible un mecanismo de reacción que incluye
distintas etapas basadas en el ataque del radical •HO , desarrollando un modelo
cinético capaz de simular la evolución de la concentración de DEET, la de sus
principales intermedios de degradación y la de los ácidos orgánicos de cadena
corta en términos de carbono orgánico total, lo que proporciona un enfoque
simplificado del proceso.
CAPÍTULO 1 Introducción y objetivos
El crecimiento demográfico e industrial en las últimas décadas ha provocado un incremento exponencial de la contaminación de los recursos hídricos, lo que motiva el desarrollo de procesos y tecnologías que permitan devolver los parámetros físicos, químicos y biológicos de las aguas residuales a su estado natural. En este capítulo se aborda brevemente la naturaleza de la contaminación hídrica, con especial atención a los denominados contaminantes emergentes. Seguidamente, se exponen algunos conceptos teóricos sobre procesos avanzados de oxidación que se proponen para eliminar dichos contaminantes del agua, haciendo referencia a líneas y resultados actuales de investigación. Por último, se indican los objetivos que se persiguen en este trabajo de Tesis, así como la estructura de la memoria que se presenta.
Introducción y objetivos
11
1.1. ANTECEDENTES Y JUSTIFICACIÓN DE LA TESIS
1.1.1. Contextualización del trabajo
El presente trabajo de Tesis Doctoral se ha desarrollado dentro del grupo de
investigación TRATAGUAS del Departamento de Ingeniería Química y Química
Física de la Universidad de Extremadura [1], gracias a la beca de Formación de
Personal Investigador (Ref. PD12059) concedida por la Consejería de Empleo,
Empresa e Innovación (Junta de Extremadura) y el Fondo Social Europeo.
En los últimos años, la actividad del grupo se ha centrado en la eliminación
de contaminantes emergentes presentes en las aguas mediante procesos
avanzados de oxidación y, más recientemente, en la preparación de
catalizadores para su aplicación en dichos procesos. El desarrollo de las líneas
de investigación anteriores ha sido posible gracias a la financiación recibida del
Ministerio de Economía, Industria y Competitividad y del Fondo Europeo de Desarrollo
Regional (FEDER) a través de los proyectos del Programa Nacional de
Investigación “Integración de procesos de fotocatálisis solar en la depuración biológica
de aguas residuales para la eliminación de contaminantes emergentes”
(CTQ2009/13459/C05/05); “Preparación de catalizadores y su aplicación en la
eliminación de contaminantes refractarios de aguas residuales mediante ozonización
fotocatalítica” (CTQ2012/35789/C02/01); y “Led y fotocatalizadores polifuncionales
basados en grafeno y estructuras metal‐orgánicas para el tratamiento de aguas por
ozonización fotocatalítica” (CTQ2015‐64944‐R). De entre los anteriores, el trabajo
de Tesis Doctoral presentado en esta memoria se encuadra fundamentalmente
dentro del proyecto CTQ2012/35789/C02/01, si bien se han abordado algunos
objetivos de los otros dos proyectos citados.
1.1.2. El agua en el mundo actual
El agua es primordial para la salud y la supervivencia de los seres vivos.
Estos necesitan el agua para realizar sus funciones vitales y sin ella no habría
vida. El agua representa entre el 50 y el 90 % de la masa de los seres vivos,
CAPÍTULO 1
12
siendo un 75 % en el caso de los seres humanos. Además, el agua promueve el
crecimiento económico y el desarrollo social de una región, de manera que la
relativa abundancia y seguridad del suministro hídrico en los países
industrializados ha sido, en gran medida, el factor que ha facilitado su
desarrollo económico [2]. A su vez, el aumento en el nivel de vida lleva
aparejado un mayor consumo de agua, por lo que ha dejado de ser un recurso
abundante y disponible en cantidad y calidad suficiente para sus diversos usos.
De toda el agua existente en la Tierra solo una pequeña parte está disponible
para su empleo y consumo, siendo, por tanto, un recurso limitado. La gran
mayoría del agua en el planeta, casi el 97,5 %, se encuentra en los mares y los
océanos y, dada su alta salinidad, no puede aprovecharse de forma directa como
recurso para la población. Del total de agua dulce (2,5 %), más de dos tercios se
encuentra acumulada en los polos. Así pues, solo una pequeña fracción, inferior
al 1 % del total del agua en la Tierra, se encuentra accesible en los ríos o en los
acuíferos, tal y como muestra la Figura 1.1 [3].
Figura 1.1. Distribución de los recursos hídricos en la Tierra [3].
Introducción y objetivos
13
La distribución de los recursos hídricos se basa en el ciclo del agua, en el que
el océano es el origen de la mayor parte de las precipitaciones del planeta. La
lluvia sobre la tierra satisface casi todas las necesidades de agua dulce de la
población, junto a una pequeña cantidad, continuamente en aumento, de agua
proveniente de la desalación [4,5]. No obstante, la disponibilidad de agua de
calidad no es la misma en todas partes; su distribución no es homogénea así
como tampoco lo es el sistema de lluvias en la Tierra. De hecho, existen países
húmedos y con escasa población que disponen de mayor cantidad de agua per
cápita que otros áridos y más populosos, algunos de los cuales sufren graves
problemas de escasez del recurso [6] (ver Figura 1.2). Esto se ve agravado hoy en
día por efecto del calentamiento global, el cual se prevé que dará lugar a graves
inundaciones y sequías en amplias zonas del planeta [7].
Figura 1.2. Porcentaje de agua e índice de población de cada continente sobre el total
mundial según datos de la UNESCO y el Departamento de Asuntos Económicos y
Sociales de las Naciones Unidas [6].
A la distribución heterogénea de los recursos hídricos y las consecuencias del
cambio climático se suma el aumento de la población mundial a un ritmo de 80
millones de personas al año, lo que ha triplicado las extracciones de agua en los
CAPÍTULO 1
14
últimos 50 años según cifras del Programa Mundial de Evaluación de los
Recursos Hídricos de las Naciones Unidas [8]. Si las tendencias actuales de
demografía y consumo continúan, se prevé que en 2025 dos de cada tres
personas en el mundo vivirán en zonas con estrés hídrico [8,9]. Además, los
mayores crecimientos de población estarán localizados en países en desarrollo,
tratándose en muchos casos de regiones sin acceso a agua potable ni a un
saneamiento adecuado.
De forma paralela, los diversos usos que se han dado al agua, muy
influenciados por el aumento de la densidad poblacional y los modelos de
consumo actuales, han provocado una disminución en la calidad del agua dulce
disponible. Hasta hace relativamente pocos años, gracias a su conocida
capacidad de auto‐depuración, los ríos han sido con frecuencia los receptores de
las aguas residuales generadas por el hombre. En muchos casos, los vertidos han
sobrepasado con creces la capacidad auto‐depuradora del agua, provocando una
contaminación que altera sus características y hace inservible este recurso
natural para determinados usos. El principal problema lo constituyen los
vertidos residuales procedentes de la industria, la agricultura y la ganadería,
aunque los desechos domésticos también juegan un papel destacado en la
contaminación del medio ambiente acuático [10].
En cifras reales, el 70 % del consumo total de agua en el mundo se destina a
uso agrícola y ganadero, el 22 % a uso industrial y un 8 % al gasto doméstico,
según se constata en la edición del año 2016 del Informe de las Naciones Unidas
sobre el desarrollo de los recursos hídricos en el mundo [10]. El sector productor
no sólo es el que más consume, también es el que más contamina. La necesidad
de alimentos ha llevado a la expansión del riego y a una utilización cada vez
mayor de fertilizantes y plaguicidas con el fin de lograr y mantener
rendimientos superiores. Estas sustancias son arrastradas por el agua de lluvia y
de riego, movilizando consigo desde compuestos de nitrógeno, fósforo y azufre
hasta trazas de otros compuestos tóxicos, como por ejemplo los organoclorados.
El gran contenido en nutrientes de estas escorrentías causa la eutrofización en
Introducción y objetivos
15
aguas superficiales y la contaminación de aguas subterráneas por infiltración en
el subsuelo [11].
Por su parte, el sector industrial produce un impacto ambiental aún mayor
debido a la introducción de una gran variedad de sustancias que pueden
contaminar el medio hídrico aportando materia orgánica, metales pesados,
radiactividad, aceites, grasas, ácidos, bases, temperatura extrema, etc. Entre las
industrias más contaminantes destacan las petroquímicas, energéticas,
papeleras, metalúrgicas, alimentarias, textiles y mineras [12].
En cuanto a la contaminación de origen doméstico, las aguas residuales que
se generan constituyen una fuente de patógenos microbianos, sólidos en
suspensión, nutrientes como el nitrógeno y el fósforo y materia orgánica, tanto
coloidal como disuelta, que consumirá oxígeno [13]. Además, cada vez es más
frecuente la aparición en estas aguas de productos químicos presentes en la
composición de fármacos, cosméticos o productos de limpieza [14].
Aparte de los anteriores, existen otros sectores fuertemente dependientes del
agua y, por tanto, fuentes de contaminación de la misma, entre los que se
encuentran la silvicultura, la pesca y la acuicultura continental, la eliminación de
desechos y la minería y la extracción de recursos [10].
En conclusión, el desarrollo demográfico e industrial de las últimas décadas
ha incrementado el uso indiscriminado de los recursos hídricos hasta niveles
insostenibles. En este escenario, surge la necesidad de tomar medidas
legislativas para proteger las aguas tanto en términos cualitativos como
cuantitativos y garantizar así su sostenibilidad. Esta legislación, cada vez más
estricta, obliga a asumir una idea clave para un nuevo enfoque de la gestión
hídrica: considerar el problema de forma global, esto es, el “ciclo integral del
agua”. De esta forma, la captación ya no es algo separado de la potabilización y
la depuración, de manera que si el agua residual se trata y se purifica
suficientemente se convierte en agua regenerada, que puede y debe ser de nuevo
utilizada regresando de este modo al punto inicial del ciclo. No se trata ya de
CAPÍTULO 1
16
buscar nuevas fuentes de agua, sino de recuperar la que se malgasta, regenerar
la que se contamina y reutilizarla [15].
1.1.2.1. Marco legal del agua
La Directiva Marco Europea del Agua (DMA) nace como respuesta a la
necesidad de unificar las actuaciones en materia de gestión de agua en la Unión
Europea. La DMA es el fruto de un proceso extenso de discusión, debate y
puesta en común de ideas entre un amplio abanico de expertos, usuarios del
agua, medioambientalistas y políticos que, por consenso, sentaron los principios
fundamentales de la gestión moderna de los recursos hídricos y que constituyen
hoy por hoy los cimientos de esta directiva. Tras un largo periodo de gestación
de más de cinco años entró en vigor el 22 de diciembre de 2000 (2000/60/CE)
[16], siendo posteriormente modificada por la Directiva 2008/105/CE [17] y, más
recientemente, por la Directiva 2013/39/UE [18].
Esta legislación europea persigue como objetivo último conseguir restablecer
las características naturales de las masas de agua, eliminando de ellas, entre
otras, las sustancias consideradas prioritarias y peligrosas. Para ello, se tienen en
cuenta aspectos como el principio de acción preventiva o el control de la
contaminación en su origen, mediante la fijación de valores límite de emisión y
de normas de calidad ambiental (NCA), entre otros. Las sustancias prioritarias
están definidas por la Directiva 2000/60/CE [16]. A través del artículo 16 de la
misma, se exige el establecimiento de un listado de estas sustancias en base a su
capacidad para representar un riesgo significativo para o a través del medio
acuático. De esta forma, la Decisión 2455/2001/CE [19] estableció la primera lista
de sustancias prioritarias, que debía ser sometida a revisión cada 4 años y
actualizada en función de las nuevas sustancias que pudieran ser identificadas
como prioritarias. Así, el anexo II de la Directiva 2008/105/CE presenta una lista
de 33 sustancias [17] entre las que figuran los metales cadmio, plomo, mercurio
y níquel y sus compuestos, el benceno, los hidrocarburos aromáticos policíclicos
y varios pesticidas [20]. De estas sustancias, 11 fueron identificadas como
peligrosas y, por tanto, sujetas a cese o eliminación progresiva de sus emisiones,
Introducción y objetivos
17
descargas y pérdidas en un plazo de tiempo no superior a 20 años. Además, en
el Anexo I de esta directiva también se establecen las NCA relativas a valores
máximos de concentración media y de concentración máxima admisibles de
cada sustancia, debiendo velar los Estados miembros porque se cumplan estos
estándares de calidad [17].
Con la publicación de la Directiva modificativa 2013/39/UE, a la lista inicial se
sumaron 12 sustancias más, contando por tanto con un total de 45 sustancias
prioritarias de las cuales 21 son identificadas como peligrosas. Además, en base
a los últimos conocimientos técnicos y científicos sobre sus propiedades, la
directiva actualiza las NCA de 7 de las 33 sustancias prioritarias originales [18].
Desde el 22 de diciembre de 2015 los planes hidrológicos deben tener en cuenta
estas NCA revisadas con el objeto de alcanzar un buen estado químico de las
aguas superficiales en relación con estas sustancias, a más tardar el 22 de
diciembre de 2021. Asimismo, las 12 sustancias prioritarias más recientes y sus
NCA deberán tenerse en cuenta en la elaboración de los programas de
seguimiento suplementarios y en programas preliminares de medidas a
presentar antes de que finalice 2018, con el fin de lograr un buen estado químico
de las aguas superficiales en relación con estas nuevas sustancias a más tardar el
22 de diciembre de 2027.
Para cada demarcación hidrográfica, los Estados miembros deben elaborar
un inventario de emisiones, vertidos y pérdidas de las sustancias identificadas
por esta directiva. En España, esto se gestiona a través de programas de control
de Sustancias Peligrosas, dependientes del Ministerio de Agricultura y Pesca,
Alimentación y Medioambiente [21].
Por otra parte, la Directiva 2013/39/UE introduce una estipulación según la
cual debe establecerse una lista de observación de sustancias sobre las que han
de recabarse datos de seguimiento a nivel de la Unión Europea, para que sirvan
de base a futuros ejercicios de asignación de prioridad de conformidad [18]. En
este sentido, la Decisión 495/2015/UE [22] estableció una primera lista de
CAPÍTULO 1
18
observación que contenía 10 sustancias y que deberá actualizarse cada dos años.
Se tienen en cuenta, en particular, aquellas sustancias que estuvieron a punto de
ser consideradas prioritarias en la última revisión pero para las que los datos de
seguimiento siguen siendo insuficientes como para confirmar un riesgo
significativo. También se examinaron otras sustancias nuevas sobre las que no se
disponía de datos de seguimiento recientes o estos eran insuficientes. El riesgo
que supone cada una de esas sustancias se determina a partir de la información
disponible sobre su peligrosidad intrínseca y la exposición del medioambiente a
ellas. Por su parte, dicha exposición se estima a partir de los datos sobre el
alcance de la producción y utilización, teniendo en cuenta datos de seguimiento
reales.
Cabe destacar la incorporación en la primera lista de observación de los
fármacos diclofenac, 17‐‐estradiol y 17‐‐etinilestradiol. Los fármacos se
encuentran entre las sustancias conocidas como “contaminantes emergentes o
contaminantes de preocupación emergente”, junto a otros compuestos
empleados en la industria y en la vida diaria tales como productos de higiene y
cuidado personal, drogas de abuso, los metabolitos y/o productos de
degradación de las sustancias anteriores, surfactantes, aditivos industriales,
pesticidas, retardantes de llama, esteroides y hormonas [23].
La aparición de contaminantes emergentes en las aguas se justifica en base al
cambio en los hábitos de consumo de la sociedad actual (mayor densidad y
envejecimiento de la población, mayor cuidado de la imagen exterior, aumento
de enfermedades psicológicas o de cáncer, etc.), y su presencia se ha hecho
evidente gracias al avance de las técnicas analíticas [24]. Se producen a escala
mundial, tienen multitud de aplicaciones y han llegado a ser imprescindibles
para la sociedad moderna. En la mayoría de los casos, estos compuestos son
introducidos en el medio hídrico a través de las aguas residuales urbanas,
aunque también pueden proceder de las aguas empleadas en actividades
industriales o agrícolas y ganaderas que, como puede observarse en la Figura
1.3, se descargan en las estaciones depuradoras de aguas residuales (EDAR) o
Introducción y objetivos
19
directamente a los ríos, receptores a su vez de los efluentes de las EDAR.
También pueden introducirse en las aguas subterráneas a través de la lixiviación
de los vertidos o de las filtraciones de las aguas superficiales. Así, se ha
identificado la presencia de un amplio número de contaminantes emergentes en
aguas superficiales, en efluentes de estaciones depuradoras de aguas residuales
[25‐28], e incluso en aguas potables de consumo humano [29‐31], claro indicador
de que los tratamientos convencionales resultan ineficaces en la eliminación de
muchos de estos compuestos [14,32‐34]. Por este motivo, si el agua regenerada
contiene estos compuestos y es utilizada en la irrigación, algunos de los
contaminantes pueden adsorberse en el suelo e incluso llegar a contaminar
aguas subterráneas [35].
Figura 1.3. Introducción de los contaminantes emergentes en el medio hídrico (adaptada
de [36]).
Si bien la concentración de estos compuestos en las aguas es baja (del orden
de ng L‐1), la exposición crónica debida a su introducción continuada puede
representar un nuevo problema medioambiental [37]. Aunque la presencia de la
mayoría de ellos en el medioambiente aún no se analiza de forma rutinaria, su
toxicidad frente a distintos organismos y microorganismos sí ha sido objeto de
CAPÍTULO 1
20
estudio en los últimos años, concluyéndose que la exposición crónica puede
causar alteraciones en la vida acuática, afectando al comportamiento y
reproducción de las especies, a la composición de las comunidades biológicas,
etc. [31,38]. Más aún, algunos trabajos revelan efectos dañinos en el crecimiento,
proliferación o material genético de células humanas [39,40]. El efecto de estos
contaminantes depende no solo de su concentración en el medio hídrico, sino
también de factores tales como su solubilidad en grasas, su persistencia,
bioacumulación, tiempo de exposición o mecanismos de biotransformación y
eliminación [20]. Así, aunque se encuentren en concentraciones muy bajas, la
exposición crónica y la acción sinérgica de estos compuestos o sus metabolitos
con otros contaminantes, unido en muchos casos a su baja biodegradabilidad y a
su posible bioacumulación en la cadena trófica, podría constituir a largo plazo
una amenaza potencial para los ecosistemas acuáticos y terrestres [14,29,41].
Cada vez con más asiduidad, los medios de comunicación se hacen eco de los
resultados de los trabajos realizados por la comunidad científica sobre la
presencia de este tipo de contaminantes en las aguas y sus posibles efectos. En
2005 la prensa española sacó a la luz pública un estudio en el que se ponían de
manifiesto fenómenos de feminización en peces en la cuenca del río LLobregat
como resultado de la presencia de compuestos con actividad estrogénica,
capaces de actuar como alteradores endocrinos [42]. En 2008 los medios
informaban sobre la presencia de medicamentos en el agua potable de un gran
número de ciudades de Estados Unidos [43]. En base a los estudios realizados
por investigadores de distintas instituciones españolas, en 2009 varios medios
alertaban sobre la presencia de drogas y sus metabolitos en el Ebro [44]. En 2012,
la prensa española recogía los resultados obtenidos por investigadores de la
Facultad de Ciencias de la Salud de la Universidad Rey Juan Carlos sobre la
presencia de fármacos en el agua embotellada, según los cuales 5 de las 10
marcas comerciales evaluadas presentaron nicotina en concentraciones de 7 a 15
ng L‐1 [45]. Más recientemente, también en nuestro país los medios
comunicación se han hecho eco del estudio realizado por expertos del Instituto
de Diagnóstico Ambiental y Estudios del Agua (IDAEA), sobre la acumulación
Introducción y objetivos
21
de retardantes de llama en delfines que habitan en el Golfo de Cádiz y el
estrecho de Gibraltar, productos capaces de alterar la función de las hormonas
de los cetáceos [46].
La toma de conciencia del riesgo que ocasiona la presencia de este tipo de
contaminantes en el medio ambiente es relativamente reciente. Teniendo en
cuenta que según las tendencias de consumo actuales, ni la producción y uso de
estos compuestos ni la consecuente continua introducción de los mismos en el
medio ambiente van a reducirse, se requiere una investigación más urgente. Es
necesario obtener datos suficientes para valorar de forma apropiada el impacto
de los contaminantes emergentes, determinar si pueden considerarse como
sustancias peligrosas y, en ese caso, establecer niveles máximos para su
concentración en las aguas. Actualmente, este estudio se encuentra entre las
líneas de investigación prioritarias de los principales organismos dedicados a la
protección de la salud pública y medioambiental, tales como la Organización
Mundial de la Salud (OMS), la Agencia de los Estados Unidos para la Protección
del Medioambiente (US‐EPA), o la Comisión Europea.
1.1.2.2. Conservación de los recursos hídricos
Continuamente aparecen nuevos contaminantes potencialmente peligrosos y
la demanda de agua no deja de ir en aumento. Al tratarse de una problemática
directamente relacionada con el desarrollo y la salud pública, se hace necesario
establecer una serie de pautas y acciones en materia de depuración de las aguas
residuales que permita, no solo cumplir con la normativa, sino también
implantar un ciclo de reutilización de las aguas tratadas.
La Directiva 91/271/CEE de 21 de Mayo de 1991 [47], por la que se regula el
tratamiento de las aguas residuales urbanas antes de su vertido, establecía que a
partir del año 2005 todos los núcleos urbanos de más de 2000 habitantes debían
disponer de instalaciones para la recogida y el tratamiento de las aguas
residuales. Posteriormente, en el año 2000, la Directiva Marco del Agua vino a
exigir no solo la depuración de las aguas residuales, sino conseguir que el grado
CAPÍTULO 1
22
de depuración cumpliera los parámetros establecidos para asegurar el buen
estado ecológico de todas las masas de agua, fijando finales del año 2015 como
fecha límite de cumplimiento [16]. Esto implicaba que ya no era suficiente que
una población contara con depuradora, sino que esta tenía que funcionar
adecuadamente.
En España, en respuesta a la Directiva 91/271/CEE sobre el tratamiento de las
aguas residuales, se elaboró el Plan Nacional de Calidad de las Aguas:
Saneamiento y Depuración (PNCA) [48], con un escenario temporal
comprendido entre 2007 y 2015, haciendo coincidir la fecha límite con la de la
Directiva Marco del Agua para así cumplir con la política europea en materia de
medioambiente hídrico. Entre las principales líneas de actuación, el Plan
contemplaba, además de la construcción, explotación y mantenimiento de las
depuradoras, el acondicionamiento de los sistemas de depuración a una
reducción eficaz de nutrientes mediante tratamiento terciario en las zonas
declaradas sensibles. El Plan lleva retraso en su ejecución debido en parte a la
fuerte caída de la inversión pública en infraestructuras del agua. A finales de
2015 operaban en España un total de 2940 plantas EDAR, las cuales depuraron
durante el ejercicio un volumen de 5160 hm3. Es necesaria la renovación y
adecuación tecnológica del parque de EDAR existente, así como la ampliación
futura de la red, de cara a incrementar el grado de conformidad de la carga
contaminante tratada, de acuerdo con la normativa europea [49]. La meta fijada
por el Ministerio para alcanzar los objetivos ambientales es conseguir que en
2021 el 74 % de las masas de agua alcance el buen estado y el 93 % antes de
finalizar 2027 [50].
Por lo general, en la depuración de aguas residuales urbanas suelen aplicarse
únicamente tratamientos primarios y secundarios, con los que se consigue
reducir en gran medida la contaminación de estos efluentes, pero sin lograr a
menudo cumplir la normativa vigente, cada vez más estricta. Además, tal como
se ha comentado anteriormente, muchos de los contaminantes emergentes son
refractarios a los tratamientos convencionales de las EDAR. La tendencia actual
Introducción y objetivos
23
es la utilización de tratamientos terciarios, entre los que destacan una o varias
etapas de oxidación química, tanto para cumplir la normativa, como para lograr
reutilizar el agua, que es el gran objetivo del futuro [51].
Con la Directiva Marco del Agua, la reutilización se plantea como una
posible medida complementaria de protección de los recursos hídricos [16]. En
este sentido, en nuestro país el Real Decreto 1620/2007 [52], por el que se
establece el Régimen Jurídico de la Reutilización de las Aguas Depuradas, ha
facilitado su desarrollo de manera que, actualmente, España se sitúa en el
primer puesto de la Unión Europea en reutilización de efluentes de depuradora
[50]. En el citado decreto se aclaran varios conceptos relativos a la reutilización
del agua hasta entonces empleados indistintamente, se decretan los usos
permitidos (urbanos, agrícolas, recreativos, industriales y ambientales) y los
prohibidos (entre los que destaca el consumo humano), fijando para cada uno de
ellos los parámetros de calidad y valores máximos permitidos y determinando el
régimen de control y responsabilidades en relación al mantenimiento de la
calidad, normalizándose los procedimientos administrativos para la obtención
del derecho al uso. Entre los parámetros para evaluar la calidad se encuentran la
carga microbiológica (nematodos intestinales, Escherichia coli, Legionella spp, etc.),
los sólidos en suspensión, la turbidez y el contenido en nitrógeno y fósforo.
En España se comenzó a tramitar el Plan Nacional de Reutilización de Aguas
(PNRA, 2012) [53] como herramienta de planificación y gestión para el fomento
de la reutilización con dos horizontes: 2015 y largo plazo. La capacidad de
regeneración de aguas depuradas prevista para 2015 se estimaba en 1000 hm3 al
año. Actualmente, los planes hidrológicos sitúan la cifra de agua regenerada en
unos 400 hm3 anuales [50], aún muy lejos de la cantidad proyectada por la
Administración. El precio de la regeneración es la principal barrera para el uso
del agua regenerada en España (solo se reutiliza el 11 % de las aguas residuales)
[50]. El PNRA se planteó sin compromisos de financiación y parece haber
sucumbido a los cambios de gobierno. Sin embargo, continúan desarrollándose
multitud de iniciativas de reutilización a escala regional y local. Gracias a estas
CAPÍTULO 1
24
iniciativas, tres cuartas partes del agua regenerada (75 %) se destina a la
agricultura, el 12 % al riego de campos de golf y jardines, el 6 % a servicios
urbanos, el 4 % a recarga de acuíferos y en torno al 3 % a un uso industrial,
según los últimos datos disponibles [50].
Los beneficios del empleo de aguas regeneradas en la agricultura son
notables, no solo por el ahorro de agua dulce que supone sino porque, además,
el agua regenerada puede contener nutrientes provechosos para la tierra de
cultivo dado el valor fertilizante de los mismos. No obstante, la reutilización no
está exenta de riesgos. Puede existir una falta de idoneidad agronómica para
riego si las aguas depuradas presentan una alta salinidad y elevadas
concentraciones de iones fitotóxicos [54]. Independientemente del uso que se le
vaya dar al agua regenerada, existen también riesgos sanitarios y
medioambientales debidos a la posible presencia de microorganismos patógenos
[55] y de compuestos orgánicos capaces de producir efectos perjudiciales a largo
plazo. Cabe destacar nuevamente la presencia de contaminantes emergentes en
las aguas regeneradas en las depuradoras, aspecto que puede poner en peligro el
futuro de la reutilización al constituir una entrada potencial de estos compuestos
en el medioambiente. Así, algunos estudios han demostrado que los
contaminantes emergentes presentes en las aguas reutilizadas son adsorbidos
por el suelo, pudiendo a continuación acumularse en las plantas cultivadas y
llegar a través de la cadena alimentaria al ser humano [29,56,57].
Por los motivos expuestos, al objeto de posibilitar la reutilización de forma
segura de los efluentes depurados y de evitar la llegada de contaminantes
emergentes a los cuerpos de agua, se hace recomendable la aplicación de un
tratamiento terciario adicional a los tratamientos convencionales de las EDAR.
El desarrollo de procesos alternativos y avanzados ha dado lugar a un amplio
campo de estudio en las últimas décadas [58].
Introducción y objetivos
25
1.1.3. Procesos avanzados de oxidación
Los procesos avanzados de oxidación (PAO), definidos por Glaze et al. en
1987 [59], son capaces de oxidar de manera no selectiva un amplio grupo de
contaminantes resistentes a los procesos convencionales de tratamiento de aguas
[14,60]. Los PAO se basan en la producción “in situ” de especies transitorias
energéticas, principalmente radicales hidroxilo (•
HO ), que tienen un alto poder
oxidante. Así, bajo condiciones favorables, es posible lograr la mineralización
completa (transformación hasta CO2, H2O y sales inorgánicas inocuas) de los
contaminantes presentes en el agua [61].
Los distintos PAO se diferencian entre sí por el mecanismo de producción
del radical hidroxilo. Así, se puede diferenciar entre los métodos fotoquímicos,
donde estas especies activas se generan inducidas por algún tipo de radiación; y
los no fotoquímicos, en los cuales el radical hidroxilo se produce sin necesidad
de una fuente de radiación. Esta clasificación se recoge en la Tabla 1.1.
Tabla 1.1. Principales procesos avanzados de oxidación [62].
Procesos no fotoquímicos Procesos fotoquímicos
Ozonización en medio alcalino (O3/HO•) Radiación ultravioleta con peróxido de
hidrógeno (UV/H2O2)
Ozonización en presencia de peróxido de
hidrógeno (O3/H2O2) Radiación ultravioleta y ozono (UV/O3)
Ozonización catalítica (O3/catalizador) Radiación ultravioleta con ozono y
peróxido de hidrógeno (UV/O3/ H2O2)
Procesos Fenton (Fe(II)/H2O2) y
relacionados Fotocatálisis heterogénea (UV/catalizador)
Oxidación electroquímica o
electrocatalítica
Foto‐Fenton (Fe(II)/H2O2/UV) y
relacionados
Descarga electrohidráulica‐ultrasonido UV/Fe(III) y relacionados
(UV/ferricomplejos)
Oxidación húmeda y supercrítica Fotólisis en el ultravioleta de vacío (UVV)
Radiólisis γ y tratamiento con haces de
electrones
CAPÍTULO 1
26
Cabe destacar que la combinación de dos o más PAO ha mostrado
frecuentemente efectos sinérgicos en relación con la velocidad de degradación
de los contaminantes [63,64].
El principal inconveniente de los PAO radica en el alto coste de los reactivos
necesarios tales como el ozono, el peróxido de hidrógeno o las fuentes
artificiales de radiación, por lo que su implementación se restringe a situaciones
en las que los procesos biológicos no son eficaces en la eliminación de
determinados contaminantes orgánicos [65]. No obstante, en el caso de los PAO
fotoquímicos podría emplearse la radiación natural proveniente del sol, lo que
supondría un ahorro energético considerable. Muchos de los procesos
fotoquímicos se pueden denominar procesos fotocatalíticos ya que emplean
algún tipo de catalizador. Entre estos se distinguen procesos de fotocatálisis
homogénea (e.g., sistema foto‐Fenton) y procesos de fotocatálisis heterogénea
(e.g., fotocatálisis con TiO2). Puesto que en la presente investigación se ha
aplicado la fotocatálisis heterogénea mediante el uso de distintos catalizadores y
fuentes de radiación, así como su combinación con ozono, a continuación se
describen detalladamente ambos procesos.
1.1.3.1. Fotocatálisis heterogénea
Desde la primera referencia a los procesos fotocatalíticos en 1910 por J.
Plotnikow [66], la fotocatálisis heterogénea ha sido ampliamente estudiada en el
ámbito de degradación de contaminantes, particularmente en fase acuosa.
La base de la fotocatálisis heterogénea es la foto‐excitación de un
semiconductor sólido como consecuencia de la absorción de radiación
electromagnética, procedente de una fuente lumínica con una longitud de onda
determinada (UV e incluso visible). Los semiconductores poseen una estructura
electrónica caracterizada por tener una banda de valencia (BV) llena y una
banda de conducción (BC) vacía, lo que les permite actuar en los procesos redox
inducidos por la luz. Cuando un material semiconductor adecuado es excitado
por fotones que poseen energía de magnitud suficiente para superar la energía
Introducción y objetivos
27
correspondiente al salto de banda (band gap), se produce la transición de
electrones desde la banda de valencia hacia la banda de conducción
generándose pares portadores de carga electrón‐hueco positivo [67]. Puesto que
los huecos de la banda de valencia son oxidantes poderosos y los electrones de
la banda de conducción son buenos reductores, estos portadores de carga son
capaces de inducir reacciones de oxidación o reducción, respectivamente [68].
La excitación del semiconductor puede tener lugar de dos formas: por
excitación directa del semiconductor (el caso más habitual y al que generalmente
hace referencia el término de fotocatálisis heterogénea), de manera que es este el
que absorbe los fotones que llegan a su superficie; o por excitación de moléculas
adsorbidas en la superficie del catalizador que, a su vez, ceden electrones al
semiconductor. En estas últimas reacciones se suele hablar de foto‐
sensibilizadores para referirse a los compuestos orgánicos adsorbidos capaces de
ceder electrones al semiconductor.
En la Figura 1.4 se muestra un esquema del mecanismo global del proceso de
fotocatálisis heterogénea que tiene lugar en una partícula de un semiconductor
de banda ancha, en disolución acuosa aireada en presencia de materia orgánica
[69]. Como se observa en el esquema, cuando un semiconductor sólido es
irradiado con luz de energía suficiente se generan pares electrón‐hueco que, en
cuestión de nanosegundos, pueden migrar a la superficie y reaccionar con las
especies adsorbidas (procesos c y d). Los que no lo hacen, se recombinan entre sí
y la energía se disipa, pudiendo tener lugar la recombinación en la superficie o
en el seno del catalizador (procesos a y b, respectivamente). Los huecos y los
electrones generados dan lugar, a través un mecanismo de reacción complejo, a
especies altamente oxidantes que desempeñan un papel clave en la degradación
de contaminantes orgánicos en disolución acuosa. Obviamente, la
recombinación del par electrón‐hueco es perjudicial para la eficacia del proceso
fotocatalítico al reducir la posibilidad de generación de especies activas.
CAPÍTULO 1
28
Figura 1.4. Esquema del proceso fotocatalítico de oxidación de materia orgánica en
disolución acuosa [69].
Profundizando algo más en el proceso, una vez generado el par
electrón/hueco en las bandas de conducción y de valencia, respectivamente
( BC BVe / h , reacción (1.1)), estas especies móviles pueden migrar a la superficie
del catalizador ( s se / h ) y/o ser fácilmente atrapadas formando estados de
menor movilidad ( T Te / h ), de acuerdo a las reacciones (1.2) y (1.3). A su vez,
estas especies pueden dar lugar a la reacción de recombinación (1.4), en la que
e y h representan todas las formas de electrones y huecos (BC, BV,
superficiales o atrapados). La recombinación supone un problema serio en el
desarrollo de las tecnologías fotocatalíticas al limitar severamente los
rendimientos cuánticos que pueden alcanzarse [69].
BC BVSemiconductor h e h (1.1)
BC s Te e e (1.2)
BV s Th h h (1.3)
Introducción y objetivos
29
e h Semiconductor calor (1.4)
Existe una gran controversia sobre la naturaleza de los huecos atrapados
[70,71]. En general, se ha asumido que el agua adsorbida sobre la superficie del
catalizador puede ser foto‐reducida por los sh dando lugar a radicales hidroxilo
superficiales, de acuerdo con la reacción (1.5). Sin embargo, otros autores han
demostrado que esta reacción está cinética y termodinámicamente impedida
[70], proponiéndose la reacción (1.6) para explicar la formación de huecos menos
móviles [72]. En ella, los sh son atrapados por iones oxígeno terminales de la
red del catalizador ( 2sO ) formando radicales terminales protonados o
desprotonados, dependiendo del pH.
s s 2 s sh HO (H O ) HO (1.5)
2s s s sh O HO O (1.6)
Por otro lado, en caso de existir oxígeno en la superficie de las partículas de
catalizador, este actuará como aceptor de electrones a través de la reacción (1.7),
en la que e representa los electrones en todas sus formas. Los radicales anión
superóxido ( 2O ) formados podrían recombinarse dando lugar a la aparición de
peróxido de hidrógeno (reacción (1.8)), capaz de reaccionar con electrones y/o
con radicales superóxido adicionales y generar radicales libres hidroxilo a través
de la reacción (1.9).
2 2e O O (1.7)
2 2 2 2 2O O 2H H O O (1.8)
‐2 2 2 2e O H O HO HO O (1.9)
En este mecanismo de reacción complejo (pero simplificado), los radicales
libres hidroxilo, •
HO , y/o los huecos atrapados ( T s sh HO / O ), pueden ser
CAPÍTULO 1
30
responsables de la oxidación no selectiva de la materia orgánica presente en el
medio acuoso y su posterior mineralización (reacción (1.10)). La fotocatálisis
heterogénea es uno de los PAO para el tratamiento de aguas más estudiados,
demostrando ser una tecnología muy eficiente en el tratamiento de sustancias
difícilmente biodegradables y altamente refractarias a los tratamientos
convencionales de las EDAR [69,73,74].
T T 2 2HO h C C ʹ HO h . . . CO H O (1.10)
Entre las variables que pueden afectar al proceso fotocatalítico, aparte de
parámetros propios tales como la intensidad y longitud de onda de la radiación
empleada, la geometría del reactor y la concentración de contaminantes y de
catalizador, algunos investigadores postulan que, si bien el H2O2 generado en la
reacción (1.8) podría actuar como captador de electrones a través de (1.9), es
crítico el aporte continuo de oxígeno al sistema para el mantenimiento de la
reacción (1.7), evitando así la recombinación del par e / h . Por otra parte, existe
un claro acuerdo en el efecto dominante del pH del medio dada su influencia
sobre el estado de la superficie del semiconductor, la banda de potencial y el
grado de disociación de los contaminantes.
1.1.3.2. Combinación de ozono y fotocatálisis heterogénea: ozonización fotocatalítica
La eficacia del proceso fotocatalítico puede mejorarse, tanto en términos
cinéticos (mayor velocidad de degradación de contaminantes), como
económicos (menores costes energéticos), con la presencia de otras especies
oxidantes o mediante su combinación, secuencial o simultánea, con otras
operaciones físicas o químicas de tratamiento de aguas.
En este trabajo de Tesis se ha estudiado la combinación simultánea del
proceso de fotocatálisis heterogénea empleando distintos semiconductores y
ozono (ozonización fotocatalítica). Esta combinación ha demostrado ser un PAO
muy eficaz al aumentar el rendimiento de formación de especies oxidantes en
comparación con los procesos individuales [61,75], y ha sido aplicada con éxito a
Introducción y objetivos
31
la degradación de diversos contaminantes emergentes en agua [76‐80].
El ozono (O3) es un agente muy reactivo de elevado poder oxidante que
puede actuar frente a compuestos que poseen determinados grupos funcionales.
En algunos casos, estas reacciones dan lugar a la formación de radicales libres
que se propagan por el medio generando, entre otros, radicales hidroxilo que,
como ya se ha comentado, son extremadamente reactivos frente a la materia
orgánica y algunos de los compuestos inorgánicos presentes en el agua [81]. Por
esta razón, las reacciones del ozono en el agua se pueden clasificar como
reacciones directas e indirectas. Las reacciones directas, también llamadas
reacciones de ozono molecular, son aquellas en las que el ozono ataca de forma
altamente selectiva determinadas moléculas orgánicas (compuestos aromáticos y
aromáticos sustituidos, moléculas con insaturaciones ‐C=C‐, ‐C≡C‐, ‐C=N, ‐C=O,
etc.) e incluso puede reaccionar con iones simples [63]. Por otro lado, las
reacciones indirectas son las reacciones en las que participan los radicales
hidroxilo generados a partir de la descomposición del ozono o de su reacción
directa con algunas especies químicas [82]. A diferencia de las reacciones
directas, las reacciones indirectas vía radicales hidroxilo son muy poco
selectivas, puesto que estas especies atacan a cualquier tipo de compuesto
oxidable presente en el agua [83]. La contribución de cada vía (reacción directa
con el ozono o a través de los radicales •
HO generados en su descomposición),
dependerá de factores tales como la naturaleza de los contaminantes, el pH del
medio y la dosis de ozono empleada, principalmente.
En el tratamiento de aguas el ozono puede emplearse como desinfectante de
agentes patógenos, como oxidante clásico para eliminar iones Fe(II) y Mn(II),
algas, olores, sabores, contaminantes orgánicos como pesticidas, detergentes,
fenoles y aminas, entre otros, y como tratamiento previo para mejorar la eficacia
de otras unidades de operación (coagulación, floculación, sedimentación,
oxidación biológica, etc.) [84]. Independientemente de la finalidad que se
persiga, en la aplicación del ozono en el tratamiento de aguas se distinguen
cuatro etapas claramente diferenciadas: 1) generación de ozono “in situ”; 2)
CAPÍTULO 1
32
transferencia del ozono desde la corriente de gas a la fase acuosa; 3) reacción
química; y 4) eliminación del ozono residual del gas de salida dada su toxicidad.
Tanto en fase gas como en disolución acuosa el ozono es relativamente
inestable. En agua destilada, a pH 7 y 20 °C los valores de vida media del ozono
varían entre 20 ‐ 30 minutos y 160 minutos, aumentando su inestabilidad en
medio básico. Esto hace que no pueda almacenarse y distribuirse para su uso,
por lo que debe ser generado en el momento de su empleo a partir de oxígeno
molecular, según se indica en la reacción (1.11) [85].
‐12 33O 2O H 284,5 kJmol
(1.11)
Se trata de una reacción endotérmica y no espontánea (∆G° = 161,3 kJ mol‐1)
[85], por lo que el ozono no solo no puede ser generado por activación térmica
del oxígeno, sino que además descompone fácilmente por calentamiento. El
método de formación de ozono de mayor aplicación es mediante descarga
eléctrica directa, generando radicales de oxígeno atómico que reaccionan con el
oxígeno molecular y forman una molécula de ozono. También puede generarse
mediante otros métodos electrolíticos o fotoquímicos.
Por otro lado, las reacciones del ozono pueden desarrollarse en fase
homogénea, encontrándose tanto el ozono como los compuestos
orgánicos/inorgánicos disueltos en agua, o en fase heterogénea, con aporte de
ozono en fase gas, en cuyo caso hay que considerar la etapa de transferencia de
materia gas‐líquido. La velocidad a la que se transfiera el ozono a la fase acuosa
dependerá de las propiedades que influyen en la absorción del ozono tales como
el pH, concentración de contaminante, fuerza iónica y temperatura, por lo que
también deben considerarse en el diseño del sistema. Por su parte, los
parámetros característicos de la transferencia de materia en un sistema de
ozonización son los coeficientes individual y volumétrico de transferencia de
materia (kL y kLa, respectivamente) y la constante de Henry del sistema, He, que
expresa la relación de equilibrio entre la concentración de ozono en el seno del
gas y su solubilidad en agua, es decir, la concentración de este en la interfase
Introducción y objetivos
33
gas‐líquido.
Una vez transferido el ozono a la fase acuosa o de forma simultánea a esta
etapa tienen lugar las distintas reacciones, entre ellas las reacciones indirectas
por la acción de radicales libres, formados en las etapas de iniciación o
propagación de las reacciones del ozono con otros compuestos tales como el
peróxido de hidrógeno o radiación UV. Así, se han propuesto varios
mecanismos de descomposición de ozono en agua, siendo uno de los más
aceptados el desarrollado por Staehelin, Hoigné y Buhler (SHB) [86] (reacciones
de (1.13) a (1.23)). El mecanismo se muestra a continuación incluyendo, además,
las reacciones directas del ozono y del radical hidroxilo con un contaminante
externo C (reacciones (1.12) y (1.24)).
Reacción directa:
3 oxC zO C (1.12)
Reacciones indirectas:
‐1 1k 70M s
3 2 2O HO HO O (1.13)
6 ‐1 1k 2,2x0 M s ‐
3 2 3 2O HO O HO (1.14)
pK 4,8
2 2HO H O
(1.15)
9 ‐1 1k 1,6x10 M s
3 2 3 2O O O O (1.16)
10 1 1
2 1
k 5x10 M s
3 3k 3,3x10 s
O H HO
(1.17)
5 1k 1,1x10 s
3 2HO HO O (1.18)
5 1 1
7 1 1k 8,3x10 M s
o bien k 9,7x10 M s
2 2 2 2 2 2 2HO HO obienO O H H O O
(1.19)
9 ‐1 1k 2x10 M s
3 2 2O HO HO O (1.20)
CAPÍTULO 1
34
pK 11,3
2 2 2H O HO H (1.21)
7 ‐1 1k 2,7x10 M s
2 2 2 2H O HO HO H O (1.22)
9 ‐1 1k 7 ,5x10 M s
2 2 2HO HO O H O (1.23)
ox 2 2C HO C . . . CO H O (1.24)
En estas reacciones pueden participar, a su vez, una serie de especies
llamadas iniciadoras, promotoras e inhibidoras, que conducen a la
descomposición o estabilización del ozono en agua [87]. Las sustancias
iniciadoras, que como su propio nombre indica inician el mecanismo de
descomposición en cadena, son aquellas que reaccionan directamente con el
ozono, como por ejemplo el ion hidroperóxido ( 2HO) que, a través de la
reacción (1.14), da lugar a la formación del radical ozónido ( •
3O ) y el radical
hidroperóxido (•
2HO ) el cual, de acuerdo con (1.15), se encuentra en equilibrio
con el radical superóxido (•
2O ). Este radical reacciona rápidamente con el ozono
generando más radicales ozónido a través de (1.16). El •
3O formado a través de
(1.14) y (1.16) promueve la aparición del radical •
3HO según el equilibrio (1.17) y,
con ello, la formación de radicales hidroxilo a través de (1.18). Por su parte, las
sustancias promotoras son compuestos orgánicos o inorgánicos que reaccionan
con el radical hidroxilo dando lugar al radical superóxido o hidroperóxido
(dependiendo del pH del medio). Dichos radicales •
2O /
•
2HO se incorporan al
mecanismo de reacción en cadena de descomposición del ozono,
promoviéndolo. Algunos promotores son el ion fosfato, el metanol o el ácido
fórmico. Finalmente, las sustancias inhibidoras son aquellas que reaccionan con
el radical hidroxilo y terminan con las reacciones en cadena, es decir, no dan
lugar al radical hidroperóxido o superóxido, como por ejemplo el ter‐butanol,
determinados iones como carbonatos y bicarbonatos o ciertas sustancias
húmicas.
Introducción y objetivos
35
En este tipo de sistemas hay que tener en cuenta que no todo el ozono en el
gas de alimentación se transfiere al agua, existiendo un remanente en el gas de
salida que, por sus propiedades tóxicas, no puede liberarse directamente a la
atmósfera y debe ser eliminado. Para ello, las plantas de tratamiento que
emplean ozono disponen de unos reactores catalíticos de lecho fijo a través de
los cuales se hace circular el gas de salida, destruyendo el ozono e impidiendo
así su emisión al exterior.
Las reacciones directas entre los contaminantes orgánicos y el ozono suelen
ser de segundo orden (primer orden respecto a cada reactante). Aunque existen
compuestos refractarios al ozono, como es el caso del tricloroacetato (kO3‐C =
3x10‐5 M‐1 s‐1), el ter‐butanol (kO3‐C = 3x10‐3 M‐1 s‐1), el etanol (kO3‐C = 0,37 M‐1 s‐1) o
el benceno (kO3‐C = 2 M‐1 s‐1), entre otros [88], los compuestos que en su estructura
presentan insaturaciones o anillos aromáticos (y en especial si en estos últimos
existen sustituyentes activantes como ocurre, por ejemplo, con los fenoles),
reaccionan de forma rápida con el ozono con valores de kO3‐C del orden de 102 ‐
107 M‐1 s‐1. Por su parte, las reacciones entre los contaminantes orgánicos y el
radical hidroxilo, también de segundo orden, tienen lugar a velocidades muy
superiores, con valores de kHO‐C comprendidos entre 107 ‐ 1010 M‐1 s‐1 [82].
Volviendo al conjunto de reacciones (1.12) a (1.24), es importante destacar el
papel del peróxido de hidrógeno en el mecanismo de descomposición del ozono,
tanto por reacción directa entre su forma iónica y el ozono molecular a través de
(1.14), actuando como iniciador, como en la captura de radicales hidroxilo
promoviendo la descomposición mediante (1.22) y (1.23). Por tanto, la
combinación de peróxido de hidrógeno y ozono constituye por sí sola un PAO,
al igual que la combinación de ion hidróxido (HO‾) y ozono (ozonización en
medio básico), según indica la reacción (1.13).
Por otro lado, si el ozono es irradiado con luz UV suficientemente energética
(longitudes de onda inferiores a 310 nm), puede generarse oxígeno atómico en
estado excitado (O(1D) (reacción (1.25)), especie capaz de generar peróxido de
CAPÍTULO 1
36
hidrógeno al reaccionar con el agua a través de (1.26) [89]. A su vez, la fotólisis
del peróxido de hidrógeno podría dar lugar a la formación del radical hidroxilo
a través de la reacción (1.27). Sin embargo, aunque el rendimiento cuántico de
esta reacción en el intervalo de longitudes de onda de 200 ‐ 400 nm es elevado (1
mol einstein‐1, [90]), la absortividad molar del H2O2 para valores de λ superiores
a 310 nm es muy débil, inferior a 0,51 M‐1 cm‐1 [91]. En consecuencia, el
desarrollo de estas reacciones dependerá de la longitud de onda de la radiación
empleada, de manera que a longitudes de onda superiores a 310 nm la reacción
(1.27) perderá importancia aunque la generación de peróxido de hidrógeno
seguirá favoreciendo la descomposición del ozono igualmente.
h 310nm 1
3 2O O O D
(1.25)
12 2 2O D H O H O (1.26)
h 200 ‐ 280nm
2 2H O 2HO
(1.27)
Por su parte, el oxígeno atómico formado a través de (1.25) no solo es capaz
de reaccionar con el agua (reacción (1.26), k = 1,8x1010 M‐1 s‐1), sino que presenta
valores de k del orden de 109 ‐ 1010 M‐1 s‐1 con prácticamente todos los sustratos
[92].
Además de la formación de oxígeno atómico, hay evidencias de la aparición
en fase gas de otros estados del oxígeno a través de 5 posibles vías, en función
de la longitud de onda:
309nm 1 13 2 gO O D O
(1.28)
3411nm 13 2 gO O D O
(1.29)
31180nm 33 2 gO O P O
(1.30)
Introducción y objetivos
37
1463nm 33 2 gO O P O
(1.31)
612nm 3 13 2 gO O P O
(1.32)
En disolución acuosa estas reacciones también podrían ocurrir. A longitudes
de onda inferiores a 300 nm los productos principales son el estado excitado
(O(1D)) (con un rendimiento cuántico próximo a 0,9 mol einstein‐1), el oxígeno
singlete (O2(1∆g)) y el oxígeno en su estado fundamental (O2(3∑g‾)) (reacciones
(1.28) y (1.29)) [93‐96]. En comparación, las reacciones (1.30) y (1.31) son menos
importantes ya que el rendimiento cuántico de la fotólisis del ozono disminuye
hasta valores cercanos a 0,1 mol einstein‐1 para longitudes de onda superiores a
320 nm.
En cuanto al oxígeno singlete, este puede reaccionar con el ozono y generar
más especies atómicas:
1 32 g 3 2O O 2O O P (1.33)
El O(3P) formado puede, a su vez, desaparecer por reacción con el O2(3∑g‾)
(reacción (1.34)) [97] o con el ozono (reacción (1.35), k ≈ 10‐3 M‐1 s‐1), aunque se ha
comprobado que en el agua esta última reacción no tiene lugar [98].
9 1 13 k 4x10 M s32 g 3O P O O
(1.34)
33 2O P O 2O (1.35)
Finalmente, cuando el ozono está presente en un proceso de fotocatálisis
heterogénea tendrán lugar, además de los mecanismos de reacción
correspondientes a cada sistema por separado, otros mecanismos de generación
de radicales por interacción entre los distintos sistemas. En este sentido, destaca
el papel del ozono como captador de los e generados en la fotoexcitación del
semiconductor con formación del anión radical ozónido (reacción (1.36)),
potenciando así las reacciones (1.17) y (1.18) y, con ello, la aparición de radicales
CAPÍTULO 1
38
•HO [99,100].
3 3e O O (1.36)
La captura de los e por parte del ozono tiene lugar con velocidades
superiores a las que presenta el O2, de manera que mientras que la constante
cinética de la reacción (1.36) a pH comprendido entre 4 y 6 es de 3,6x1010 M‐1 s‐1,
en el caso del O2 (reacción (1.7)) esta resulta ser de 1,76x1010 M‐1 s‐1 [67]. En
consecuencia, en presencia de ozono el grado de recombinación de electrones y
huecos positivos puede verse disminuido, por una parte gracias a la
participación de (1.36) y, por otra, a una mayor contribución de la reacción (1.9)
dada la mayor concentración esperada de H2O2 en el medio procedente tanto de
las reacciones directas con ozono como de la reacción (1.19) [101‐103]. Además
de lo anterior, en presencia de oxígeno y ozono el ion radical superóxido,
generado tanto en la fotocatálisis como en la descomposición de ozono
(reacciones (1.7), (1.14) y (1.15)), puede reaccionar con ozono y generar el ion
radical ozónido a través de la reacción (1.16), desembocando en la formación
adicional de radicales hidroxilo.
En conclusión, atendiendo a los mecanismos de reacción indicados cabe
esperar la existencia de sinergia entre el ozono y la fotocatálisis heterogénea,
dada la gran cantidad de especies oxidantes formadas y la minimización del
proceso de recombinación de electrones y huecos.
1.1.4. Optimización de los procesos fotocatalíticos: empleo de radiación solar y
fotocatalizadores apropiados
Uno de los principales inconvenientes a los que se enfrentan los procesos
fotocatalíticos es el alto coste derivado del uso de lámparas UV. Por este motivo,
el empleo de radiación natural proveniente del sol en reactores con un diseño
adecuado supondría un ahorro energético considerable, minimizando los costes
de operación asociados a este tipo de tratamientos [69]. No obstante, hay que
Introducción y objetivos
39
tener en cuenta que, a diferencia de la luz artificial, la intensidad de la radiación
solar que llega a la superficie terrestre no es constante, sino que se ve afectada
por factores meteorológicos, varía diaria y estacionalmente, depende de la
localización geográfica, etc. La interacción entre la radiación solar y los distintos
componentes atmosféricos (O3, O2, CO2, H2O, aerosoles, etc.) provoca que la
radiación que llega a la superficie de la tierra quede restringida a longitudes de
onda comprendidas entre 300 y 3000 nm [104]. De toda esta radiación, un 5 %
corresponde a la región ultravioleta (95% UVA y 5% UVB), un 46 % a la visible y
un 49 % a la del infrarrojo cercano [105].
Los resultados de las investigaciones sobre el uso de luz solar en los procesos
fotocatalíticos de detoxificación de aguas, iniciada en los años 90 en la
Plataforma Solar de Almería [106], reflejan la eficacia de estos procesos en la
degradación de compuestos en agua, siendo el proceso Fenton y la fotocatálisis
heterogénea los tratamientos más estudiados. Ambos sistemas se han ensayado
con éxito para la eliminación de multitud de contaminantes, muchos de ellos no
biodegradables y/o que por su toxicidad pueden impedir el correcto
funcionamiento de los tratamientos biológicos [69,107,108]. En los últimos años,
la problemática asociada a los contaminantes emergentes ha dado lugar a un
aumento significativo en el número de estudios sobre su eliminación mediante
detoxificación solar [37,109‐112]. En el caso concreto del proceso de ozonización
fotocatalítica solar, estudios recientes informan del éxito de su aplicación en la
degradación de diversos contaminantes emergentes en agua [78,113‐118].
En cuanto a los catalizadores, existen diversos materiales semiconductores
tales como TiO2, ZnO, CdS, FexOy, WO3, CeO2, etc. Se trata de materiales
económicamente asequibles que, en su mayoría, pueden excitarse con radiación
de no muy alta energía (UVA e incluso visible), lo que incrementa el interés en
su empleo en procesos de detoxificación solar. En la Figura 1.5 se muestran las
posiciones de la banda de conducción y de valencia de varios semiconductores,
a partir de las cuales es posible deducir su energía del salto de banda (Eg).
CAPÍTULO 1
40
5
4
3
2
1
0
-1
-2g-C3N4
1,80
-0,90 -0,40
CdS 2,00
-0,75
2,65
0,79
-Fe2O3 2,99
3,47
WO3 0,77
3,75
CeO2
3,13
-0,07
ZnO 3,41
0,21 -0,16
︵rutilo ︶
TiO2
3,04
0,04
pH=0
V vs. ENH
H2O/HO·Eº = +2,73 V
Eº = +0,08 V O2/·O2-
︵anatasa ︶
TiO2
3,04
SnO2
0,25
Figura 1.5. Posiciones de banda (arriba BV y abajo BC) de varios semiconductores
(adaptado de [128]).
En los últimos 30 años, el dióxido de titanio (TiO2) en forma de polvo
policristalino ha sido el semiconductor empleado por excelencia en la
investigación de la descontaminación de aguas mediante procesos fotocatalíticos
debido a su actividad, bajo coste, estabilidad y no toxicidad [69,113,119‐121]. El
producto comercial más conocido es el dióxido de titanio P25 de Degussa
(actualmente Evonik Degussa), constituido por un 20 % de rutilo y un 80 % de
anatasa, siendo esta última la fase cristalina más activa en fotocatálisis. Posee un
área superficial específica de 50 ± 15 m2 g‐1 y un tamaño medio de partícula
cristalina de 20 a 40 nm [122]. Este material ha resultado ser uno de los más
activos en la degradación fotocatalítica de contaminantes del agua empleando
diferentes fuentes de radiación ultravioleta [123‐125]. En lo que respecta a su
empleo en la degradación de compuestos mediante ozonización fotocatalítica,
distintos trabajos ponen de manifiesto su efectividad [78,79,117,118,126,127].
Sin embargo, el TiO2 presenta varios inconvenientes para su aplicación a gran
escala empleando luz solar. El principal de ellos es que al corresponder su
energía del salto de banda (3,0 ‐ 3,2 eV) con longitudes de onda de 387 ‐ 410 nm,
la energía del espectro solar que puede aprovecharse se ve reducida
aproximadamente a un 5 % de la radiación que llega a la Tierra, concretamente a
Introducción y objetivos
41
la fracción UV [69,129]. A esto se suma la dificultad de reutilización del
catalizador en polvo debido a la difícil separación del líquido y, además, una
alta capacidad de recombinación. Como se ha comentado en párrafos anteriores,
esto último podría minimizarse mediante el empleo de ozono como oxidante
[69,79,80].
Actualmente se están estudiando distintas estrategias que permitan solventar
o reducir los problemas anteriores. En cuanto al aprovechamiento de la
radiación solar, existen otros semiconductores alternativos al TiO2 que, por su
salto de banda, son aplicables bajo luz solar. Entre ellos se encuentran el óxido
de zinc (ZnO) y el sulfuro de cadmio (CdS) [129], si bien en disolución acuosa
estos catalizadores han presentado algunos problemas de lixiviación del catión
metálico al medio por efectos de fotocorrosión [129].
El óxido de wolframio (WO3), presenta también un salto de banda menor que
el TiO2 (2,4 ‐ 2,8 eV) que permite la absorción de radiación visible. Además, su
fácil preparación, toxicidad nula y su estabilidad en procesos de oxidación
fotocatalíticos hacen de este semiconductor un buen candidato frente al
fotocatalizador tradicional [130‐133]. Sin embargo, dado que el nivel energético
de su banda de conducción es más positivo que el nivel de reducción del
oxígeno (ver Fig. 1.5), este semiconductor por sí solo no resulta eficiente en
fotocatálisis empleando oxígeno, al carecer este último del potencial suficiente
para capturar de forma eficaz los electrones y evitar la recombinación electrón‐
hueco [130,131]. Sin embargo el ozono, que posee un mayor poder oxidante, sí es
capaz de reaccionar rápidamente con los electrones de la banda de conducción
del WO3 y, por tanto, este material sí resulta eficiente en el proceso de
ozonización fotocatalítica como ya han demostrado los estudios de Nishimoto et
al. [134,135].
Por otra parte, el óxido de cerio (CeO2) es un semiconductor que puede
absorber radiación en el ultravioleta cercano e incluso ligeramente en la región
del visible ya que, en función de la morfología y el tamaño de partícula del
CAPÍTULO 1
42
semiconductor, presenta un salto de banda que varía entre 2,9 y 3,2 eV [136‐138].
Además, los pares electrón‐hueco fotogenerados poseen una vida útil lo
suficientemente larga como para desencadenar reacciones fotocatalíticas tanto
en fase líquida como en fase gas [139].
El CeO2 se ha utilizado como fotocatalizador en la degradación numerosos
contaminantes orgánicos en agua [137,140‐143]. Los trabajos realizados ponen de
manifiesto, como se ha indicado anteriormente, la importancia de la morfología
y del tamaño de partícula en su comportamiento como fotocatalizador. Por otra
parte, el CeO2 ha resultado activo en la ozonización catalítica de compuestos
orgánicos de diferente naturaleza [144‐146]. Dada la eficiencia de los procesos
individuales de fotocatálisis y ozonización catalítica empleando CeO2, es de
esperar un efecto sinérgico en el proceso combinado de ozonización
fotocatalítica, si bien no existe hasta la fecha ningún estudio al respecto.
Además del empleo de semiconductores distintos del TiO2, otra posible
estrategia para mejorar la absorción en el visible es sintetizar materiales
compuestos de dos semiconductores, por lo general acoplando al TiO2 un
semiconductor con una energía del salto de banda menor que la suya [69]. Así,
por ejemplo, al ser la banda de conducción del WO3 menos negativa que la del
TiO2 los electrones pueden transferirse desde la banda del TiO2 a la del WO3,
produciéndose una separación de cargas electrón/hueco más efectiva y evitando
así en cierta medida el proceso de recombinación [147,148]. En la Figura 1.6 se
muestra un esquema del proceso fotocatalítico que tendría lugar en el material
compuesto TiO2‐WO3.
Este tipo de materiales se ha aplicado en la eliminación fotocatalítica de
distintos compuestos en agua, demostrando tener una actividad superior al TiO2
bajo radiación visible [149,150]. Sin embargo, no existen estudios sobre su
aplicación en el proceso de ozonización fotocatalítica.
Introducción y objetivos
43
Figura 1.6. Proceso de excitación en el sistema fotocatalítico TiO2‐WO3 (adaptada de [69]).
Por otra parte, en los últimos años existe un interés creciente por controlar la
morfología de los fotocatalizadores a niveles nanométricos. Si se les da otra
forma y otro tamaño, es decir, si se modifican sus dimensiones, es de esperar
que varíen sus propiedades fotocatalíticas. Se trata de intentar obtener un
semiconductor con una mayor relación área superficial/volumen para aumentar
con ello la eficiencia fotocatalítica. A modo de ejemplo, el TiO2 en forma de
nanotubos ha demostrado poseer una alta superficie específica, una estrecha
distribución de tamaño de poro, unas propiedades electrónicas que mejoran el
transporte de carga y una alta capacidad de intercambio iónico [151]. Todas
estas propiedades pueden favorecer de una forma u otra su actividad
fotocatalítica. Los nanotubos de titanio pueden sintetizarse mediante diversos
métodos, llegándose a conseguir áreas específicas de hasta 500 m2 g‐1 [151‐153].
En cuanto a la estructura cristalina de los nanotubos, en función de las
condiciones de síntesis se ha encontrado la forma tetragonal anatasa [153] y
también otras estructuras monoclínicas u ortorrómbicas [154,155]. Aunque los
resultados obtenidos empleando nanotubos de TiO2 como fotocatalizadores en
el tratamiento de aguas (utilizando colorantes como compuesto a degradar
CAPÍTULO 1
44
[151]), no han sido muy prometedores, sí son interesantes como soporte de
distintas nanopartículas metálicas u otros semiconductores [156]. En lo que
respecta a su aplicación en ozonización fotocatalítica hasta la fecha se ha
publicado un único trabajo [157].
Por último, es posible favorecer la recuperación y reutilización de los
catalizadores mediante la síntesis de materiales soportados, si bien el
fotocatalizador pierde eficiencia al disminuir la superficie activa [69]. Entre los
soportes comúnmente utilizados se encuentran la sílice, vidrios de distinta
morfología, algunos polímeros y membranas poliméricas, carbón activado, etc.
[158]. Una estrategia muy atractiva que se está investigando en los últimos años
es soportar el catalizador (TiO2 generalmente) sobre partículas magnéticas, ya
sean nanopartículas de magnetita [159,160] o carbón activado magnetizado
[78,161,162], lo que permite separar el catalizador mediante un imán externo.
Los materiales carbonosos introducen, además, las propiedades adicionales del
carbón tales como una alta superficie específica y gran volumen de poros,
proporcionando una elevada capacidad de adsorción de compuestos orgánicos
[162,163]. Asimismo, se ha demostrado que algunos materiales carbonosos
podrían actuar donando electrones al semiconductor, inhibiendo los procesos de
recombinación [164].
Aunque todas estas alternativas se han desarrollado y aplicado en la
oxidación fotocatalítica de distintos contaminantes, la bibliografía sobre su
aplicación en la ozonización fotocatalítica es escasa. Por ello, el trabajo que
recoge esta memoria de Tesis persigue el desarrollo de materiales activos y
estables, con vistas a su aplicación en la eliminación de compuestos emergentes
en las aguas mediante ozonización fotocatalítica empleando luz solar.
1.2. OBJETIVOS Y ALCANCE DEL TRABAJO
Dada la preocupación mundial por la presencia de contaminantes
emergentes en los ecosistemas acuáticos, esta investigación tiene como objetivo
principal contribuir al desarrollo de estrategias más eficientes para su
Introducción y objetivos
45
degradación. Se pretende, además, que los posibles procesos a aplicar resulten
atractivos desde el punto de vista económico, empleando para ello la luz solar
como fuente de radiación y sintetizando fotocatalizadores que permitan un
mayor aprovechamiento de la misma.
El trabajo de Tesis se ha centrado en el proceso de ozonización fotocatalítica,
si bien ha sido necesaria la aplicación de otros sistemas con fines comparativos
(fotólisis, ozonización, ozonización fotolítica, adsorción, oxidación fotocatalítica
y ozonización catalítica). Todos los ensayos se han realizado a escala de
laboratorio, empleando luz solar simulada y fotocatalizadores apropiados
previamente sintetizados y/o comerciales.
Para lograr el objetivo general del presente trabajo, se han planteado los
siguientes objetivos específicos:
OBJETIVO 1: Evaluar y comparar la producción de especies oxidantes
fotogeneradas durante los procesos de fotocatálisis y de ozonización
fotocatalítica, determinando el rendimiento cuántico de ambos procesos con
vistas a establecer la sinergia entre el ozono y el proceso de fotocatálisis.
OBJETIVO 2: Preparar y caracterizar fotocatalizadores que puedan ser
fotoexcitados bajo radiación solar, así como mejorar el aprovechamiento de esta
radiación por parte del fotocatalizador más comúnmente utilizado, TiO2,
sintetizando materiales compuestos, evaluando el efecto de las distintas
variables de síntesis a fin de encontrar las formulaciones óptimas para su
empleo en procesos de ozonización fotocatalítica solar.
OBJETIVO 3: Estudiar la degradación de distintos contaminantes emergentes en
agua mediante ozonización fotocatalítica con luz solar, empleando los
catalizadores preparados.
OBJETIVO 4: Desarrollar un modelo cinético para el proceso de ozonización
fotocatalítica solar en base a la identificación de los intermedios de degradación
CAPÍTULO 1
46
de un compuesto modelo y de las principales especies involucradas en el
proceso.
1.3. ORGANIZACIÓN DE LA MEMORIA
Este trabajo de Tesis se estructura en 9 capítulos. En este primero se ha
introducido el trabajo y se han definido los objetivos del mismo. El segundo
capítulo está destinado a las instalaciones y técnicas experimentales empleadas
durante la investigación. En los siguientes capítulos se presentan, como una
colección de artículos de investigación ya publicados en revistas científicas
internacionales, los resultados de los estudios realizados para la consecución de
los objetivos propuestos. Dichos capítulos son los siguientes:
‐ Capítulo 3: On ozone‐photocatalysis synergism in black‐light induced
reactions: Oxidizing species production in photocatalytic ozonation versus
heterogeneous photocatalysis.
‐ Capítulo 4: Influence of structural properties on the activity of WO3
catalysts for visible light photocatalytic ozonation.
‐ Capítulo 5: Visible light photocatalytic ozonation of DEET in the presence
of different forms of WO3.
‐ Capítulo 6: Nanostructured CeO2 as catalysts for different AOPs based in
the application of ozone and simulated solar radiation.
‐ Capítulo 7: WO3‐TiO2 based catalysts for the simulated solar radiation
assisted photocatalytic ozonation of emerging contaminants in a municipal
wastewater treatment plant effluent.
‐ Capítulo 8: Reaction mechanism and kinetics of DEET visible light assisted
photocatalytic ozonation with WO3 catalyst.
Finalmente, en el capítulo 9 se indican las conclusiones más relevantes del
estudio realizado.
Introducción y objetivos
47
BIBLIOGRAFÍA
[1] www.unex.es/investigacion/grupos/trataguas (consultada en febrero de
2017).
[2] Radovic, L.R.; Moreno‐Castilla; C.; Rivera‐Utrilla; J. “Carbon materials as
adsorbents in aqueous solutions” en: “Chemistry and Physics of Carbon”.
Radovic L.R.; Dekker M. (eds.), New York (2000), 27, 227‐403.
[3] Sánchez‐Vila, X. en “Aguas continentales: Gestión de recursos hídricos,
tratamiento y calidad del agua”. Informes CSIC, D. Barceló (coord.), CSIC,
Madrid (2008), Cap. 2.
[4] Li, C.; Goswami, Y.; Stefanakos, E. “Solar assisted sea water desalination: A
review”. Renew. Sust. Energy Rev. 19 (2013) 136‐163.
[5] Xiao, C.; Wang, X.; Ni, M.; Wang, F.; Zhu, W.; Luo, Z.; Cen, K. “A review on
solar stills for brine desalination”. Appl. Energy 103 (2013) 642‐652.
[6] www.aysa.com.ar/index.php?id_contenido=323&id_seccion=0 (consultada en
febrero de 2017).
[7] Carrera, J. en “Aguas continentales: Gestión de recursos hídricos, tratamiento
y calidad del agua”. Informes CSIC, D. Barceló (coord.), CSIC, Madrid (2008),
Cap. 1.
[8] www.unesco.org/new/es/natural‐sciences/environment/water/wwap/facts‐
and‐figures/all‐facts‐wwdr3/fact1‐demographics‐consumption/ (consultada en
febrero de 2017).
[9] Blanco, J.; Malato, S.; Fernández‐Ibáñez, P.; Alarcón, D.; Gernjak, W.;
Maldonado, M.I. “Review of feasible solar energy applications to water
processes”. Renew. Sust. Energy Rev. 13 (2009) 1437‐1445.
[10] Informe de las Naciones Unidas sobre el Desarrollo de los Recursos
Hídricos en el Mundo: Agua y Empleo (2016).
unesdoc.unesco.org/images/0024/002441/244103s.pdf (consultada en febrero de
2017).
[11] Barros, R.; Isidoro, D.; Aragüés, R. “Irrigation management, nitrogen
fertilization and nitrogen losses in the return flows of La Violada irrigation
district (Spain)”. Agr. Ecosyst. Environ. 155 (2012) 161‐171.
[12] Wasi, S.; Tabrez, S.; Ahmad, M. “Toxicological effects of major
environmental pollutants: an overview”. Environ. Monit. Assess. 185 (2013)
2585‐2593.
[13] Osorio, F.; Torres, J.C.; Sánchez, M. en “Tratamiento de aguas para la
CAPÍTULO 1
48
eliminación de microorganismos y agentes contaminantes”. Ediciones Díaz de
Santos, Madrid (2010).
[14] Klavarioti, M.; Mantzavinos, D.; Kassinos, D. “Removal of residual
pharmaceuticals from aqueous systems by advanced oxidation processes”.
Environ. Int. 35 (2009) 402‐417.
[15] Bayo, I.F.; Angulo, E. “Gota a gota”. Química e Industria 582 (2009) 16‐24.
[16] Directiva 2000/60/CE del Parlamento Europeo y del Consejo, de 23 de
octubre de 2000 por la que se establece un marco comunitario de actuación en
el ámbito de la política de aguas.
[17] Directiva 2008/105/CE del Parlamento Europeo y del Consejo, de 16 de
diciembre de 2008, relativa a las normas de calidad ambiental en el ámbito de la
política de aguas, por la que se modifica la Directiva 2000/60/CE.
[18] Directiva 2013/39/UE del Parlamento Europeo y del Consejo, de 12 de
agosto de 2013 , por la que se modifican las Directivas 2000/60/CE y 2008/105/CE
en cuanto a las sustancias prioritarias en el ámbito de la política de aguas.
[19] Decisión nº 2455/2001/CE del Parlamento Europeo y del Consejo de 20 de
noviembre de 2001, por la que se aprueba la lista de sustancias prioritarias en el
ámbito de la política de aguas, y por la que se modifica la Directiva 2000/60/CE.
[20] Esplugas, S.; Bila, D.M.; Krause, L.G.T.; Dezotti, M. “Ozonation and
advanced oxidation technologies to remove endocrine disrupting chemicals
(EDCs) and pharmaceuticals and personal care products (PPCPs) in water
effluents”. J. Hazard. Mater. 149 (2007) 631‐642.
[21] www.mapama.gob.es/es/cartografia‐y‐sig/ide/descargas/agua/red‐
programa‐control.aspx (consultada en febrero de 2017).
[22] Decisión de Ejecución (UE) 2015/495 de la Comisión de 20 de marzo de 2015
por la que se establece una lista de observación de sustancias a efectos de
seguimiento a nivel de la Unión en el ámbito de la política de aguas, de
conformidad con la Directiva 2008/105/CE del Parlamento Europeo y del
Consejo.
[23] Díaz‐Cruz, M.S.; García‐Galán, M.J.; Guerra, P.; Jelic, A.; Postigo, C.;
Eljarrat, E.; Farré, M.; López de Alda, M.J.; Petrovic, M.; Barceló, D. “Analysis of
selected emerging contaminants in sewage sludge”. Trends Anal. Chem. 28
(2009) 1263‐1275.
[24] Rodríguez‐Mozaz, S.; López de Alda, M.J.; Barceló, D. “Advantages and
limitations of on‐line solid phase extraction coupled to liquid chromatography‐
mass spectrometry technologies versus biosensors for monitoring of emerging
Introducción y objetivos
49
contaminants in wáter”. J. Chromatogr. A 1152 (2007) 97‐115.
[25] Karthikeyan, K.G; Meyer, M.T. “Occurrence of antibiotics in wastewater
treatment facilities in Wisconsin, USA”. Sci. Total Environ. 361 (2006) 196‐207.
[26] Lishman, L.; Smyth, S. A.; Sarafin, K.; Kleywegt, S.; Toito, J.; Peart, J.; Lee, B.;
Servos, M.; Beland, M.; Seto, P. “Occurrence and reductions of pharmaceuticals
and personal care products and estrogens by municipal wastewater treatment
plants in Ontario, Canada”. Sci. Total Environ. 367 (2006) 544‐558.
[27] Rosal, R.; Rodríguez, A.; Perdigón‐Melón, J.A.; Mezcua, M.; Hernando,
M.D.; Letón, P.; García‐Calvo, E.; Agüera, A.; Fernández‐Alba, A.R. “Removal
of pharmaceuticals and kinetics of mineralization by O3/H2O2 in a biotreated
municipal wastewater.” Water Res. 42 (2008) 3719‐3728.
[28] Rosal, R.; Rodríguez, A.; Perdigón‐Melón, J.A.; Petrea, A.; García‐Calvo, E.;
Gómez, M.J.; Agüera, A.; Fernández‐Alba, A.R. “Occurrence of emerging
pollutants in urban wastewater and their removal through biological treatment
followed by ozonation”. Water Res. 44 (2010) 578‐588.
[29] Rodríguez‐Gil, J.L.; Catalá, M.; González Alonso, S.; Romo Maroto, R.;
Valcárcel, Y.; Segura, Y.; Molina, R.; Melero, J.A.; Martínez, F. “Heterogeneous
photo‐Fenton treatment for the reduction of pharmaceutical contamination in
Madrid rivers and ecotoxicological evaluation by a miniaturized fern spores
bioassay”. Chemosphere 80 (2010) 381‐388.
[30] Kasprzyk‐Hordern, B.; Dinsdale, R.M.; Guwy, A.J. “The removal of
pharmaceuticals, personal care products, endocrine disruptors and illicit drugs
during wastewater treatment and its impact on the quality of receiving
waters”. Water Res., 43 (2009) 363‐380.
[31] Hazelton, P.D.; Cope, W.G.; Mosher, S.; Pandolfo, T.J.; Belden, J.B.; Barnhart,
M.C.; Bringolf, R.B. “Fluoxetine alters adult freshwater mussel behavior and
larval metamorphosis”. Sci. Total Environ. 445‐446 (2013) 94‐10.
[32] Ikehata, K.; Jodeiri Naghashkar, N.; Gamal El‐Din, M. “Degradation of
aqueous pharmaceuticals by ozonation and advanced oxidation processes: A
review”. Ozone Sci. Eng. 6 (2006) 353‐414.
[33] Oulton, R.L.; Kohn, T.; Cwiertny, D.M. “Pharmaceuticals and personal care
products in effluent matrices: A survey of transformation and removal during
wastewater treatment and implications for wastewater management”. J.
Environ. Monit. 12 (2010) 1956‐1978.
[34] Verlichi, P.; Al Aukidy, M.; Zambello, E. “Ocurrence of pharmaceutical
compounds in urban wastewater: removal, mass load and environmental risk
CAPÍTULO 1
50
after a secondary treatment ‐ A Review”. Sci. Total Environ. 429 (2012) 123‐155.
[35] Xu, J.; Wu, L.; Chang, A.C. “Degradation and adsorption of
selectedpharmaceuticals and personal care products (PPCPs) un agricultural
soils”. Chemosphere 77 (2009) 1299‐1305.
[36] Petrovic, M.; González, S.; Barceló, D. “Analysis and removal of emerging
contaminants in wastewater and drinking water”. Trends Anal. Chem. 22 (2003)
685‐696.
[37] Bernabeu, A.; Vercher, R.F.; Santos‐Juanes, L.; Simón, P.J.; Lardín, C.;
Martínez, M.A.; Vicente, J.A.; González, R.; Llosá, C.; Arques, A.; Amat, A.M.
“Solar photocatalysis as a tertiary treatment to remove emerging pollutants from
wastewater treatment plant effluents”. Cat. Today 161 (2010) 235‐240.
[38] Muñoz, I.; López‐Doval, J.C.; Ricart, M.; Villagrasa, M.; Brix, R.; Geiszinger,
A.; Ginebreda, A.; Guasch, H.; López De Alda, M.J.; Romani, A.M.; Sabater, S.;
Barceló, D. “Bridging levels of pharmaceuticals in River water with biological
community structure in the Llobregat River basin (Northeast Spain)”. Environ.
Toxicol. Chem. 28 (2009) 1706‐2714.
[39] Pomati, F.; Castigloni, S.; Zuccato, E.; Fanelli, R.; Vigetti, D.; Rossetti, C.;
Calamari, D. “Effects of a complex mixture of therapeutic drugs at
environmental levels on human embryonic cells”. Environ. Sci. Technol. 40
(2006) 2442‐2447.
[40] Zegura, B.; Heath, E.; Cernosa, A.; Filipic, M. “Combination of in vitro
bioassays for the determination of cytotoxic and genotoxic potential of
wastewater, Surface water and drinking water samples”. Chemosphere 75 (2009)
1453‐1460.
[41] Muñoz, I.; Rodríguez, A.; Rosal, R.; Fernández‐Alba, A.R. “Life Cycle
Assessment of urban wastewater reuse with ozonation as tertiary treatment. A
focus on toxicity‐related impacts”. Sci. Total Environ. 407 (2009) 1245‐125.
[42] www.abc.es/hemeroteca/historico‐26‐10‐2005/abc/Sociedad/el‐rio‐llobregat‐
tiene‐contaminantes‐que‐da%C3%B1an‐el‐sistema‐reproductor‐de‐los‐
peces_611832583748.html (consultada en febrero de 2017)‐
[43] usatoday30.usatoday.com/news/nation/2008‐03‐10‐drugs‐tap‐water_N.htm
(consultada en febrero de 2017).
[44] elpais.com/diario/2009/10/28/sociedad/1256684408_850215.html (consultada
en febrero de 2017).
[45] www.abc.es/sociedad/20121126/abci‐trazas‐nicotina‐agua‐embotellada‐
201211261705.html (consultada en febrero de 2017).
Introducción y objetivos
51
[46] http://www.lavanguardia.com/vida/20150417/54429985933/delfines‐del‐
estrecho‐de‐gibraltar‐acumulan‐contaminantes‐nocivos‐para‐salud.html
(consultada en febrero de 2017).
[47] Directiva 91/271/CEE del Consejo, de 21 de mayo de 1991, sobre el
tratamiento de las aguas residuales urbanas.
[48] www.mapama.gob.es/es/agua/planes‐y‐estrategias/
PlanNacionalCalidadAguas_tcm7‐29339.pdf (consultada en febrero de 2017).
[49] www.iagua.es/noticias/espana/dbk/16/05/18/negocio‐depuracion‐aguas‐
alcanza‐1230‐millones‐2015 (consultada en febrero de 2017).
[50]www.observatoriosostenibilidad.com/documentos/SOS16_v23_PDF_final.pd
f (consultada en febrero de 2017).
[51] Prat, N.; Rieradevall, M.; Barata, C.; Munné, A. “The combined use of
metrics of biological quality and biomarkers to detect the effects of reclaimed
water on macroinvertebrate assemblages in the lower part of a polluted
Mediterranean river (Llobregat River, NE Spain)”. Ecol. Indic. 24 (2013) 167‐176.
[52] Real Decreto 1620/2007, de 7 de diciembre, por el que se establece el
régimen jurídico de la reutilización de las aguas depuradas.
[53] http://www.mapama.gob.es/es/calidad‐y‐evaluacion‐ambiental/
participacion‐publica/pp_2009p006.aspx (consultada en febrero de 2017).
[54] www.i‐ambiente.es/?q=blogs/aspectos‐considerar‐en‐la‐reutilizacion‐de‐
aguas‐depuradas‐para‐riego (consultada en febrero de 2017).
[55] Scott, T.M.; MccLaughlin, M.R.; Harwood, V.J.; Chivukula, V.; Levine, A.;
Gennaccaro, A.; Lukasick, J.; Farrah, S.R.; Rose, J.B. “Reduction of pathogens,
indicator bacteria, and alternative indicators by wastewater treatment and
reclamation processes”. Water Sci. Technol. Water Supply 3 (2003) 247‐252.
[56] Fatta‐kassinos, D.; Kalavrouziotis, I.K.; Koukoulakis, P.H.; Vasquez, M.I.
“The risks associated with wastewater reuse and xenobiotics in the
agroecological environment”. Sci. Total Environ. 409 (2011) 3555‐3563.
[57] Wu, C.; Spongberg, A.L.; Witter, J.D.; Fang, M.; Czajkowski, K.P. “Uptake
of pharmaceutical and personal care products by soybean plants from soils
applied with biosolids and irrigated with contaminated”. Water Environ. Sci.
Technol. 44 (2010) 6157‐6161.
[58] Wang, J.; Wang, S. “Removal of pharmaceuticals and personal care products
(PPCPs) from wastewater: A review”. J. Environ. Manage. 182 (2016) 620‐640.
[59] Glaze, W. H.; Kang, J.W.; Chapin, D.H. “The chemistry of water treatment
CAPÍTULO 1
52
processes involving ozone, hydrogen peroxide and ultraviolet radiation”. Ozone
Sci. Eng. 9 (1987) 335‐352.
[60] Andreozzi, R.; Caprio, V.; Insola, A.; Marotta, R. “Advanced oxidation
processes (AOP) for water purification and recovery”. Catal. Today 53 (1999) 51‐
59.
[61] Augugliaro, A.; Litter, M.; Palmisano, L: Soria, J. “The combination of
heterogeneous photocatalysis with chemical and physical operations: A tool for
improving the photoprocess performance”. J. Photochem. Photobiol. C
Photochem. Rev. 7 (2006) 127‐144.
[62] Domènech, X.; Jardim, W.F.; Litter, M.I. “Procesos avanzados de oxidación
para la eliminación de contaminantes”. CNEA‐CAC‐UAQ #95‐Q‐03‐05 (2005).
[63] Gogate P.R.; Pandit A.B. “A review of imperative technologies for
wastewater treatment I: oxidation technologies at ambient conditions”. Adv.
Env. Res. 8 (2004) 504‐551.
[64] Beltrán, F.J.; Aguinaco, A.; García‐Araya, J.F; Oropesa, A. “Ozone and
photocatalytic processes to remove the antibiotic sulfamethoxazole from water”.
Water Res. 42 (2008) 3799‐3808.
[65] Esplugas S.; Giménez J.; Contreras S.; Pascual E.; Rodríguez M.
“Comparison of different advanced oxidation processes for phenol
degradation”. Water Res. 36 (2002) 1034‐1042.
[66] J. Plotnikow. Zeitschrift fur Phisikalische chemie‐Stochiometrie und
verwandtschaftslehre 75 (1910) 337‐356.
[67] Bahnemann, D.; Hart, E.J. “Rate constants of the reaction of the
hydratedelectron and hydroxyl radical with ozone in aqueous solution”. J. Phys.
Chem. 86 (1982) 252‐255.
[68] Grätzel, M. “Heterogeneous Photochemical Electron Transfer”. CRC Press,
Boca Raton, FL. (1989).
[69] Malato, S.; Fernández‐Ibáñez, P.; Maldonado, M.I.; Blanco, J.; Gernjak, W.
“Decontamination and disinfection of water by solar photocatalysis: Recent
overview and trends”. Catal. Today 147 (2009) 1‐59.
[70] Salvador, P. “On the nature of photogenerated radical species active in the
oxidative degradation of dissolved pollutants with TiO2 aqueous suspensions: A
revision in the light of the electronic structure of adsorbed water”. J. Phys.
Chem. C 111 (2007) 17038‐17043.
[71] Henderson, M.A. “A surface science perspective on TiO2 photocatalysis”.
Surf. Sci. Rep. 66 (2011) 185‐297.
Introducción y objetivos
53
[72] Goldstein, S.; Behar, D.; Rabani, J. “Mechanism of visible light
photocatalytic oxidation of methanol in aerated aqueous suspensions of carbon‐
doped TiO2”. J. Phys. Chem. C. 112 (2008) 15134‐15139.
[73] Choina, J.; Fischer, G.U.; Flechsig, H.; Kosslick, V,A.; Tuan, N.D.; Tuyen,
N.A.; Tuyen, A.S. “Photocatalytic properties of Zr‐doped titania in the
degradation of the pharmaceutical ibuprofen”. J. Photochem. Photobiol. A 274
(2014) 108‐116
[74] Zhang, H.; Zhang, P.; Ji, Y.; Tian, J.; Du, Z. “Photocatalytic degradation of
four non‐steroidal anti‐inflammatory drugs in water under visible light by P25‐
TiO2/tetraethyl orthosilicate film and determination via ultra performance liquid
chromatography electrospray tandem mass spectrometry”. Chem. Eng. J. 262
(2015) 1108‐1115.
[75] Agustina, T.E.; Ang, H.M.; Vareek, V.K. “A review of synergistic effect of
photocatalysis and ozonation on wastewater treatment”. J. Photochem.
Photobiol. C Photochem. Rev. 6 (2005) 264‐273.
[76] Rivas, F.J.; Beltrán, F.J.; Encinas, A. “Removal of emergent contaminants:
Integration of ozone and photocatalysis”. J. Environ. Manag. 100 (2012) 10‐15.
[77] Rodríguez, E.M.; Fernández, G.; Álvarez, P.M.; Beltrán, F.J. TiO2 and Fe (III)
photocatalytic ozonation processes of a mixture of emergent contaminants of
water. Water Res. 46 (2012) 152‐166.
[78] Quiñones, D.H.; Rey, A.; Álvarez, P.M.; Beltrán, F.J.; Plucinski, P.K.
“Enhanced activity and reusability of TiO2 loaded magnetic activated carbon for
solar photocatalytic ozonation”. Appl. Catal. B Environ. 144 (2014) 96‐106.
[79] Xiao, J.; Xie, Y.; Cao, H. “Organic pollutants removal in wastewater by
heterogeneous photocatalytic ozonation (review)”. Chemosphere 121 (2015) 1‐
17.
[80] Mehrjouei, M.; Müller, S.; Möller, D. “A review on photocatalytic ozonation
used for the treatment of water and wastewater”. Chem. Eng. J. 263 (2015) 209‐
219.
[81] Buxton G.V.; Greenstock C.L.; Helman W.P.; Ross A.B. “Critical review of
rate constant for reactions or hydrated electrons, hydrogen atoms and hydroxyl
radicals (∙OH/∙O‐) in aqueous solutions”. J. Phys. Chem. Ref. Data 17 (1988) 513‐
886.
[82] Beltrán, F.J. “Ozone reaction kinetics for water and wastewater systems”.
Lewis Publishers, CRC Press. Boca Ratón, Florida (YS), 2004.
[83] Hoigné, J.; Bader, H. “The role of hydroxyl radical reactions in ozonation
CAPÍTULO 1
54
processes in aqueous solutions”. Water Res. 10 (1976) 377‐386.
[84] Degremont, s.a. “Water Treatment Handbook”. Sixth edition. 1991.
[85] Portolés, J.J.S. “El ozono atmosférico, ¿benefactor o malhechor?”. Caderno
Brasileiro de Ensino de Física 18 (2001) 360‐363.
[86] Staehelin, J., Buhler, R.E., and Hoigné, J., “Ozone decomposition in water
studied by pulse radiolysis. 2. Hydroxyl and hydrogen tetroxide (HO4) as chain
intermediates”. J. Phys. Chem. 88 (1984) 5999‐6004.
[87] Staehelin, J.; Hoigné, J. “Decomposition of ozone in water in the presence of
organic as promoters and inhibitors of radical chain reactions”. Environ. Sci.
Technol. 19 (1985) 1206‐1213.
[88] Hoigné, J.; Bader, H. “Rate constants of reactions of ozone with organic and
inorganic compounds in water‐ I: Non‐dissociating organic compounds”. Water
Res. 17 (1983) 173‐183.
[89] Peyton, G.R.; Glaze, W.H. “Destruction of pollutants in water with ozone in
combination with ultraviolet radiation. Photolysis of aqueous ozone”. Environ.
Sci. Technol. 22 (1988) 761‐767.
[90] Goldstein, S.; Aschengrau, D.; Diamant, Y.; Rabani J. “Photolysis of aqueous
H2O2: Quantum yield and applications for polychromatic UV actinometry in
photoreactors”. Environ. Sci. Technol. 41 (2007) 7486‐7490.
[91] Chu L.; Anastasio C. “Formation of hydroxyl radical from the photolysis of
hydrogen peroxide in ice”. J. Phys. Chem. A 109 (2005) 6264‐6271.
[92] Biedenkapp, D.; Hartshorn, L.G.; Bair E. “The O(1D) + H2O reaction” J.
Chem. Phys. Lett. 5 (1970) 379‐380.
[93] Wayne, R.P. “The photochemistry of ozone” Atmos. Environ. 21 (1987)
1683‐1694.
[94] Taniguchi, N.; Takahashi, K.; Matsumi, Y.J. “Photodissociation of O3 around
309 nm”. J. Phys. Chem. A 104 (2000) 8936‐8944.
[95] Hancock G.; Tyley, P.L. “The near‐UV photolysis of ozone: quantum yields
of O(1D) between 305 and 329 nm at temperatures from 227‐298 K, and the room
temperature quantum yield of O(3P2) between 303 and 310nm, measured by
resonance enhanced multiphoton ionization.” Phys. Chem. Chem. Phys. 3 (2001)
4984‐4990.
[96] Wine, P.H.; Ravishankara, A.R. “O3 photolysis at 248 nm and O(1D2)
quenching by H2O, CH4, H2, and N2O ‐ O(3Pj) yields” Chem. Phys. 69 (1982) 365‐
373.
Introducción y objetivos
55
[97] Kläning, U.K.; Sehested K.; Wolff T.J. “Ozone formation in laser flash‐
photolysis of oxoacids and oxoanions of chlorine and bromine” Chem. Soc.,
Faraday Trans. 1 80 (1984) 2969‐2979.
[98] Kondratiev, V.N. “Rate Constants of Gas‐Phase Reactions”. National Bureau
of Standards: Washington DC (1972).
[99] Hernández‐Alonso, M.D.; Coronado, J.M.; Maira, A.J.; Soria, J.; Loddo, V.;
Augugliaro, V. “Ozone enhanced activity of aqueous titanium dioxide
suspensions for photocatalytic oxidation of free cyanide ions”. Appl. Catal. B
Environ. 39 (2002) 257‐267.
[100] Marci, G.; García‐López, E.; Palmisano, L. “Mechanistic aspects of oxalic
acid oxidation by photocatalysis and ozonation”. J. Appl. Electrochem. 38 (2008)
1029‐1033.
[101] Rakowski, S.; Cherneva, D. “Kinetics and mechanism of the reaction of
ozone with aliphatic alcohols”. Int. J. Chem. Kinetics 22 (1990) 321‐329.
[102] Mvula, M.; von Sonntag, C. “Ozonolysis of phenols in aqueous solution”.
Org. Biomol. Chem. 1 (2003) 1749‐1756.
[103] Leitzke, A.; von Sonntag, C. “Ozonolysis of unsaturated acids in aqueous
solution: acrylic, methacrylic, maleic, fumaric and muconic acids”. Ozone Sci.
Eng. 31 (2009) 301‐308.
[104] Blanco J.; Malato S. “Solar detoxification”. Unesco Publishing, Renewable
energies series (2003).
[105] http://solarcellcentral.com/solar_page.html (consultada en febrero de
2017).
[106] www.psa.es (consultada en febrero de 2017).
[107] Robert, D.; Malato, S. “Solar photocatalysis: a clean process for water
detoxification”. Sci. Total. Environ. 291 (2002) 85‐97.
[108] Oller, I.; Malato, S.; Sánchez‐Pérez, J.A. “Combination of advanced
oxidation processes and biological treatments for wastewater decontamination‐
A review”. Sci. Total. Environ. 409 (2011) 4141‐4166.
[109] Prieto‐Rodríguez, L.; Miralles‐Cuevas, S.; Oller, I.; Agüera, A.; Puma, G.L.;
Malato, S. “Treatment of emerging contaminants in wastewater treatment plants
(WWTP) effluents by solar photocatalysis using TiO2 concentrations”. J. Hazard.
Mater. 211‐212 (2012) 131‐137.
[110] Velegraki, T.; Hapeshi, E.; Fatta‐Kassinos, D.; Poulios, I. “Solar‐induced
heterogeneous photocatalytic degradation of methyl‐paraben”. Appl. Catal. B
CAPÍTULO 1
56
Environ 178 (2015) 2‐11.
[111] Cordero‐García, A.; Turnes Palomino, G.; Hinojosa‐Reyes, L.; Guzmán‐
Mar, J.L.; Maya‐Teviño, L.; Hernández‐Ramírez, A. “Photocatalytic behaviour of
WO3/TiO2‐N for diclofenac degradation using simulated solar radiation as an
activation source”. Environ. Sci. Pollut. R. DOI: 10.1007/s11356‐016‐8157‐0.
[112] Rimoldi, L.; Meroni, D.; Falleta, E.; Pifferi, V.; Falciola, L.; Cappelletti, G.;
Ardizzone, S. “Emerging pollutant mixture mineralization by
TiO2 photocatalysts. The role of the water medium”. Photochem. Photobiol. Sci.
16 (2017) 60‐66.
[113] Araña, J.; Herrera, J.A.; Doña, J.M.; González, O.; Viera, A.; Pérez, J.;
Marrero, P.M.; Espino, V. “TiO2‐photocatalysis as a tertiary treatment of
naturally treated wastewater”. Catal. Today 76 (2002) 279‐289.
[114] Bayarri, B.; González, O.; Maldonado, M.I.; Giménez, J.; Esplugas,
S. “Comparative study of 2,4‐dichlorophenol by using different photocatalytic
techniques”. J. Sol. Energy Eng. 129 (2007) 60‐67.
[115] Oyama, T.; Yanigisawa, I.; Takeuchi, M; Koike, T.; Serpone, N.; Hidaka, H.
“Remediation of simulated aquatic sites contaminated with recalcitrant
substrates by TiO2/ozonation under natural sunlight”. Appl. Catal. B Environ. 91
(2009) 242‐246.
[116] Oyama, T.; Otsu, T.; Hidano, Y.; Koike, T.; Serpone, N.; Hidaka, H.
“Enhanced remediation of simulated wastewaters contaminated with 2‐
chlorophenol and other aquatic pollutants by TiO2‐photoassisted ozonation in a
sunlight driven pilot‐plant scale photoreactor”. Sol. Energy 85 (2011) 938‐944.
[117] Márquez, G.; Rodríguez, E.M.; Beltrán, F.J.; Álvarez, P.M. “Solar
photocatalytic ozonation of a mixture of pharmaceutical compounds in water”.
Chemosphere 113 (2014) 71‐78.
[118] Quiñones, D.H.; Álvarez, P.M.; Rey, A.; Contreras, S.; Beltrán, F.J.
“Application of solar photocatalytic ozonation for the degradation of emerging
contaminants in water in a pilor plant”. Chem. Eng. J. 260 (2015) 399‐410.
[119] Kabra, K.; Chaudhary, R.; Sawhney, R.L. “Treatment of hazardous organic
and inorganic compounds through aqueous‐phase photocatalysis: A review”.
Ind. Eng. Chem. Res. 43 (2004) 7683‐7696.
[120] Chong, M. N.; B. Jin, C. W. K. Chow, C. Saint. “Recent developments in
photocatalytic water treatment technology: A review”. Water Res. 44 (2010)
2997‐3027.
[121] Chen, J.; Qiu, F.; Xu, W.; Cao, S.; Zhu, H. “Recent progress in enhancing
Introducción y objetivos
57
photocatalytic efficiency of TiO2‐based materials”. Appl. Catal. A Gen. 495
(2015) 131‐140.
[122] Degussa. “Highly dispersed metallic oxides produced by AEROSIL®
process”. Technical Bulletin Pigments, 1990.
[123] Reyes, C.; Fernández, J.; Freer, J.; Mondaca, M.A.; Zaror, C.; Malato, S.;
Mansilla, H.D. “Degradation and inactivation of tetracycline by TiO2
photocatalysis”. J. Photochem. Photobiol. A Chem. 184 (2006) 141‐146.
[124] Xekoukoulotakis, N.P.; Drosou, C.; Brebou, C.; Chatzisymeon, E.; Hapeshi,
E.; Fatta‐Kassinos, D.; Mantzavinos, D. “Kinetics of UV‐A/TiO2 photocatalytic
degradation and mineralization of the antibiotic sulfamethoxazole in aqueous
matrices”. Catal.Today 161 (2011) 163‐168.
[125] Xiong, P.; Hu, J. “Decomposition of acetaminophen (Ace) using
TiO2/UVA/LED system”. Catal. Today 282 (2017) 48‐56.
[126] Beltrán, F.J.; Aguinaco, A.; Rey, A.; García‐Araya, J.F. “Kinetic studies on
black light photocatalytic ozonation of diclofenac and sulfamethoxazole in
water”. Ind. Eng. Chem. Res. 51 (2012) 4533‐4544.
[127] Rodríguez, E.M.; Márquez, G.; León, E.A.; Álvarez, P.M.; Amat, A.M.;
Beltrán, F.J. “Mechanism considerations for photocatalytic oxidation, ozonation
and photocatalytic ozonation of some pharmaceutical compounds in water”. J.
Environ. Manage. 127 (2013) 114‐124.
[128] Ren, H.; Koshy, P.; Chen, W.; Qi, S.; Sorrell, C.C. “Photocatalytic materials
and technologies for air purification”. J. Hazard. Mater. 325 (2017) 340‐366.
[129] Hernández‐Alonso M.D.; Fresno F.; Suarez S.; Coronado J.M.
“Development of alternative photocatalysts to TiO2: Challenges and
Opportunities”. Energ. Environ. Sci. 2 (2009) 1231‐1257.
[130] Zhao, Z.G; Miyauchi, M. “Nanoporous‐walled tungsten oxide nanotubes
as highly active visible‐light‐driven photocatalysts”. Angew. Chem. Int. Ed. 47
(2008) 7051‐7055.
[131] Kim, J.; Lee, C.W.; Choi, W. “Platinized WO3 as an environmental
photocatalyst that generates OH radicals under visible light” Environ. Sci.
Technol. 44 (2010) 6849‐6854.
[132] Sun, S.M.; Wang, W.Z.; Zeng, S.Z.; Shang, M.; Zhang, L.; “Preparation of
ordered mesoporous Ag/WO3 and its highly efficient degradation of
acetaldehyde under visible‐light irradiation” J. Hazard. Mater. 178 (2010) 427‐
433.
[133] Morales, W.; Cason, M.; Aina, O.; Tacconi, N.R.D.; Rajeshwar, K.
CAPÍTULO 1
58
“Combustion synthesis and characterization of nanocrystalline WO3” J. Am.
Chem. Soc. 130 (2008) 6318‐6319.
[134] Nishimoto S.; Mano T.; Kameshima Y.; Miyake M. “Photocatalytic water
treatment over WO3 under visible light irradiation combined with ozonation”
Chem. Phys. Lett. 500 (2010) 86‐89.
[135] Mano, T.; Nishimoto, S.; Kameshima, Y.; Miyake, M. “Investigation of
photocatalytic ozonation treatment of water over WO3 under visible light
irradiation”. J. Ceram. Soc. Jpn. 119 (211) 822‐827.
[136] Huang, W.; Gao, Y. “Morphology‐dependent surface chemistry and
catalysis of CeO2 nanocrystals”. Catal. Sci. Technol. 4 (2014) 3772‐3784.
[137] Aslam, M.; Qamar, M.T.; Tahir‐Soomro, M.; Ismail, I.M.I.; Salah, N.;
Almeelbi, T.; Gondal, M.A.; Hameed, A. “The effect of sunlight induced surface
defects on the photocatalytic activity of nanosized CeO2 for the degradation of
phenol and its derivatives”. Appl. Catal. B Environ. 180 (2016) 391‐402.
[138] Huang, X.; Beck, M.J. “Size‐dependent appearance of intrinsic
Oxq “Activated Oxygen” molecules on ceria nanoparticles”. Chem. Mater. 27
(2015) 5840‐5844.
[139] Hernández‐Alonso, M.D.; Hungría, A.B.; Martínez‐Arias, A.; Fernández‐
García, M.; Coronado, J.M.; Conesa, J.C.; Soria, J. “EPR study of the
photoassisted formation of radicals on CeO2 nanoparticles employed for toluene
photooxidation”. Appl. Catal. B Environ. 50 (2004) 167‐175.
[140] Deng, W.; Chenn, D.; Chenn, L. “Synthesis of monodisperse CeO2 hollow
spheres with enhanced photocatalytic activity”. Ceram. Int. 41 (2015) 11570‐
11575.
[141] Muduli, S.K.; Wang, S.; Chen, S.; Fan‐Ng, C.; Huan, C.H.A.; Sum, T.C.;
Soo, H.S. “Mesoporous cerium oxide nanospheres for the visible‐light driven
photocatalytic degradation of dyes”. Beilstein J. Nanotechnol. 5 (2014) 517‐523.
[142] Yu, Y.; Zhu, Y.; Meng, M. “Preparation, formation mechanism and
photocatalysis of ultrathin mesoporous single‐crystal‐like CeO2 nanosheets”.
Dalton Trans. 42 (2013) 12087‐12092.
[143] Karunakaran, C.; Dhanalakshmi, R. “Semiconductor‐catalyzed degradation
of phenols with sunlight”. Sol. Energ. Mat. Sol. C. 92 (2008) 1315‐1321.
[144] Da Silva, M.F.P.; Soeira, L.S.; Daghastanli, K.R.P.; Martins, T.S.; Cuccovia,
I.M.; Freire, R.S.; Isolani, P.C. “CeO2‐catalyzed ozonation of phenol: The role of
cerium citrate as precursor of CeO2”. J. Therm. Anal. Calorim. 102 (2010) 907‐
913.
Introducción y objetivos
59
[145] Faria, P.C.C.; Orfao, J.J.M.; Pereira, M.F.R. “A novel ceria‐activated carbon
composite for the catalytic ozonation of carboxylic acids”. Catal. Comm, 9 (2008)
2121‐2126.
[146] Gonçalves, A.G.; Orfao, J.J.M.; Pereira, M.F.R. “Ozonation of bezafibrate
over ceria and ceria supported on carbon materials”. Environ. Technol. 36 (2015)
776‐785.
[147] Keller, V.; Garin, F. “Photocatalytic behavior of a new composite ternary
system: WO3/SiC‐TiO2. Effect of the coupling of semiconductors and oxides in
photocatalytic oxidation of methylethylketone in the gas phase”. Catal.
Commun. 4 (2003) 377‐383.
[148] Dan, S.; Jingyu, W.; Yupan, T. “Constructing WO3/TiO2 composite
structure towards sufficient use of solar energy”. Chem. Commun. 47 (2011)
4231‐4233.
[149] Ismail, M.; Bousselmi, L.; Zahraa, O. “Photocatalytic behavior of WO3‐
loaded TiO2 systems in the oxidation of salicylic acid” J. Photochem. Photobiol.
A Chem. 222 (2011) 314‐322.
[150] Lana‐Villarreal T.; Monllor‐Satoca D.; Rodes A.; Gómez R. “Photocatalytic
behavior of suspended and supported semiconductor particles in aqueous
media: Fundamental aspects using catechol as model molecule” Catal. Today
129 (2007) 86‐95.
[151] Hernández‐Alonso, M.D.; García‐Rodríguez, S.; Sánchez, B.; Coronado,
J.M. “Revisiting the hydrothermal synthesis of titanate nanotubes: new insights
on the key factors affecting the morphology”. Nanoscale 3 (2011) 2233‐2240.
[152] Jung, J.H.; Lovayashi, H.; Van Bommel, K.J.C.; Shinkai, S.; Shimizu, T.
“Creation of novel helical ribbon and double‐layered nanotube TiO2 structures
using an organogel template”. Chem. Mater. 14 (2002) 1445‐1447.
[153] Kasuga, T.; Hiramatsu, M.; Hoson, A.; Sekino, T.: Niihara, K. “Formation of
titanium oxide nanotube”. Langmuir 14 (1998) 3160‐3163.
[154] Chen, Q.; Du, G.; Zhang, S.; Peng, L.M. ”The structure of trititanate
nanotubes”. Acta Crystallogr. Sect. B 58 (2002) 587‐593.
[155] Wang, W.; Varghese O.K.; Paulose M.; Grimes C.A. ”A study on the
growth and structure of titania nanotubes”, J. Mater. Res. 19 (2004) 417‐422.
[156] Capula, S.I. “Síntesis, caracterización y evaluación de la actividad catalítica
de nanopartículas Pt‐Ir sobre nanotubos de titania”. Tesis Doctoral, Escuela
Superior de Ingeniería Química e Industrias Extractivas, México D.F, México
(2007).
CAPÍTULO 1
60
[157] Pan, L.; Ji, M.; Lu, B.; Wang, X.; Zhao, L. ”Degradation of humic acid by
TiO2 nanotubes/UV/O3”. International conference on environmental science and
information application technology, Vol II (2009) Proceedings, 654‐657.
[158] Shan, A.Y.; Ghati, T.I.M.; Rashid, S.A. “Immobilisation of titanium dioxide
onto supporting materials in heterogeneous photocatalysis: A review”. Appl.
Catal. A Gen. 389 (2010) 1‐8.
[159] Belessi, V.; Lambropoulou, D.; Konstantinou, I.; Zboril, R.; Tucek, J.; Jancik,
D.; Albanis, T.; Petridis, D. “Structure and photocatalytic performance of
magnetically separable titania photocatalysts for the degradation of propachlor”.
Appl. Catal. B Environ. 87 (2009) 181‐189.
[160] Álvarez, P.M.; Jaramillo, J.; López‐Piñeiro, F.; Plucinski, P.K. “Preparation
and characterization of magnetic TiO2 nanoparticles and their utilization for the
degradation of emerging pollutants in water”. Appl. Catal. B Environ. 100 (2010)
338‐345.
[161] Wang, S.; Zhou, S. “Titania deposited on soft magnetic activated carbon as
a magnetically separable photocatalyst with enhanced activity”. Appl. Surf. Sci.
256 (2010) 6191‐6198.
[162] Ao, Y.H.; Xu, J.J.; Fu, D.G.; Shen, X.W.; Yuan, C.W. “Low temperature
preparation of anatase TiO2‐coated activated carbon”. Coll. Surf. A
Physicochem. Eng. Asp. 312 (2008) 125‐130.
[163] Li Puma, G.; Bono, A.; Krishnaiah, D.; Collin, J.C. “Preparation of titanium
dioxide photocatalyst loaded onto activated carbon support using chemical
vapor deposition: A review paper”. J. Hazard. Mater. 157 (2008) 209‐219.
[164] Wang, W.D.; Serp, P.; Kalck, P.; Faria, J.L. “Photocatalytic degradation of
phenol on MWNT and titania composite catalysts prepared by a modified sol‐
gel method”. Appl. Catal. B Environ. 56 (2005) 305‐312.
CAPÍTULO 2 Materiales y métodos experimentales
En este capítulo se describen las instalaciones experimentales empleadas y el procedimiento general de uso de las mismas. Asimismo, se detallan los métodos y equipos de análisis utilizados tanto para el seguimiento de la eficacia de los tratamientos de eliminación de contaminantes en agua como para la caracterización de los catalizadores sintetizados. Finalmente, se indican los productos químicos empleados a lo largo de esta investigación, clasificados en función de su aplicación.
Materiales y métodos experimentales
63
2.1. INSTALACIONES Y PROCEDIMIENTOS EXPERIMENTALES
En este apartado se describen de forma detallada las instalaciones
experimentales empleadas y el procedimiento general de uso de las mismas,
agrupadas en función de la finalidad que persiguen dentro de la presente
investigación.
2.1.1. Tratamientos de eliminación de contaminantes en agua
2.1.1.1. Ensayos realizados empleando lámparas de luz UVA (luz negra) como fuente de
radiación
En la Figura 2.1 se muestra un esquema de la instalación experimental en la
que se ha llevado a cabo el estudio de la degradación de compuestos orgánicos
en agua mediante diferentes procesos (fotólisis, oxidación fotocatalítica u
ozonización fotocatalítica), utilizando como fuente de radiación lámparas de luz
UVA cercana al visible (luz negra).
Figura 2.1. Instalación experimental en ensayos que emplean luz UVA como fuente de
radiación. 1: Botella de oxígeno; 2: Generador de ozono; 3: Rotámetro; 4: Entrada de gas;
5: Barra magnética; 6: Toma de muestra; 7: Termómetro; 8: Lámpara de luz negra y
soporte; 9: Agitador magnético; 10: Salida de gas; 11: Analizador de ozono; 12: Reactor
fotoquímico; 13: Caja negra.
1
2
3
4
5
6
7
8
9
10
11
12
13
CAPÍTULO 2
64
La instalación estaba formada por una caja de madera pintada de color negro
y de dimensiones 50 x 30 x 30 cm en cuyo interior se situaba el reactor. Se trataba
de un reactor cilíndrico de vidrio borosilicato 3.3, de 1 L de capacidad (18 cm de
altura y 10 cm de diámetro), situado sobre un agitador magnético (Agimatic‐E,
P. Selecta®), que inducía el movimiento de una barra magnética recubierta de
teflón situada en el interior del recipiente, garantizando una buena mezcla del
medio de reacción. En dos esquinas opuestas de la caja negra y a una distancia
de 5 cm del reactor, se situaban sendas lámparas de luz negra (LAMP15TBL
HQPOWERTM de 15 W de potencia nominal cada una de ellas, de dimensiones
45 cm de longitud y 2,5 cm de diámetro, fabricadas por Velleman®). El espectro
de emisión de este tipo de lámparas abarca, tal como se muestra en la Figura 2.2,
el intervalo de longitudes de onda comprendido entre 350 ‐ 400 nm, con máximo
de emisión a 365 nm. Este espectro se determinó con ayuda de un
espectrorradiómetro UV‐Vis compacto de fibra óptica que medía en el rango de
longitudes de onda 190 ‐ 850 nm (Black‐Comet modelo C de StellarNet).
200 300 400 500 600 700 800
(nm)
INT
EN
SID
AD
DE
LA
SE
ÑA
L (u
.a.)
Figura 2.2. Espectro de emisión de las lámparas de luz negra empleadas.
La instalación experimental disponía, además, de un sistema de alimentación
de oxígeno que permitía el burbujeo directo de la disolución a través de un
Materiales y métodos experimentales
65
difusor, fijando el caudal de gas con ayuda de un rotámetro (Gilmont). La tapa
del reactor contaba con diferentes bocas adaptadas para la entrada‐salida de
gases, la medida de la temperatura y la toma de muestras. En el caso de los
ensayos en los que se empleó ozono, a la instalación experimental descrita se le
acoplaba un sistema de generación de ozono y un equipo de análisis y
destrucción de ozono en la corriente de gas de salida. El ozono era generado a
partir de oxígeno puro en un ozonizador Sander Labor‐Ozonisator (modelo
301.7), fijando la concentración requerida, según el ensayo. La concentración de
ozono en el gas de salida del ozonizador (antes del inicio de la reacción), y en el
gas de salida del reactor (una vez abierto el paso de gases al reactor), se
analizaba en continuo mediante un analizador Anseros Ozomat GM6000 Pro
dotado de un sistema de destrucción catalítica del ozono.
El procedimiento experimental se iniciaba encendiendo las lámparas y
esperando un tiempo de 30 minutos (tiempo suficiente para que las lámparas
alcanzaran el régimen estacionario de emisión), antes de introducir el reactor en
la instalación. Mientras tanto, al reactor se adicionaba la disolución acuosa de
compuesto a degradar, se ajustaba o tamponaba el medio al valor requerido de
pH mediante la adición de ácidos o bases y se ponía en marcha el sistema de
agitación. En los ensayos en los que se empleaba catalizador se tomaba una
muestra inicial del reactor antes de la adición del mismo, se añadía la cantidad
requerida de catalizador y se mantenía la mezcla en agitación y en la oscuridad
durante 30 minutos, con la finalidad de establecer el equilibrio de adsorción del
compuesto a degradar sobre el catalizador. Pasado ese tiempo, se tomaba otra
muestra (que era considerada como la muestra a tiempo cero), y se filtraba
empleando filtros de jeringa de 0,45 μm (Macherey‐Nagel PET). Finalmente, una
vez alcanzado el régimen estacionario de emisión de las lámparas se introducía
el reactor en la instalación y se iniciaba el burbujeo de gas a través de un difusor
sumergido en la disolución (ver Fig. 2.1). A partir de ese instante, se ponía en
marcha el cronómetro y a distintos intervalos de tiempo se tomaban muestras, se
filtraban y se recogían en distintos viales para su análisis posterior. Si se
empleaba ozono en el ensayo, la forma de proceder era la misma pero poniendo
CAPÍTULO 2
66
especial cuidado en evitar fugas de gas, para lo cual debían sellarse con silicona
las uniones esmeriladas de la tapa del reactor. Además, en este tipo de ensayos
las muestras que se extraían del reactor se dividían en dos fracciones, una
dedicada al análisis de la concentración de ozono disuelto y otra al análisis del
resto de especies involucradas en el proceso. Esta última fracción era burbujeada
inmediatamente con aire para eliminar, por arrastre, el posible ozono disuelto
que hubiera en la muestra, evitando con ello que el ozono siguiera reaccionando.
La duración de los experimentos fue de 60 min.
2.1.1.2. Ensayos realizados empleando luz solar artificial como fuente de radiación.
El estudio a escala de laboratorio de la degradación de contaminantes en
agua mediante distintos procesos empleando luz solar artificial, se llevó a cabo
en una instalación experimental como la que se muestra de forma esquemática
en la Figura 2.3.
Figura 2.3. Instalación experimental en ensayos que emplean luz solar simulada como
fuente de radiación. 1: Botella de oxígeno; 2: Generador de ozono; 3: Rotámetro; 4:
Entrada de gas; 5: Barra magnética; 6: Toma de muestra; 7: Lámpara de Xe; 8: Agitador
magnético; 9: Salida de gas; 10: Analizador de ozono; 11: Reactor fotoquímico; 12:
Simulador solar.
La instalación constaba de un simulador solar Suntest CPS+ (Atlas) de
dimensiones 90 x 35 x 35 cm que disponía de una cámara dotada de una
lámpara xenón refrigerada por aire de 1500 W de potencia, siendo el área total
de exposición dentro de la cámara de 560 cm2. La presencia de filtros de cuarzo y
vidrio restringen el espectro de emisión a longitudes de onda superiores a 300
nm. Dicho espectro, medido con ayuda del espectrorradiómetro UV‐Vis
Materiales y métodos experimentales
67
indicado en el apartado anterior, se muestra en la Figura 2.4 junto con el
espectro solar. En todos los casos, la intensidad de radiación se fijaba en 550
W m‐2 (valor promedio de la radiación solar global en un día soleado), oscilando
la temperatura en el interior de la cámara entre 20 y 40 °C a lo largo de los
experimentos. En la cámara se introducía el reactor, consistente en un reactor
esférico de vidrio borosilicato 3.3 de 0,5 L de capacidad (10 cm de diámetro). La
distancia de la lámpara a la superficie del reactor era de 15 cm
aproximadamente.
200 300 400 500 600 700 800
Simulador Solar RadiaciÓn Solar
(nm)
INT
EN
SID
AD
DE
LA
SE
ÑA
L (u
.a.
)
Figura 2.4. Espectro de emisión característico del sistema lámpara de Xe + filtro de cuarzo
+ filtro de vidrio del simulador solar Suntest CPS+, junto con el espectro solar.
La instalación disponía de un sistema de alimentación de oxígeno con el que
se burbujeaba directamente la disolución acuosa a tratar a través de un difusor,
fijándose el caudal de gas con ayuda de un rotámetro (Gilmont). El reactor
contaba con diversas bocas para la entrada‐salida de gases y la toma de
muestras. En caso de emplear ozono se procedía de la forma descrita en el
apartado anterior, esto es, generándolo a partir de oxígeno puro en un
ozonizador Sander Labor‐Ozonisator (modelo 301.7) y analizando en continuo
su concentración, tanto en el gas de entrada, con ayuda de un analizador GM‐
CAPÍTULO 2
68
600‐OEM de Anseros, como en el gas de salida empleando un equipo Anseros
Ozomat GM6000 Pro dotado de un sistema de destrucción catalítica del ozono.
El procedimiento seguido fue similar al descrito para los ensayos con luz
UVA: el reactor, cargado con la disolución acuosa de compuesto/s a degradar, se
colocaba en el interior de la cámara del simulador solar soportado sobre un
agitador magnético. En caso de emplear catalizador, se tomaba una muestra
inicial de la disolución y a continuación se añadía la cantidad requerida de
catalizador, dejando la mezcla en agitación y en la oscuridad durante 30 minutos
al objeto de establecer el equilibrio de adsorción. Posteriormente, se tomaba una
nueva muestra (considerada tiempo cero), se filtraba y se recogía en un vial para
su análisis posterior. Se encendía la lámpara del simulador solar (se comprobó
previamente que no era necesario esperar un tiempo para alcanzar el régimen
estacionario de emisión), y se abría el paso del gas (oxígeno o mezcla
oxígeno/ozono), al reactor. En ese instante se ponía en marcha el cronómetro y a
distintos tiempos se tomaban muestras y se procedía de la forma descrita en el
apartado anterior. La duración de los experimentos varió en función del objetivo
perseguido con cada tipo de tratamiento.
2.1.2. Preparación de los catalizadores empleados
A continuación se describen los equipos utilizados en la preparación de
distintos catalizadores indicándose, de forma general, la función de cada uno de
ellos en el procedimiento de síntesis.
2.1.2.1. Agitador
En varias de las etapas seguidas para llevar a cabo la síntesis de catalizadores
(hidrólisis sol‐gel, homogeneización, lavado de los materiales sintetizados, etc.),
se empleó un agitador magnético Agimatic‐E de Selecta®. Este inducía el
movimiento de una barra magnética recubierta de teflón situada en el interior
del vaso de precipitados de vidrio que contenía la mezcla de la síntesis.
Materiales y métodos experimentales
69
2.1.2.2. Autoclave
En el caso de los catalizadores sintetizados por vía hidrotermal, el
tratamiento se realizó en una autoclave de acero inoxidable de Parr Instrument
Company, que contenía un vaso de teflón de 125 mL. En el vaso se introducía la
mezcla de reacción evitando que el volumen de la misma superara el 75 % del
volumen total del vaso. Las autoclaves constituyen sistemas sellados en los que
el recinto de reacción se encuentra sometido a condiciones de presión y
temperatura elevadas, como las características de los procesos hidrotermales.
2.1.2.3. Centrífuga
Tras cada etapa de lavado de los materiales sintetizados, las partículas de
catalizador fueron separadas de los disolventes utilizados (ácido clorhídrico,
agua y etanol), con ayuda de una centrífuga Alresa de 200 W de potencia y una
frecuencia de rotación de 50 Hz.
2.1.2.4. Estufa
El secado final del material sintetizado se llevó a cabo en una estufa Conterm
150 L de Selecta®, dotada de un limitador fijo de sobrecalentamiento y un
termostato de seguridad regulable. La estufa permite trabajar con temperaturas
comprendidas entre 45 °C y 250 °C, distribuyendo el calor en su interior por
convección natural. A 100 °C la homogeneidad de esta estufa es de ± 1 °C y la
estabilidad térmica de ± 0,5 °C.
2.1.2.5. Horno mufla
Las partículas de catalizador sintetizadas, así como los propios materiales
precursores en el caso de obtención de catalizadores por descomposición
térmica, se introducían en cápsulas de porcelana y se calcinaban en un horno
mufla Raypa modelo HM, en el que la calefacción se produce mediante placas
termo‐cerámicas con resistencias eléctricas de níquel y cromo. La temperatura se
regula mediante un microprocesador, pudiendo alcanzar hasta 1150 °C, con una
CAPÍTULO 2
70
estabilidad térmica de ± 1 °C y una homogeneidad térmica de ± 5 °C. Asimismo,
los tratamientos hidrotermales de obtención de catalizadores también se
llevaron a cabo en el horno mufla, introduciendo la autoclave en su interior a la
temperatura requerida.
2.2. EQUIPOS Y MÉTODOS DE ANÁLISIS
Dentro del desarrollo experimental de la presente investigación, se pusieron
a punto una serie de métodos de análisis y se emplearon diversos equipos. Estos
se detallan a continuación, clasificados según el objetivo que persiguen.
2.2.1. Seguimiento de la eficacia de los tratamientos de eliminación de
contaminantes en agua
En la Tabla 2.1 se resumen los métodos analíticos y los equipos utilizados
para llevar a cabo la determinación de la eficacia de los distintos tratamientos
ensayados.
Tabla 2.1. Métodos de análisis y equipos utilizados para el seguimiento de la eficacia de
los tratamientos de eliminación de compuestos orgánicos en agua.
Parámetro Método de análisis Equipo
Concentración de
contaminantes modelo
en agua
Cromatografía líquida de alta
resolución (HPLC) ‐ Detector de haz de
diodos (DAD)
Cromatógrafo HPLC‐
DAD Elite LaChrom
Identificación de
productos intermedios
de degradación
Cromatografía líquida de alta
resolución (HPLC) ‐ espectrómetro de
masas (MS) con analizador Q‐TOF
Cromatógrafo HPLC‐
MS‐QTOF Agilent
Concentración de
ácidos carboxílicos e
iones
Cromatografía iónica
Cromatógrafo iónico
Metrohm 881 Compact
Pro
Concentración de
carbono orgánico total
(COT)
Oxidación + espectroscopía infrarroja
Analizador de COT
Shimadzu, TOC‐V
CSH/CSN
Materiales y métodos experimentales
71
Tabla 2.1 (Continuación). Métodos de análisis y equipos utilizados para el seguimiento
de la eficacia de los tratamientos de eliminación de compuestos orgánicos en agua.
Parámetro Método de análisis Equipo
Concentración de
ozono en disolución
acuosa
Oxidación + Espectrofotometría [1]
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
Concentración de
ozono en fase gas Espectrofotometría
Analizador de ozono
Anseros Ozomat
GM6000 Pro
Concentración de
peróxido de hidrógeno
Oxidación + Complejación +
Espectrofotometría [2]
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
Concentración de
hierro total
Reducción + Complejación +
Espectrofotometría [3]
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
Concentración de
hierro (II) Complejación + Espectrofotometría [4]
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
Intensidad de
radiación
Actinometría con ferrioxalato
(Actinómetro de Parker) [5]
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
Concentración de
fomaldehído Reacción + Espectrofotometría [6]
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
Demanda química de
oxígeno (DQO) Oxidación + Espectrofotometría [7]
Digestor LT200.
Espectrofotómetro
DR2800. Cubetas test
LCK 414 (Hach Lange)
Biodegradabilidad.
Demanda biológica de
oxígeno (DBO)
Oxidación + absorción de oxígeno +
medida manométrica Biómetros Oxitop®
Aromaticidad.
Absorción de
radiación UV de 254
nm
Espectrofotometría
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
CAPÍTULO 2
72
Tabla 2.1 (Continuación). Métodos de análisis y equipos utilizados para el seguimiento
de la eficacia de los tratamientos de eliminación de compuestos orgánicos en agua.
Parámetro Método de análisis Equipo
Fosfatos
Complejación + Reducción +
Espectrofotometría (Merck
Spectroquant kit)
Espectrofotómetro
ThermoSpectronic,
Heλios‐α. Cubeta de
cuarzo de 1 cm.
Turbidez Nefelometría Turbidímetro Hanna
HI 93414
pH Potenciometría
pH‐metro Crison GLP
21+ con electrodo
combinado
Conductividad Conductimetría Conductímetro Crison
524
Temperatura Termometría Termómetro de
contacto
2.2.1.1. Determinación de la concentración de contaminantes modelo en agua
La concentración de contaminantes modelo en agua se determinó mediante
cromatografía líquida de alta resolución (HPLC), empleando un equipo Elite
LaChrom dotado de desgasificador, bomba cuaternaria Hitachi L‐2130, sistema
de inyección automático Hitachi L‐2200 y detector de haz de diodos Hitachi
L‐2455. El equipo consta, además, de un software (EZ Chrom Elite) para el
registro y tratamiento de datos.
La fase estacionaria utilizada fue una columna de fase reversa Phenomenex
Gemini C18 de tamaño de partícula 5 m y tamaño de poro 100 Å, de 15 x 0,3
cm. Los compuestos eran separados por elución con agua acidificada al 0,1 %
v/v con ácido fosfórico o ácido fórmico (disolvente A), y acetonitrilo (disolvente
B). En la Tabla 2.2 se resumen las condiciones de cada método empleado para el
análisis de la concentración de los contaminantes. Previamente, a partir del
análisis cromatográfico de muestras patrón de distinta concentración, se obtuvo
la recta de calibrado correspondiente a cada compuesto. Los límites de detección
Materiales y métodos experimentales
73
se muestran también en la Tabla 2.2.
Tabla 2.2. Condiciones de los métodos de HPLC empleados.
Método Condiciones Compuestos tR
(min) (nm)
LD
(g L‐1)
1
Elución isocrática a 0,7 mL min‐1,
40 % disolvente A (ácido
fosfórico). VINY = 50 L. IBP 6,1 220 161,2
2
Elución con gradiente a 0,2
mL min‐1, 90 ‐ 0 % disolvente A
(ácido fórmico) en 40 min y 20
min de re‐equilibrado. VINY = 50
L.
AAP
MTP
CAF
HCT
ANT
SFM
CAR
KET
DCF
10,5
14,8
15,28
16,9
18,2
23,1
27,1
29,4
36,8
244
225
273
271
240
260
285
320
275
7,1
5,4
33,6
11,5
7,0
4,4
10,3
6,6
5,3
3
Elución con gradiente a 0,5
mL min‐1, 95 ‐ 40 % disolvente A
(ácido fosfórico) en 25 min y 10
min de re‐equilibrado. VINY = 99
L.
CAF
MTP
IBP
13,5
15,1
27,5
220
220
220
12
8,4
11,4
4
Elución isocrática a 0,6 mL min‐1,
70 % disolvente A (ácido
fórmico). VINY = 50 L.
DEET 10,4 220 176,1
tR: tiempo de retención; : longitud de onda de detección; LD: límite de detección; VINY:
volumen de inyección
2.2.1.2. Identificación de productos intermedios de degradación
La identificación de productos intermedios de degradación de alguno de los
contaminantes seleccionados en agua fue realizada con ayuda del Servicio de
Análisis Elemental y Molecular (SAEM) de los Servicios de Apoyo a la
Investigación de la Universidad de Extremadura (SAIUEx), el cual dispone de
un equipo de cromatografía líquida con detector de masas tipo Q‐TOF de
Agilent que permite detectar y cuantificar multitud de compuestos orgánicos,
incluso en concentraciones del orden ng L‐1, sin necesidad de preconcentrar la
muestra. El equipo consta de desgasificador (1260‐‐Degasser), bomba binaria
CAPÍTULO 2
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(1260 bin Pump), automuestreador (HiP‐ALS SL+), termostatizador (TCC SL) y
detector de masas tipo Q‐TOF (6520 Accurate‐Mass Q‐TOF LC/MS).
En la separación mediante cromatografía líquida la fase estacionaria fue la
columna Zorbax SB C18 de Agilent (3,5 m, 4,6 x 150 mm), empleándose como
fase móvil una mezcla de acetonitrilo (disolvente A) y agua acidificada con
ácido fórmico 25 mM (disolvente B). Para la separación de los intermedios se
empleó un gradiente de 90 % de B en el minuto 0, a 100 % de A en el minuto 40,
con un caudal de 0,2 mL min‐1. El volumen de inyección fue de 10 L.
Por su parte, el detector Q‐TOF operaba en modo positivo bajo los siguientes
parámetros: voltaje del capilar, 3000 V; presión del nebulizador, 30 psig; caudal
de gas, 10 L min‐1; temperatura del gas, 350 °C; voltaje del fragmentador, 175 V;
voltaje de skimmer, 65 V; Octopolo RF, 750 V. El rango de adquisición de masas
de los compuestos fue 30‐1000 m/z. Los datos fueron adquiridos y procesados
con el programa Agilent MassHunter WorkStation (versión B.04.00). La
adquisición precisa de las masas se realizaba con una fuente de doble
calibración, introduciendo junto al flujo del cromatógrafo una solución estándar
de iones de calibración de purina (C5H4N4, m/z 121,0509) y HP‐921
(C18H18O6N3P3F24, m/z 922,0098).
2.2.1.3. Determinación de la concentración de ácidos carboxílicos de bajo peso molecular
y aniones inorgánicos en disolución
La determinación de la concentración de distintos ácidos orgánicos de cadena
corta (en forma de sus aniones correspondientes) y de algunos aniones
inorgánicos se llevó a cabo mediante cromatografía iónica. Para ello, se utilizó
un cromatógrafo iónico con supresor químico, modelo 881 compact IC Pro
(Metrohm), equipado con una columna de aniones modelo Metrosep A Supp 7
de 4 x 150 mm como fase estacionaria (5 m de tamaño de partícula,
termostatizada a 45 °C). Como fase móvil se empleó una disolución de
Materiales y métodos experimentales
75
carbonato sódico, aplicándose un gradiente en el que la concentración variaba
de 0,6 a 14,6 mM con un caudal constante de 0,7 mL min‐1. La supresión química
se realizó mediante una disolución 250 mM de ácido sulfúrico. La detección de
los aniones se efectuó mediante medidas de conductividad, determinándose la
concentración de los mismos con ayuda del software MagIC Net 2.4 S, a partir
del calibrado previo realizado con disoluciones patrón que contenían entre 0 y
10 mg L‐1 de cada anión. Los tiempos de retención y los límites de detección de
los aniones fueron los siguientes: acetato, tR = 7,3 min y LD = 46 g L‐1; formiato,
tR = 9,6 min y LD = 83 g L‐1; piruvato, tR = 10,3 min y LD = 275 g L‐1; cloruro, tR
= 13,1 min y LD = 143 g L‐1; nitrato, tR = 25,8 min y LD = 272 g L‐1; fosfato, tR =
34,5 min y LD = 350 g L‐1; sulfato, tR = 35,7 min y LD = 268 g L‐1; y oxalato, tR =
39,0 min y LD = 75 g L‐1.
2.2.1.4. Determinación de la concentración de carbono orgánico total en disolución
(COT)
El análisis del contenido en COT se realizó en un analizador TOC‐V
CSH/CSN de Shimadzu, dotado de un sistema de muestreo e inyección
automático. El método analítico se basa en la formación de dióxido de carbono
por combustión catalítica de la muestra a 680 °C, empleando un catalizador de
Pt/Al2O3. El dióxido de carbono generado se analiza posteriormente mediante
un detector de infrarrojos, determinándose así el carbono total (CT) de la
muestra. Por otro lado, para la determinación de carbono inorgánico (CI), se
inyecta una nueva alícuota en un depósito donde es acidificada con ácido
ortofosfórico al 25 %. En estas condiciones, todos los carbonatos/bicarbonatos
disueltos se transforman en dióxido de carbono, cuya concentración es de nuevo
cuantificada en el detector de infrarrojo. El contenido en COT se obtiene por
diferencia (COT = CT ‐ CI).
Previamente, se prepararon disoluciones patrón de ftalato ácido de potasio
con un contenido en carbono comprendido entre 0 y 100 mg L‐1, y disoluciones
de carbonato/bicarbonato de sodio de 0 a 40 mg L‐1 de carbono, obteniéndose a
CAPÍTULO 2
76
partir del análisis de las mismas las rectas de calibrado del CT y el CI,
respectivamente. Los límites de detección fueron 114 g L‐1 para CT y 247 g L‐1
para CI.
2.2.1.5. Determinación de la concentración de ozono en disolución acuosa
Para la determinación de la concentración de ozono disuelto en las muestras
de reacción se utilizó el método colorimétrico propuesto por Bader y Hoigné en
1981 [1], basado en la decoloración que sufre el 5,5’,7‐indigotrisulfonato de
potasio (índigo carmín) cuando reacciona con el ozono.
Para ello, a un vial que contenía 4 mL de disolución 10‐4 M de índigo carmín
de pH 2 se añadían 2 mL de la muestra analizar (fracción de muestra no
burbujeada con aire, tal como se indicó en el apartado 2.2.1.1). Se agitaba la
mezcla, se filtraba en el caso de los ensayos con catalizador, se trasvasaba a una
cubeta espectrofotométrica de 1 cm de camino óptico y se determinaba la
absorbancia de la mezcla a 600 nm. El blanco se preparaba en idénticas
condiciones pero sustituyendo la muestra por 2 mL de agua ultrapura.
De acuerdo con la Ley de Beer, en ausencia de cualquier otra especie que
absorba radiación a 600 nm se tendrá que:
3dis
0 m TO
600nm m
Abs Abs VC
c V
(2.1)
donde CO3dis es la concentración (M) de ozono disuelto en la muestra analizada;
Abs0 la absorbancia del blanco; Absm la absorbancia de la muestra; 600nm la
absortividad molar del índigo carmín a 600 nm (2x104 M‐1 cm‐1); c el camino
óptico (1 cm); VT el volumen total de mezcla (6 mL) y Vm el volumen de muestra
(2 mL). El límite de detección del método es de 10‐6 M de O3.
La disolución de índigo carmín de concentración 10‐3 M se preparaba por
pesada disolviéndolo en agua ultrapura y en medio ácido fosfórico 2x10‐3 M. A
partir de esta disolución madre se preparaba por dilución 1:20 en agua ultrapura
Materiales y métodos experimentales
77
otra de concentración 10‐4 M en medio ácido fosfórico/fosfato monobásico de
potasio de pH 2 (28/35 m/m). Esta disolución se conservaba en la oscuridad y a
una temperatura comprendida entre 2 y 8 °C, siendo estable durante
aproximadamente 4 meses.
2.2.1.6. Determinación de la concentración de ozono en fase gas
Tal como se expuso en el apartado 2.1.1 al describir las instalaciones
experimentales, para determinar la concentración de ozono en fase gas en la
corriente oxígeno/ozono a la entrada y salida del reactor se emplearon los
equipos GM‐600‐OEM y GM‐6000‐PRO de Anseros, respectivamente. La
medida, realizada en continuo durante todo el tiempo de duración del ensayo,
se basa en el análisis espectrofotométrico a 254 nm, longitud de onda a la cual el
oxígeno no absorbe y el ozono presenta un máximo de absorción. Con ayuda de
una lámpara de mercurio de baja presión situada en el interior de una celda
óptica, una fracción de la corriente gaseosa es irradiada a 254 nm. Por aplicación
de la Ley de Beer, la medida de la atenuación de la señal permite determinar la
concentración de ozono en el gas. Los equipos empleados están diseñados para
trabajar a presiones superiores a 2 bar en el rango de concentraciones de O3
comprendido entre 0 y 50 mg L‐1.
2.2.1.7. Determinación de la concentración de peróxido de hidrógeno en disolución
El método empleado ha sido el propuesto por Masschelein et al. en 1977 [2],
basado en la oxidación de Co(II) a Co(III) por parte del peróxido de hidrógeno y
la posterior determinación espectrofotométrica del complejo de bicarbonato‐Co
(III) formado a 260 nm, longitud de onda a la cual el complejo presenta una
elevada absortividad molar (2,664x104 M‐1 cm‐1). Este método es válido para
concentraciones de peróxido de hidrógeno comprendidas entre 3x10‐7 y 5x10‐5
M.
Para llevar a cabo el análisis, en un vial de vidrio que contenía 0,5 mL de
disolución de 4 g L‐1 de Co(II) (preparada disolviendo 1,61 g de cloruro de
CAPÍTULO 2
78
cobalto (II) hexahidrato en 100 mL de agua ultrapura), 10 mL de disolución
saturada de NaHCO3 y 0,5 mL de disolución de 10 g L‐1 de calgón (preparada
disolviendo 1 g de hexametafosfato de sodio en 100 mL de agua ultrapura), se
añadía 1 mL de la muestra analizar. Al objeto de tener en cuenta la posible
contribución de la matriz a la absorbancia a 260 nm, se preparaba otra serie de
viales en los que se sustituían los 0,5 mL de disolución de Co(II) por agua
ultrapura. Los blancos, uno para cada tipo de vial (con y sin cobalto), se
preparaban siguiendo el mismo procedimiento pero reemplazando la muestra
por agua ultrapura. Finalmente, se agitaban las mezclas y se medía la
absorbancia a 260 nm pasados 10‐15 min.
Teniendo en cuenta la Ley de Beer, la concentración molar de peróxido de
hidrógeno en la muestra, CH2O2, puede determinarse mediante la siguiente
expresión:
2 2
2 2
Co(II) H O Co(II) H Om 0 T
H O260nm m
Abs Abs Abs Abs VC
c V
(2.2)
donde (AbsCo(II))0 y (AbsH2O)0 son las absorbancias del blanco en presencia y
ausencia de Co(II), respectivamente; (AbsCo(II))m y (AbsH2O)m las absorbancias de
la muestra en presencia y ausencia de Co(II), respectivamente; 260nm la
absortividad molar del complejo formado a 260 nm; c el camino óptico (1 cm);
VT el volumen total (12 mL) y Vm el volumen de muestra (1 mL).
2.2.1.8. Determinación de la concentración de hierro total en disolución
Para determinar la concentración de hierro total se utilizó un kit comercial
(Spectroquant, Merck). El método se basa en la reducción del Fe(III) existente a
Fe(II) y posterior formación de un complejo estable de color violeta entre el
Fe(II) y la ferrozina presente en el reactivo comercial [3]. Dicho complejo se
analiza mediante espectrofotometría de absorción a 565 nm.
Para el análisis, 3 gotas de reactivo comercial se añadían a 5 mL de muestra o
Materiales y métodos experimentales
79
de muestra diluida, según fuera necesario en función de la concentración de
hierro total en la muestra. Pasados quince minutos se analizaba la absorbancia
de las mezclas a 565 nm.
De acuerdo con la Ley de Beer, la concentración molar de especies de hierro
en disolución en la mezcla, CFeT, viene dada por la ecuación (2.3):
T
m 0 TFe
565nm m
Abs Abs VC
c V
(2.3)
donde Abs0 la absorbancia del blanco (5 mL de agua y 3 gotas de reactivo); Absm
la absorbancia de la muestra; 565nm la absortividad molar del complejo a 565 nm
(27044 M‐1 cm‐1, [8]); c el camino óptico (1 cm); VT el volumen total de la muestra
final (5 mL); y Vm el volumen de muestra no diluida.
2.2.1.9. Determinación de la concentración de hierro (II) en disolución
La concentración de Fe(II) en disolución se determinó mediante el método
propuesto por Zuo en 1995 [4], basado en la medida de la absorbancia a 510 nm
del complejo coloreado formado entre el ion ferroso y la o‐fenantrolina en medio
ácido acético/acetato de pH comprendido entre 3 y 4. A 510 nm el coeficiente de
extinción molar de este complejo es 11023 M‐1 cm‐1 [9].
Para llevar a cabo el análisis, en primer lugar se preparaban los reactivos
necesarios: tampón de ácido acético/acetato 0,1 M de pH entre 3 y 4 (preparado
tomando 572 L de ácido acético glacial y enrasando a 100 mL con agua
ultrapura); disolución de 1,10‐fenantrolina al 0,2 % en peso (preparada
disolviendo 0,2 g de la sal en 100 mL de agua ultrapura, manteniéndose la
disolución refrigerada y preservada de la luz); y disolución de fluoruro de
amonio 2 M (preparada disolviendo 7,4 g de la sal en 100 mL de agua ultrapura).
Una vez preparados, en un vial que contenía 1,5 mL de tampón acético/acetato y
1 mL de o‐fenantrolina, se añadían 5 mL de la muestra a analizar (diluida si fue
necesario). Se agitaba la muestra y se adicionaba 1 mL de fluoruro amónico.
Transcurridos al menos 20 minutos se medía la absorbancia de la mezcla a 510
CAPÍTULO 2
80
nm. El blanco se preparaba en idénticas condiciones pero sustituyendo la
muestra por 5 mL de agua ultrapura.
De acuerdo con la Ley de Beer, en ausencia de cualquier otra especie que
absorba radiación a 510 nm la concentración molar de Fe(II) en la muestra
analizada, CFe(II) , vendrá dada por:
m 0 T
Fe(II)510nm m
Abs Abs VC
c V
(2.4)
donde Abs0 es la absorbancia del blanco; Absm la absorbancia de la muestra;
510nm la absortividad molar del complejo o‐fenantrolina‐Fe(II) a 510 nm; c el
camino óptico (1 cm); VT el volumen total (8,5 mL) y Vm el volumen de muestra
no diluída (5 mL).
2.2.1.10. Determinación de la intensidad de radiación
En la práctica, la actinometría se considera la herramienta que permite
determinar la cantidad total de la luz absorbida en un proceso fotoquímico y,
con ello, la determinación de la intensidad de la luz que las lámparas empleadas
ofrecen y la calibración de los equipos de irradiación. Así, la intensidad de
radiación del sistema dotado con lámparas UVA mostrado en la Fig. 2.1 se
determinó con ayuda del actinómetro de Parker [5], basado en la reducción
fotoquímica del Fe(III) presente en el complejo ferrioxalato [Fe(C2O4)3]3‐ a Fe(II).
La fotorreducción tiene lugar con un rendimiento cuántico de 1 ‐ 1,2
mol einstein‐1 en el intervalo de longitudes de onda comprendidas entre 250 ‐
450 nm [10]. El rendimiento cuántico de un determinado proceso fotoquímico es
la relación entre la cantidad de reactante consumido (o de producto formado) en
la fotorreacción, y la cantidad total de luz absorbida por el sistema a la longitud
de onda utilizada en dicho proceso.
Para llevar a cabo la actinometría, en primer lugar se preparaba 1 L de
disolución de ácido perclórico/perclorato de fuerza iónica 0,03 M y pH = 2
(ajuste con hidróxido sódico). Por otra parte, en sendos matraces aforados de 100
Materiales y métodos experimentales
81
mL de capacidad se preparaba una disolución de concentración 1,5 M en ácido
oxálico y otra de concentración 5x10‐2 M en Fe(III), ambas en medio ácido
perclórico/perclorato. El reactor se cargaba con 800 mL de la disolución inicial de
ácido perclórico/perclorato y se añadían los 100 mL de la disolución de ácido
oxálico. A continuación, se ponía en marcha el sistema de agitación y se
eliminaba el posible oxígeno disuelto mediante el burbujeo de una corriente de
nitrógeno. Pasados unos 20 minutos se tomaba una muestra inicial y se añadían
los 100 mL de la disolución de Fe(III), obteniéndose así una disolución final de
actinómetro Na3[Fe(C2O4)3] de concentración 5x10‐3 M. Alcanzado el régimen
estacionario de emisión de las lámparas, se introducía el reactor en la instalación
y a distintos intervalos de tiempo se extraían alícuotas y se procedía al análisis
del Fe(II) generado en la fotorreducción del Na3[Fe(C2O4)3].
En ausencia de cualquier otra especie susceptible de fotorreaccionar, la
velocidad con la que fotorreduce el ferrioxalato o, lo que es lo mismo, la
velocidad con la que aparece el Fe(II), viene dada por la expresión:
Fe(III) Fe(II)0 act
dC dCI ∙ ∙ 1 exp 2,303∙L∙ ∙C
dt dt (2.5)
donde I0 es la intensidad de la radiación incidente; φ el rendimiento cuántico de
la reacción fotoquímica a la longitud de onda de la radiación; L el paso efectivo
de radiación a través del reactor; ԑ el coeficiente de extinción molar del
ferrioxalato (500 M‐1 cm‐1 a 365 nm; [11]); y Cact la concentración de actinómetro.
Para valores del término 2,303∙L∙ԑ∙Cact superiores a 2, como ocurre en las
condiciones de trabajo elegidas, la ecuación (2.5) se simplifica en la (2.6):
Fe(III) Fe(II)
0
dC dCI ∙
dt dt (2.6)
de manera que la representación de la evolución de la concentración de Fe(II)
con el tiempo debe dar lugar a una línea recta de pendiente I0∙φ. Dado que el
valor de φ es conocido, puede obtenerse finalmente el valor de la intensidad o
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caudal de radiación incidente, I0.
De esta forma se obtuvo para la instalación de luz negra de la Fig. 2.1 un
valor de I0 = 4,43x10‐7 einstein s‐1.
2.2.1.11. Determinación de la concentración de formaldehído en disolución
El método empleado para la determinación de la concentración de
formaldehído fue el propuesto por Nash en 1953 [6], basado en el análisis
espectrofotométrico a 412 nm de la 3,5‐diaceti‐1,4‐dihidrolutidina, producto de
la reacción entre el formaldehído y la acetilacetona en presencia de iones de
amonio, conocida como reacción de Hantzsch.
En un vial de vidrio se introducían 2 mL de una mezcla de acetilacetona y
acetato de amonio (preparada por adición de 0,2 mL de acetilacetona, 3 mL de
ácido acético y 25 g de acetato de amonio a un matraz aforado de 100 mL de
capacidad, enrasando con agua ultrapura), y se añadían entre 1 y 5 mL de
muestra a analizar, completándose con agua ultrapura hasta los 5 mL si fuera
necesario. Las mezclas obtenidas eran calentadas a 50 °C durante 30 minutos en
la oscuridad, atemperadas y analizadas espectrofotométricamente. El blanco se
preparaba de la misma forma añadiendo 5 mL de la muestra a tiempo cero
(muestra que no contenía formaldehído).
Previamente, mediante el análisis de disoluciones patrón de formaldehído
obtenidas a partir del reactivo comercial, se determinó para la
diacetilhidrolutidina una absortividad molar de 7890 M‐1 cm‐1 a 412 nm.
Asimismo, se calculó un valor del límite de detección de este método de
6,57x10‐7 M.
En base a la Ley de Beer, la concentración molar de formaldehído en la
muestra, CCH2O, puede determinarse mediante la siguiente expresión:
Materiales y métodos experimentales
83
2
m 0 TCH O
412nm m
Abs Abs VC
c V
(2.7)
donde Abs0 es la absorbancia del blanco y Absm la de la muestra analizada;
412nm la absortividad molar de la diacetilhidrolutidina a 412 nm; c el camino
óptico (1 cm); VT el volumen total (7 mL) y Vm el volumen de muestra (1 ‐ 5 mL).
2.2.1.12. Determinación de la demanda química de oxígeno (DQO)
La demanda química de oxígeno de una muestra de agua (DQO) representa
la cantidad de oxígeno necesaria para oxidar la materia orgánica e inorgánica
oxidable presente en la misma. Para su análisis se ha empleado un método
colorimétrico de reflujo cerrado [7], en el que la muestra se oxida con dicromato
de potasio en medio ácido sulfúrico y en caliente (148 °C), utilizando sulfato de
plata como catalizador. La interferencia debida a la posible presencia de iones
cloruro se neutraliza adicionando sulfato de mercurio al medio de reacción, que
precipita como cloruro de mercurio. La reducción de Cr(VI) a Cr(III) que tiene
lugar se evalúa fotométricamente por la disminución del color amarillo de la
mezcla.
El análisis de DQO se llevó a cabo empleando cubetas test LCK 414 de Hach
Lange (válidas para el rango de DQO de 5 a 60 mg L‐1 de O2), añadiendo 2 mL
de muestra según el procedimiento indicado por el fabricante. A continuación
las cubetas se introducían en un termorreactor Hach Lange modelo LT200
precalentado a 148 °C y se mantenía la digestión durante 2 horas a dicha
temperatura. Transcurrido ese tiempo, se dejaba que los viales alcanzaran
temperatura ambiente y se analizaba el valor de DQO, proporcionado
directamente por un espectrofotómetro DR2800 de Hach Lange.
2.2.1.13. Determinación de la biodegradabilidad: demanda biológica de oxígeno (DBO)
La demanda biológica de oxígeno de una muestra de agua (DBO5),
representa la cantidad de oxígeno consumido por microorganismos aerobios en
la degradación de la materia orgánica biodegradable presente en la misma
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durante un tiempo de incubación de 5 días. Es, por tanto, una medida de la
biodegradabilidad de un agua en condiciones aerobias.
El análisis de DBO5 del efluente de la EDAR de Badajoz antes/después de su
tratamiento se realizó mediante el método manométrico, basado en la pérdida
de presión en el interior de un recipiente debida al consumo de oxígeno por
parte de las bacterias aerobias involucradas en el proceso. A la vez que degradan
la materia orgánica, dichos microorganismos consumen el oxígeno disuelto en el
agua y liberan dióxido de carbono que se retira del sistema por absorción sobre
hidróxido sódico. El consumo del oxígeno disuelto provoca el desplazamiento
de aire desde el espacio de cabeza al agua para restablecer el equilibrio,
creándose así una depresión tanto mayor cuanto mayor es el oxígeno
consumido. Para determinar la demanda biológica de la materia orgánica
carbonácea el período de incubación es de cinco días.
Para llevar a cabo el análisis, 432 mL de la muestra previamente
acondicionada (filtrada en caso de ensayos con catalizador y con un pH
comprendido entre 6,5 y 7,5) se introducía en un biómetro Oxitop®, dotado de
tapón con dispositivo de medida de la presión del sistema, junto a unas gotas de
disolución inhibidora de nitrificación (disolución 5g L‐1 de aliltiourea, 20 gotas
por litro), y 1 mL de inóculo, procedente de la línea de recirculación de fangos
activos de la depuradora de aguas residuales urbanas de Badajoz. A
continuación, en el interior de la cápsula de goma que forma parte del tapón del
biómetro, se introducían 2 o 3 lentejas de hidróxido sódico para absorber el
dióxido de carbono liberado. El biómetro, cerrado herméticamente, se colocaba
en una cámara de incubación termostatizada a 20 °C y dotada de agitación
individual para cada botella durante cinco días, periodo a lo largo del cual cada
24 horas se almacenaba en la memoria el valor de DBO.
2.2.1.14. Determinación de la absorbancia a 254 nm: Aromaticidad
Los compuestos orgánicos aromáticos se caracterizan por absorber
fuertemente la radiación UV, por lo que la medida de la absorbancia de una
Materiales y métodos experimentales
85
muestra de agua en esa zona del espectro es con frecuencia un indicador de la
concentración de este tipo de compuestos. Esta medida suele realizarse a 254
nm, longitud de onda que minimiza las posibles interferencias de otros
compuestos y, a su vez, proporciona las mayores bandas de absorción de los
compuestos de interés. En este trabajo, las medidas espectrofotométricas se
llevaron a cabo empleando cubetas de cuarzo de 1 cm en las que se introducían
las muestras de agua a analizar previamente filtradas.
2.2.1.15. Determinación de la concentración de fosfatos en disolución
La determinación de la concentración de fósforo en forma de fosfato (P‐PO43‐)
en el efluente de la EDAR de Badajoz antes/después de su tratamiento, se llevó a
cabo empleando el Kit LCK 349 de Lange (rango de medida comprendido entre
0,15 y 4,5 mg L‐1).
Los iones fosfato reaccionan en disolución ácida con iones molibdato y
antimonio formando un complejo antimonilfosfomolibdato, que se reduce en
presencia de ácido ascórbico a azul de fosfomolibdeno, compuesto coloreado
que puede analizarse espectrofotométricamente. La digestión previa de la
muestra a 100 °C durante 60 min (Digestor modelo LT200 de Hach Lange),
conduce a la transformación de todo el fósforo de la muestra en fosfato, lo que
permite determinar no solo el contenido en P‐PO43‐, sino también el contenido en
fósforo total (PT).
Para llevar a cabo el análisis se siguió el procedimiento indicado por el
fabricante, realizando la medición fotométrica con ayuda de un
espectrofotómetro DR2800 de Hach Lange.
2.2.1.16. Determinación de la turbidez
La turbidez es una medida del grado de pérdida de transparencia del agua
debido a la presencia de partículas en suspensión. Se mide en unidades
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nefelométricas de turbidez (NTU) y es una propiedad óptica que hace que la luz
sea dispersada y absorbida, en lugar de ser transmitida.
Para llevar a cabo el análisis se empleó un turbidímetro Hanna HI 93414, el
cual se calibraba con la ayuda de suspensiones patrón de distinta turbidez
suministradas por Hanna Instruments. Este instrumento está dotado de un
sistema óptico que incluye una lámpara con filamento de wolframio para
irradiar la muestra de agua contenida en una cubeta de vidrio, un detector de
luz dispersada y un detector de luz transmitida. Para el rango de medida (0 a
1000 NTU), el microprocesador del instrumento calcula, a partir de las señales
que llegan a los dos detectores, el valor NTU mediante un algoritmo efectivo.
2.2.1.17. Determinación de pH, conductividad y temperatura
La medida de pH se efectuó por potenciometría empleando un pH‐metro
Crison GLP 21+ con electrodo combinado, para un rango de pH comprendido
entre 0 y 14 y un intervalo de temperatura de 0 a 80 °C. El pH‐metro se calibraba
diariamente con disoluciones amortiguadoras de pH 4, 7 y 9, siguiendo el
manual del instrumento.
La conductividad de las muestras de agua se midió también por
potenciometría con un conductímetro Crison 524, el cual se calibraba con una
disolución patrón de cloruro de potasio de concentración 0,01 M (conductividad
1413 S cm‐1 a 25 °C). El microprocesador del instrumento mostraba
directamente la conductividad de la muestra en pantalla.
La medida de temperatura se realizó con un termómetro de contacto de
laboratorio que permitía lecturas en el rango de 0 a 60 °C.
2.2.2. Caracterización de los catalizadores empleados
En este apartado se describen las técnicas analíticas y los equipos que se
utilizaron para llevar a cabo la caracterización de los catalizadores ensayados
durante esta investigación, con el fin de poder correlacionar sus características
Materiales y métodos experimentales
87
químico‐físicas con su actividad catalítica y/o fotocatalítica. En la Tabla 2.3 se
resumen y clasifican todas las técnicas de caracterización empleadas en función
el tipo de información que proporcionan, detallándose cada una de ellas en los
apartados siguientes.
Tabla 2.3. Clasificación de las técnicas y equipos de caracterización empleados.
Determinación Técnica de análisis Equipo
Análisis elemental:
composición química
Espectroscopía de emisión atómica
de plasma por acoplamiento
inductivo (ICP‐OES)
Espectrofotómetro de
emisión atómica ICP‐
OES Optima 3300DV
(Perkin‐Elmer)
Análisis estructural:
fases cristalinas,
tamaño de fases
cristalinas, proporción
de material amorfo, etc
Difracción de Rayos X (XRD) Difractómetro de Rayos
X Bruker D8 Advance
Espectroscopía Raman
Espectrómetro micro‐
Raman Nicolet Almega
XR Dispersive (Thermo
Scientific)
Análisis termogravimétrico y
térmico diferencial (TGA‐DTA)
Termobalanza SETSYS
Evolution‐16 (Setaram)
Microscopía electrónica de
transmisión (TEM)
Microscopio electrónico
de transmisión JEOL
JEM‐2100F y
microscopio electrónico
de transmisión Tecnai
G20 Twin (FEI
Company)
Análisis morfológico y
textural: topografía
superficial, área
superficial, volumen y
distribución de poros.
Microscopía electrónica de barrido
(SEM)
Microscopio electrónico
de barrido Hitachi
S‐4800
Isotermas de adsorción‐desorción
de nitrógeno
Equipo Autosorb‐1
(Quantachrome)
Análisis superficial:
acidez, grupos
superficiales, estado de
oxidación y porcentajes
atómicos en superficie
Valoración másica para determinar
el pH del potencial de carga cero
(pHPZC)
pH‐metro Crison GLP
21+ con electrodo
combinado
Espectroscopía fotoelectrónica de
Rayos X (XPS)
Espectrómetro de
fotoelectrones con
fuente de Rayos X K‐ (Thermo Scientific)
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Tabla 2.3 (Continuación). Clasificación de las técnicas y equipos de caracterización
empleados.
Determinación Técnica de análisis Equipo
Análisis de la
propiedades
electrónicas: ancho de
la banda de energía
prohibida
Espectroscopía ultravioleta‐visible
de reflectancia difusa (DR‐UV‐Vis)
Espectrofotómetro
UV‐Vis‐NIR Cary 5000
(Varian‐Agilent)
2.2.2.1. Espectroscopía de emisión atómica de plasma por acoplamiento inductivo (ICP‐
OES)
La espectroscopía de emisión atómica es una técnica de análisis elemental
capaz de determinar y cuantificar la mayoría de los elementos de la tabla
periódica en concentraciones desde % en peso hasta partes por millón [12]. Esta
técnica está basada en la excitación de un electrón desde su estado fundamental
hasta un nivel energético superior, debida a la absorción de radiación
electromagnética. El átomo así excitado vuelve nuevamente a su estado
fundamental, emitiendo una radiación cuya energía es característica de cada
elemento en particular y cuya intensidad es proporcional a la cantidad de dicho
elemento en la muestra analizada [12].
Esta técnica se ha empleado para llevar a cabo el análisis del contenido total
en W en algunos catalizadores, mediante un analizador ICP‐OES Perkin Elmer,
modelo Optima 3300 DV, equipado con un detector UV. La disgregación de las
muestras se llevó a cabo mediante digestión ácida, vía microondas.
El análisis fue realizado por la Unidad de Apoyo a la Investigación del
Instituto de Catálisis y Petroleoquímica del CSIC (ICP‐CSIC).
2.2.2.2. Difracción de Rayos X (XRD)
La difracción de rayos X es una técnica muy utilizada para la identificación y
caracterización de las fases cristalinas presentes en un sólido. Aprovechando
Materiales y métodos experimentales
89
que las estructuras cristalinas poseen planos (producidos por ordenamientos
repetitivos de átomos), capaces de difractar rayos X, la medida de los ángulos en
los que un haz de rayos X de longitud de onda determinada es difractado por la
muestra permite obtener los diagramas de difracción. El espaciado entre dos
planos hkl (d), que es la base de la caracterización de fases y estructura de un
material, está relacionado con el ángulo de difracción 2θ mediante la ley de
Bragg [13]:
n
d2 sen
(2.8)
donde d es la distancia interplanar, n un número entero que representa el orden
de difracción y λ la longitud de onda de la fuente de rayos X.
Esta es una técnica muy usada por su carácter no destructivo, rapidez y
mínima cantidad de muestra necesaria para identificar las fases cristalinas,
cuantificarlas, determinar el grado de cristalinidad y el tamaño del cristal.
Para la determinación de las dimensiones del cristal se puede emplear la
ecuación de Scherrer [13]:
K
Lcos
(2.9)
siendo L el tamaño de cristal, K una constante relativa a un factor de forma del
cristal (en ausencia de información detallada sobre la forma del cristal K = 0,9 es
una buena aproximación para cristales esféricos [14]) y la anchura a la mitad
de la intensidad máxima de un pico seleccionado.
La señal de difracción emitida por un sólido cristalino es una huella de su
estructura y la intensidad de las líneas de difracción es función de la
concentración de las diferentes fases cristalinas. Por comparación de las
distancias interplanares se puede determinar la fase existente, tomando como
referencia las correspondientes a los compuestos puros [13]. Así, cada
compuesto puro tiene un difractograma específico y unívoco recogido en las
fichas estandarizadas ASTM en la base de datos de la ICDD (International
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Centre for Diffraction Data) [15].
Las medidas de Difracción de Rayos X de distintos catalizadores en polvo
fueron realizadas por el Servicio de Análisis y Caracterización de Sólidos y
Superficies (SACSS) perteneciente a los SAIUEx, empleando un difractómetro
Bruker D8 Advance, que utiliza la radiación Kα del cobre (= 0,1541 nm). Los
difractogramas se registraron para valores de 2θ comprendidos entre 20° y 80°
con paso de barrido de 0,02 s‐1 y tiempo de acumulación de 1 s por punto.
A partir de la intensidad de las líneas de difracción es posible determinar de
forma semicuantitativa el porcentaje de las diferentes fases cristalinas presentes.
Para ello, el equipo dispone del software EVA v.14 (Bruker‐AXS) que, mediante
el método RIR (Reference Intensity Ratio), determina dichos porcentajes
basándose en la relación entre las intensidades de los picos de cada fase
cristalina y la intensidad del corindón como patrón de referencia (Ip/Ipcor).
2.2.2.3. Espectroscopía Raman
La espectroscopía Raman se basa en el efecto Raman, que se produce al
irradiar una sustancia con luz de determinada longitud de onda, teniendo una
parte de la luz dispersada una longitud de onda distinta [16]. Al irradiar una
sustancia con luz monocromática, la mayoría de la dispersión de luz se produce
de forma elástica, es decir, no se produce un intercambio energético neto entre la
radiación y la muestra y, por tanto, la luz dispersada tiene la misma longitud de
onda que la luz incidente, fenómeno al que se denomina dispersión Rayleigh.
Sin embargo, una pequeña parte de la luz incidente produce un cambio en el
estado vibracional del sistema y, en este caso, la luz dispersada tiene una
longitud de onda distinta. Dentro de esta dispersión, conocida como dispersión
Raman, existen dos tipos: la dispersión Stokes, en la que la longitud de onda
final es mayor que la del haz incidente; y la anti‐Stokes, que da lugar a una luz
dispersa con longitud de onda menor que la del haz incidente [17].
De esta manera, la espectroscopía Raman mide frecuencias vibracionales
Materiales y métodos experimentales
91
como diferencias energéticas entre la luz incidente y la dispersada. Estas
diferencias son independientes de la frecuencia del haz incidente. El espectro
Raman se representa como intensidad frente a desplazamiento Raman, siendo
este la diferencia entre el número de onda de la luz dispersa y la de excitación.
La espectroscopía Raman es muy útil en la caracterización de sólidos, pues
además de aportar información sobre la simetría del sistema e identificar las
fases existentes por medio del número, intensidad y posición de las bandas
presentes en el espectro, permite realizar un análisis sobre efectos relacionados
con el tamaño de partícula y la presencia de defectos. Así, por ejemplo, un
incremento de la anchura de las bandas puede deberse a la no estequiometría de
la muestra, ocasionada por la presencia de vacantes de oxígeno o bien inducida
por la presencia de fases parcialmente reducidas minoritarias.
Los espectros Raman de algunos de los catalizadores sintetizados fueron
obtenidos por el SAEM‐SAIUEx. Dichos espectros se registraron en un equipo
Nicolet Almega XR Dispersive micro‐Raman (Thermo Scientific) con una
resolución espectral de 2 cm‐1. Como fuente de excitación se utilizó un láser de
λ = 633 nm, con la potencia del mismo al 100 %.
2.2.2.4. Análisis termogravimétrico y térmico diferencial (TGA‐DTA)
El análisis termogravimétrico (TGA) se basa en el registro de los cambios de
peso que se producen en un material al someterlo a calentamiento o
enfriamiento a una velocidad conocida y en una atmósfera controlada. Por otro
lado, en el análisis térmico diferencial (DTA) se registran los cambios de
temperatura que tienen lugar durante el tratamiento térmico con respecto a una
sustancia de referencia térmicamente inerte. Cuando se realizan conjuntamente,
estas técnicas permiten determinar pérdidas de humedad, descomposiciones y
transformaciones de fase, información a partir de la cual pueden establecerse las
condiciones óptimas para llevar a cabo el tratamiento térmico de las muestras.
Dado que una serie de los catalizadores empleados en esta investigación se
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prepararon por descomposición térmica de un precursor comercial, se llevó a
cabo el análisis termogravimétrico y térmico diferencial de dicho precursor. El
análisis, realizado por el SACSS‐SAIUEx, se efectuó en una termobalanza
SETSYS Evolution‐16 (Setaram), con una balanza que detecta cambios de peso
con una resolución de 0,04 g. La muestra, normalmente entre 10 y 20 mg, se
introducía en un crisol de alúmina, empleando un caudal de 50 mL min‐1 de aire
e incrementando la temperatura desde temperatura ambiente hasta 600 °C a una
velocidad de calentamiento de 5 °C min‐1.
2.2.2.5. Microscopía electrónica de transmisión (TEM)
La microscopía electrónica de transmisión (TEM) permite obtener
información estructural y morfológica a nivel nanométrico. Mediante el empleo
de esta técnica se pueden identificar las distintas fases existentes en las
partículas del catalizador, su cristalinidad y la dispersión en la que se encuentra
la fase activa. En esta técnica un haz de electrones se focaliza mediante dos
lentes condensadoras sobre una muestra delgada y transparente a los electrones.
Después de atravesar la muestra, los electrones son recogidos y focalizados por
la lente objetivo dentro de una imagen intermedia ampliada que, además, es
aumentada con las lentes proyectoras y, finalmente, proyectada sobre una
pantalla fluorescente o una película fotográfica [18].
Las muestras en polvo se dispersaron en etanol y se depositaron en una
rejilla de cobre recubierta con un polímero orgánico. Una vez evaporado el
alcohol se procedió al análisis en un microscopio electrónico de transmisión
JEM‐2100F (JEOL) y en otro Tecnai G20 Twin (FEI Company). Ambos equipos,
pertenecientes a ICP‐CSIC y SACSS‐SAIUEx, respectivamente, operaban con
una tensión de aceleración de 200 kV.
2.2.2.6. Microscopía electrónica de barrido (SEM)
La microscopía electrónica de barrido (SEM) es una técnica que permite
obtener información estructural y morfológica a nivel micrométrico sobre el
Materiales y métodos experimentales
93
tamaño y la forma de las partículas que constituyen un material sólido.
En un microscopio electrónico de barrido se hace incidir sobre una muestra
gruesa y opaca a los electrones un haz delgado de electrones acelerados, con
energías desde cientos de eV hasta decenas de keV. El haz es desplazado sobre
la superficie de la muestra gracias a un juego de bobinas deflectoras,
rastreándola. Cuando el haz primario entra en contacto con la superficie de la
muestra, parte de los electrones es reflejada y otra parte penetra unas pocas
capas atómicas, siguiendo una trayectoria complicada antes de volver a emerger
a la superficie. La intensidad de estas dos emisiones varía en función del ángulo
que forma el haz incidente con la superficie del material, es decir, depende de la
topografía de la muestra. Las señales emitidas se recogen mediante un detector
y se amplifican. Como resultado se obtiene una imagen topográfica de la
muestra muy ampliada [18].
El equipo utilizado para la adquisición de micrografías de SEM de los
catalizadores, perteneciente al SACSS‐SAIUEx, fue un microscopio electrónico
de barrido de la marca Hitachi S‐4800. Se trabajó con un voltaje de aceleración
de 20‐30 kV y 500 ‐ 2000 de magnificación. Las muestras se dispersaron en un
portamuestras recubierto de carbono y un adhesivo para mantener fijas las
partículas de sólido.
2.2.2.7. Isotermas de adsorción‐desorción de nitrógeno
La isoterma de adsorción de un gas se define como la cantidad de gas
adsorbido en un sólido a distintas presiones relativas de gas, manteniendo
constante la temperatura. Cada punto de la isoterma obtenido representa un
punto de equilibrio entre el volumen del gas adsorbido y la presión relativa del
gas (P/P0). Una vez alcanzadas presiones relativas próximas a la unidad, es
posible determinar la cantidad de adsorbato que permanece retenido para
valores decrecientes de P/P0, conociéndose la curva resultante como isoterma de
desorción. Las curvas de adsorción y desorción no tienen por qué coincidir en
un determinado intervalo de presiones relativas, denominándose a la diferencia
CAPÍTULO 2
94
entre ambas curvas ciclo de histéresis. Para un tipo de sólido, la forma de la
isoterma y del ciclo de histéresis están relacionados con las diferencias en la
energía de interacción entre adsorbato y adsorbente y con la estructura porosa
(micro, meso o macroporos) del sólido. La mayoría de las isotermas que se
pueden encontrar en la bibliografía científica pertenecen a uno de los cinco tipos,
denominados I a V, de la clasificación original de Brunauer, Deming, Deming y
Teller [19] o del tipo VI añadido por la IUPAC [20]. Asimismo se reconocen
cuatro tipos de bucles de histéresis [21].
Las medidas experimentales de adsorción‐desorción de nitrógeno se llevaron
a cabo a ‐196° C en un equipo Autosorb 1 (Quantachrome) perteneciente al
SACSS‐SAIUEx. Las muestras fueron desgasificadas durante 24 horas a 250 °C
con una presión de alto vacío (10‐4 Pa) para asegurar que la superficie estuviera
libre de especies adsorbidas.
Para la determinación de la superficie específica a partir de esta técnica, la
IUPAC recomienda la metodología desarrollada por Brunauer, Emmett y Teller
[22], que desarrollaron la ecuación que hoy se conoce como BET (método BET).
La ecuación BET relaciona el volumen de gas adsorbido a una determinada
presión relativa con el volumen adsorbido en una monocapa de adsorbato sobre
el sólido.
En esta investigación, se calculó el valor de la superficie específica de los
catalizadores a partir de los datos correspondientes a presiones relativas entre
0,02 y 0,15, procesando los mismos con ayuda del software AS1winTM 2.01. El
software también permite determinar el volumen de microporos a partir de la
rama de adsorción de la isoterma mediante el método “t”, aplicando la ecuación
de Halsey [23,24]. En esta determinación, de la representación del volumen
adsorbido frente al parámetro t se obtiene una recta cuya pendiente se relaciona
con el área externa no asociada a los microporos. De esta forma, el área
microporosa se obtuvo por diferencia entre el área BET y el área externa.
Finalmente, el volumen de mesoporos se determinó a partir de la cantidad de
Materiales y métodos experimentales
95
nitrógeno adsorbido a la presión relativa de 0,96 en la rama de desorción de la
isoterma, zona equivalente al llenado de todos los poros de diámetro hasta 50
nm, restando al valor obtenido el volumen de microporos calculado por el
método “t”.
2.2.2.8. Determinación del pH del potencial de carga cero (pHPZC) por valoración másica
Para las reacciones en medio acuoso, una medida que da una buena
indicación acerca del carácter ácido o básico superficial de un sólido es la
determinación del pH de la suspensión acuosa del mismo (en terminología
anglosajona “pH slurry”), el cual se relaciona con el pH del potencial de carga
cero (pHPZC) dando una buena indicación acerca de la carga electrónica
superficial del material [25].
Para la estimación del pHPZC de distintas muestras de catalizador se empleó
el método de valoración másica propuesto por Subramanian et al. en 1988 [26].
Para ello, se determinó el pH de la disolución acuosa resultante de mantener
una suspensión de sólido al 5 % en peso en agua ultrapura, con agitación
continua y en un recipiente cerrado. Transcurridas 24 horas se filtró la muestra y
se midió el pH del agua resultante con un pH‐metro Crison GLP 21+. Este
procedimiento se repitió dejando la muestra en agitación hasta que el pH
medido alcanzó un valor constante.
2.2.2.9. Espectroscopía fotoelectrónica de Rayos X (XPS)
La espectroscopía fotoelectrónica de rayos X (XPS) es una de las técnicas de
análisis químico englobadas bajo el nombre de ESCA (Electron Spectroscopy for
Chemical Analysis), utilizada para determinar el estado químico y la
composición superficial de materiales sólidos. Está basada en el efecto
fotoeléctrico: al irradiar una muestra con fotones de energía superior a la energía
de enlace de los electrones de sus átomos, dichos electrones salen del sólido con
una energía cinética igual a la diferencia de energía entre el fotón y la energía de
enlace del electrón. En este sentido, aunque los rayos X utilizados pueden
CAPÍTULO 2
96
penetrar unas pocas micras en una muestra sólida, solo los electrones generados
a unos pocos nanómetros de la superficie pueden salir del sólido. Esto se debe a
que los fotoelectrones producidos en las capas más internas sufren colisiones
inelásticas que provocan una pérdida de energía tal que no pueden abandonar la
superficie del material. Por tanto, esta técnica es especialmente sensible para
análisis químico superficial, proporcionando información química de las 5 ‐ 10
primeras capas atómicas del sólido.
El espectro de XPS es la representación del número de electrones detectados
en un intervalo de energías frente a su energía cinética o, más comúnmente,
frente a su energía de enlace. Las variaciones de energía de enlace de un
elemento respecto a su estado no combinado se deben a las diferencias en el
potencial químico y en la polaridad de los compuestos, por lo que pueden
usarse para identificar el estado químico de los elementos analizados dado que
cada elemento tiene un conjunto de energías de enlace características [27,28].
El análisis de muestras de catalizadores por espectroscopía fotoelectrónica de
rayos X fue realizado por el SACSS‐SAIUEx. Los espectros XPS se obtuvieron
utilizando un espectrómetro de fotoelectrones K‐ de Thermo Scientific
equipado con fuente de rayos X de Al Kα (hν = 1253,6 eV), que operaba a un
voltaje de 12 kV bajo vacío (2x10‐7 mbar). Se realizaron espectros de detalle de
los elementos más importantes en las muestras y, para eliminar el efecto de
carga, las energías de enlace se corrigieron utilizando como patrón el pico C1s
(Energía de enlace 284,6 eV).
En todos los casos el espectro XPS se resolvió en varias componentes,
ajustándose cada componente, mediante el programa de tratamiento de
espectros XPSPEAK 4.0, a la curva experimental con una combinación lineal de
curvas lorentzianas y gaussianas en proporciones variables y escogiendo el
mejor ajuste mediante minimización de residuos χ2. Las relaciones atómicas de
los elementos en superficie se obtuvieron a partir de la siguiente ecuación [29]:
Materiales y métodos experimentales
97
i i i
j j j
n I / S
n I / S (2.10)
donde ni es el número de átomos por cm3 del elemento i, Ii es la intensidad del
pico (generalmente expresada como área) y Si es su factor de sensibilidad
atómica.
2.2.2.10. Espectroscopía ultravioleta‐visible de reflectancia difusa (DR‐UV‐Vis)
La espectroscopía UV‐Vis se fundamenta en la absorción electrónica de la
radiación electromagnética cuando interacciona con la materia en el entorno de
longitudes de onda entre 190 nm y 800 nm. En el caso de los catalizadores
estudiados en este trabajo (semiconductores), se observa fundamentalmente la
transición de electrones desde la banda de valencia a la banda de conducción,
cuya energía corresponde a esta región del espectro electromagnético.
La medida de la reflectancia difusa se define como la fracción de radiación
incidente que es reflejada por la muestra en todas direcciones. Para ello, se
emplea un dispositivo llamado esfera integradora, consistente en una esfera
hueca recubierta en su interior de un material altamente reflectante, que envía la
luz reflejada por la muestra al detector [30]. El espectro resultante se obtiene
como tanto por ciento de reflectancia frente a la longitud de onda, fijando como
100 % de reflectancia la obtenida para una muestra de referencia que no absorba
luz en el rango de longitudes de onda utilizado (generalmente BaSO4).
Cerca del borde de absorción se ha establecido que:
ngh A h E (2.11)
siendo h la constante de Planck, equivalente a 4,136x10‐15 eV s; A una constante;
Eg la energía del salto de banda en eV; la frecuencia de la radiación en s‐1; el
exponente n que toma el valor de ½ para transiciones electrónicas directas; y el
coeficiente de absorción [31]. Este coeficiente se determina como:
‐ln R (2.12)
CAPÍTULO 2
98
donde R es el valor de reflectancia medida respecto a la unidad.
De esta manera, reordenando la ecuación (2.11) se tiene que:
2gh A h E (2.13)
Si se representa el primer término de la ecuación (h)2 frente a h se
determina el valor del salto de banda, Eg, como el corte en el eje h cuando α es
igual a cero.
Los análisis de espectroscopía UV‐visible de reflectancia difusa fueron
realizados por el personal de la Unidad de Apoyo a la Investigación del ICP‐
CSIC, en un equipo UV‐Vis‐NIR Cary 500 (Varian‐Agilent Technologies)
equipado con esfera integradora, registrando espectros de absorbancia y
porcentajes de reflectancia para valores de longitudes de onda comprendidos
entre 200 y 900 nm a intervalos de 1 nm.
2.3. REACTIVOS EMPLEADOS
En la Tabla 2.4 se muestran, clasificados en función de su aplicación, los
productos químicos empleados a lo largo de esta investigación, indicándose la
pureza de los mismos y la casa comercial que los suministró.
En todos los casos en los que fue necesaria la preparación de disoluciones
acuosas, se empleó agua ultrapura (grado de pureza milliQ‐Millipore).
Materiales y métodos experimentales
99
Tabla 2.4. Reactivos empleados.
Reactivo Pureza Fabricante Aplicación
Acetaminofeno (C8H9NO2) 99 % Sigma‐Aldrich
Compuestos
modelo y matriz
para estudios de
procesos de
degradación en
agua
Agua residual urbana EDAR Badajoz
Antipirina (C11H12N2O) 99 % Sigma‐Aldrich
Cafeína (C8H10N4O2) 99 % Sigma‐Aldrich
Carbamazepina (C15H12N2O) > 98 % Sigma‐Aldrich
Diclofenaco de sodio
(C14H10Cl2NaO2) 99 % Sigma‐Aldrich
Hidroclorotiazida (C7H8ClN3O4S2) 99 % Sigma‐Aldrich
Ibuprofeno sódico (C13H17NaO2) > 98 % Sigma‐Aldrich
Ketorolaco sal tris
(C15H13NO3∙C4H11NO3) > 99 % Sigma‐Aldrich
Metanol (CH3OH) PAI‐ACS
(grado HPLC)
Panreac
Química
Metoprolol tartrato
(C15H25NO3)2∙C4H6O6) > 99 % Sigma‐Aldrich
N,N‐Dietil‐meta‐toluamida
(C12H17NO) 97 % Sigma‐Aldrich
Sulfametoxazol (C10H11N3O3S) > 98 % Fluka
Biochemika
Ácido clorhídrico (HCl) 36,7 % Fisher
Chemical
Ajuste de pH
Ácido ortofosfórico (H3PO4) 85 % Panreac
Química
Ácido perclórico (HClO4) 70 % Panreac
Química
Hidróxido sódico (NaOH) 98 % Panreac
Química
CAPÍTULO 2
100
Tabla 2.4 (Continuación). Reactivos empleados.
Reactivo Pureza Fabricante Aplicación
Ácido cítrico (C6H8O7) 99 % Panreac
Química
Síntesis de
catalizadores
Ácido clorhídrico (HCl) 36,7 % Fisher
Chemical
Ácido nítrico (HNO3) 65 % Panreac
Química
Ácido oxálico dihidratado
(C2H2O4∙2H2O) 99,5 % Sigma‐Aldrich
Ácido wolfrámico (H2WO4) > 99 % Sigma‐Aldrich
Cloruro cálcico (CaCl2) 99 % Panreac
Química
Dióxido de titanio P25 (TiO2) > 99 % Evonik‐
Degussa AG
Disolución de amoníaco (NH4OH) 35 % Fisher
Chemical
Etanol (C2H6O) 99,5 % Panreac
Química
Hidróxido sódico (NaOH) 98 % Panreac
Química
Nitrato de cerio (III)
hexahidratado (Ce(NO3)3∙6H2O) 99 % Sigma‐Aldrich
Trióxido de wolframio (WO3) > 99 % Sigma‐Aldrich
Wolframato cálcico (CaWO4) 99,9 % Sigma‐Aldrich
Wolframato sódico (Na2WO4) > 99 % Sigma‐Aldrich
Oxígeno comprimido (O2) > 99,999 % Abelló Linde Agente oxidante y
obtención de ozono
Acetonitrilo (C2H3N) PAI‐ACS
(grado HPLC)
Panreac
Química
Determinación de
la concentración de
contaminantes
modelo en agua e
identificación de
intermedios de
degradación
Ácido fórmico (CH2O2) 98 % Panreac
Química
Ácido ortofosfórico (H3PO4) 85 % Panreac
Química
Materiales y métodos experimentales
101
Tabla 2.4 (Continuación). Reactivos empleados.
Reactivo Pureza Fabricante Aplicación
Ácido ortofosfórico (H3PO4) 85 % Panreac
Química
Determinación de
la concentración
de ozono en
disolución acuosa
Fosfato diácido de potasio
(KH2PO4) > 98 %
Panreac
Química
Índigo carmín (5,5’,7‐
indigotrisulfonato de potasio, C16H8N2Na2O8S2)
98 % Sigma‐Aldrich
Bicarbonato de sodio (NaHCO3) 99,7 % Panreac
Química Determinación de
la concentración
de peróxido de
hidrógeno en
disolución
Calgón (Polifosfato de sodio,
((NaPO3)n) 65 ‐ 70 %
Panreac
Química
Cloruro de cobalto (II)
hexahidratado (CoCl2∙6H2O) 99 %
Panreac
Química
Acetato de amonio (CH3CO2NH4) 98 % Sigma‐Aldrich
Determinación de
la concentración
de formaldehído
en disolución
Acetilacetona
(CH3COCH2COCH3) 99 % Sigma‐Aldrich
Ácido acético glacial (CH3COOH) 100 % Merck
Formaldehído (CH2O) 37 % Sigma‐Aldrich
Ácido acético glacial (CH3COOH) 100 % Merck
Determinación de
la intensidad de
radiación y de las
concentraciones de
Fe(II) y Fe total en
disolución
Ácido oxálico dihidratado
(C2H2O4∙2H2O) 99,5 % Sigma‐Aldrich
Ácido perclórico (HClO4) 70 % Panreac
Química
Fluoruro de amonio (NH4F) 98 % Sigma‐Aldrich
Nitrógeno comprimido (N2) > 99,999 % Abelló Linde
o‐Fenantrolina (C12H8N2∙H2O) 99 % Fluka
Biochemika
Perclorato de hierro(III) hidratado
(Cl3FeO12∙xH2O) PA‐ISO Sigma‐Aldrich
Reactivo para análisis fotométrico
de hierro, Spectroquant ‐ Merck
CAPÍTULO 2
102
Tabla 2.4 (Continuación). Reactivos empleados.
Reactivo Pureza Fabricante Aplicación
Ácido clorhídrico (HCl) 36,7 % Fisher
Chemical
Determinación de
la concentración
de carbono
orgánico total en
disolución
Ácido ortofosfórico (H3PO4) 85 % Panreac
Química
Aire sintético comprimido (O2/N2) 21 ± 0,5 % O2 Abelló Linde
Bicarbonato de sodio (NaHCO3) PA‐ISO Nacalai Tesque
Carbonato de sodio (Na2CO3) PA‐ISO Nacalai Tesque
Ftalato ácido de potasio
(KHC8H4O4) PA‐ISO Nacalai Tesque
Ácido acético glacial (CH3COOH) 100 % Merck
Determinación de
la concentración
de algunos ácidos
orgánicos e iones
inorgánicos en
disolución
Ácido fórmico (CH2O2) 98 % Panreac
Química
Ácido oxálico dihidratado
(C2H2O4∙2H2O) 99,5 % Sigma‐Aldrich
Ácido sulfúrico (H2SO4) 95 ‐ 98 % Panreac
Química
Carbonato de sodio (Na2CO3) > 99,8 % Sigma‐Aldrich
Cloruro de sodio (NaCl) 99 % Panreac
Química
Fosfato de potasio dibásico
(K2HPO4) > 98 %
Panreac
Química
Nitrato de sodio (NaNO3) 99 % Panreac
Química
Piruvato de sodio (C3H3NaO3) > 99 % Sigma‐Aldrich
Sulfato de sodio (Na2SO4) 99 % Panreac
Química
Ácido oxálico dihidratado
(C2H2O4∙2H2O) 99,5 % Sigma‐Aldrich
Determinación de
parámetros
cinéticos Ácido p‐clorobenzoico (C7H5ClO2) > 98 % Merck
ter‐butanol(C4H10O) 99 % Panreac
Química
Materiales y métodos experimentales
103
Tabla 2.4 (Continuación). Reactivos empleados.
Reactivo Pureza Fabricante Aplicación
Cubetas test LCK 349 ‐ Hach‐Lange
Determinación de
la concentración
de fosfatos en
disolución
Cubetas test LCK 414 ‐ Hach‐Lange
Determinación de
la demanda
química de
oxígeno
Aliltiourea (C4H8N2O) > 98 % Panreac
Química Determinación de
la demanda
biológica de
oxígeno
Hidróxido sódico (NaOH) 98 % Panreac
Química
Microorganismos aerobios ‐ EDAR Badajoz
BIBLIOGRAFÍA
[1] Bader, H.; Hoigné, J. “Determination of ozone in water by the indigo
method”. Water Res. 15 (1981) 449‐456.
[2] Masschelein, W.; Denis, M.; Ledent, R. “Spectrophotometric determination of
residual hydrogen peroxide”. Water & Sewage Works (1977) 69‐72.
[3] Stookey, L.L. “Ferrozine‐ A new spectrophotometric reagent for iron”. Anal.
Chem. 42 (1970) 779‐781.
[4] Zuo, Y. “Kinetics of photochemical/chemical cycling of iron coupled with
organic substances in cloud and fog droplets”. Geochim. Cosmochim. Act. 59
(1995) 3123‐3130.
[5] Hatchard, C.G.; Parker, C.A. “A new sensitive chemical actinometer. II.
Potassium ferrioxalate as a standard actinometer”. Proceedings of the Royal
Society, London Service A 235 (1956) 518‐536.
[6] Nash, T. “The colorimetric estimation of formaldehyde by means of the
Hantzsch reaction”. Biochemistry 55 (1953) 416‐421.
[7] Moore, W.A.; Krorner, R.C.; Ruchchoff, C.C. “Dichromate reflux method for
determination of oxygen consumed”. Anal. Chem. 21 (1948) 953‐957.
[8] Rodríguez, E.; Mimbrero, M.; Masa, F.J.; Beltrán, F.J. “Homogeneous iron‐
catalyzed photochemical degradation of muconic acid in water”. Water Res. 41
CAPÍTULO 2
104
(2007) 1325‐1333.
[9] Sandell, E.B. “Colorimetric determination of traces of metals”. Interscience
Pubs., New York 1959.
[10] Goldstein, S.; Rabani, J. “The ferrioxalate and iodide‐iodate actinometers in
the UV region”. J. Photochem. Photobiol. A Chem. 193 (2008) 50‐55.
[11] Safarzadeh‐Amiri, A.; Bolton, J.; Cater, S. “Ferrioxalate‐mediated
photodegradation of organic pollutants in contaminated water”. Water Res. 31
(1997) 787‐798.
[12] Faraldos, M. en “Técnicas de Análisis y Caracterización de Materiales”.
Faraldos, M.; Goberna, C. (eds.), CSIC, Madrid (2003), cap. 9.
[13] Murcia‐Mascarós, S. en “Técnicas de Análisis y Caracterización de
Materiales”. Faraldos, M.; Goberna, C. (eds.), CSIC, Madrid (2003), cap. 10.
[14] Klug, H.P.; Alexander, L.E. “X‐ray diffraction procedures for polycrystalline
and amorphous materials”. 2nd. Edition, John Wiley & Sons (1974).
[15] www.icdd.com (consultada en enero de 2017).
[16] Raman, C.V.; Krishnan, K.S. “A new type of secondary radiation”, Nature
121 (1928) 501‐502.
[17] Bañares, M.A.; Valenzuela, R.X. en “Técnicas de Análisis y Caracterización
de Materiales”. Faraldos, M.; Goberna, C. (eds.), CSIC, Madrid (2003), cap. 5.
[18] Díaz, I.; Landa, A.R.; Otero, L.C. en “Técnicas de Análisis y Caracterización
de Materiales”. Faraldos, M.; Goberna, C. (eds.), CSIC, Madrid (2003), cap. 11.
[19] Brunauer, S.; Deming, L.S.; Deming, W.S.; Teller, E. “On a theory of the van
der Waals adsorption of gases”. J. Am. Chem. Soc. 62 (1940) 1723‐1732.
[20] Haber, J. “Manual on catalyst characterization”. Pure Appl. Chem. 63 (1991)
1227‐1246.
[21] Sing, K.S.W.; Everett, D.H.; Haul, R.A.W.; Moscou, L.; Pierotti, R.A.;
Rouquerol, J.; Siemieniewska, T. “Reporting physisorption data for gas/solid
systems with special reference to the determination of surface area and
porosity”. Pure Appl. Chem. 57 (1985) 603‐619.
[22] Brunauer, S.; Emmett, P.; Teller, E. “Adsorption of gases in multimolecular
Layers”. J. Am. Chem. Soc. 60 (1938) 309‐319.
[23] Halsey, C.P. “Physical adsorption on non‐uniform surfaces”. J. Chem. Phys.
16 (1948) 931‐937.
[24] Lippens, B.C.; Linsen, B.G.; de Boer, J.H. “Studies on pore systems in
catalysts I. The adsorption of nitrogen; apparatus and calculation”. J. Catal. 3
Materiales y métodos experimentales
105
(1964) 32‐37.
[25] Menéndez, J.A.; Illán‐Gómez, M.J.; León y León, C.A.; Radovic, L.R. “On the
difference between the isoelectric point and the point of zero charge of carbons”.
Carbon 33 (1995) 1655‐1657.
[26] Subramanian, S., Noh, J.S., Schwarz, J.A. “Determination of the point of
zero‐charge of composite oxides”. J. Catal. 114 (1988) 433‐439.
[27] Moulder, J.F.; Stickle, W.F.; Sobol, P.E.; Bomben, K.D. Handbook of X‐Ray
Photoelectron Spectroscopy: A Reference Book of Standard Spectra for
Identification and Interpretation of XPS Data (Hardcover), Physical Electronics,
Reissue ed. (1995).
[28] Campos, J.M. en “Técnicas de Análisis y Caracterización de Materiales
Faraldos, M.; Goberna, C. (eds.), CSIC, Madrid (2003), cap. 12.
[29] Wagner, C.D.; Davis, L.E.; Zeller, M.V.; Taylor, J.A.; Raymond, R.H.; Gale,
L.H. “Empirical atomic sensitivity factors for quantitative analysis by electron
spectroscopy for chemical analysis”. Surf. Interf. Anal. 3 (1981) 211‐225.
[30] Springsteen, A. en “Applied Spectroscopy”, Workman, J.; Springsteen, A.
(eds.), Academic Press, San Diego (1998), cap.6.
[31] Serpone, N.; Lawless, D.; Khairutdinovt R. “Size effects on the
photophysical properties of colloidal anatase TiO2 particles: size quantization or
direct transitions in this indirect semiconductor?” J. Phys. Chem. 99 (1995)
16646‐16654.
CAPÍTULO 3 (CHAPTER 3) PAPER 1: On ozone-photocatalysis synergism in black-light induced reactions: Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
E. Mena, A. Rey, B. Acedo, F.J. Beltrán, S. Malato
Chemical Engineering Journal 204-206 (2012) 131-140
ABSTRACT. The synergism produced between ozone and TiO2 black light photocatalytic oxidation of methanol has been studied following the rate of formaldehyde formation during photocatalytic oxidation, ozonation and photocatalytic ozonation experiments. Methanol was selected as a model compound due to its low reaction rate with molecular ozone and its scavenging character for both, free hydroxyl radicals and trapped holes. TiO2-P25 was used as photocatalyst and black light blue lamps (emitting with a maximum at 365 nm) as radiation source. The effect of ozone concentration and pH was evaluated. Absorbed light intensity by the photocatalyst was also determined to calculate the quantum yields of photocatalytic reactions. Three main processes need to be considered during photocatalytic ozonation: direct ozone-methanol reaction, indirect ozone reactions and photocatalytic reactions, which allow calculating the quantum yield of photo-generated oxidizing species. The presence of ozone exerts a positive effect in the reaction rate of oxidizing species formation due to light induced reactions also enhancing the quantum yield from 0.34 to 0.80 mol einstein-1 at pH = 3 (where indirect ozone reactions are negligible). This parameter increased from 0.29 to 3.27 mol einstein-1 at pH = 7 likely due to indirect ozone reactions that cannot be disregarded. The positive effect of ozone in the photocatalytic induced reactions has been attributed to the reaction of dissolved ozone and hydrogen peroxide (formed upon methanol direct ozonation) as electron acceptors, thus reducing the recombination process on the catalyst surface to some extent. A simplified economic study is also presented.
Keywords: Photocatalytic ozonation, ozone, TiO2, synergism, methanol, formaldehyde.
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
109
3.1. INTRODUCTION
Wastewater is one of the most important problems in industrialized regions.
Many of these effluents contain high concentrations of organic pollutants that
must be removed in order to fulfill the required limits of the increasingly
restrictive legislation. In this context, photocatalytic detoxification treatments
based on TiO2 semiconductor have been the focus of numerous investigations in
the last 30 years for the destruction of undesirable contaminants in water [1‐3].
In heterogeneous photocatalysis, photoinduced holes and electrons in
semiconductor particles, through a complex reaction mechanism, give place to
highly oxidizing species which play a key role in degradation of organic
pollutants. The most commonly used semiconductor has been polycrystalline
powders of titanium dioxide due to unique properties such as chemical stability,
safety and low cost [2]. However, its photohole‐electron recombination is a
serious problem for the development of photocatalytically based technologies
since it severely limits the quantum yields achievable [2‐4]. Several strategies
have been proposed to minimize this problem and increase the process
efficiency (ion doping, different semiconductors coupling, using chemical
oxidants or combining photocatalysis with other advanced oxidation processes
(AOPs) [2]). Among them, the combination of ozone and heterogeneous
photocatalysis with TiO2 (photocatalytic ozonation) has demonstrated to be an
efficient treatment enhancing the formation of oxidizing species compared to the
single ozonation or photocatalytic processes [5,6].
When a semiconductor (e.g. TiO2) is irradiated with a photon of energy
greater than its band gap energy an electron/hole pair is formed in the
conduction band (CB) and valence band (VB), respectively. These mobile species
can migrate to the TiO2 surface and/or can be readily trapped forming less
mobile states (to simplify e‐/h+ stand for all the forms of holes and electrons).
2TiO h e h (3.1)
In addition, these species can give place to the recombination reaction (3.2):
CAPÍTULO 3 (CHAPTER 3)
110
2e h TiO (3.2)
The nature of trapped holes has been controversially discussed [7,8]. It has
been generally assumed that adsorbed water could be photo‐oxidized giving
place to surface bounded hydroxyl radicals ( sHO ). However, it has been
reported that this reaction was kinetic and thermodynamically hindered [7] and
the trapping phenomena by terminal oxygen ions of the TiO2 lattice (O ‐22TiO ) has
been proposed, forming terminal protonated or deprotonated radicals
(depending on pH) [9,10]. Regardless of the reaction considered, the formation
of hydroxyl radicals in the TiO2 surface can be described as:
2‐
2 TiO2H O/O
‐s sh HO (O ) (3.3)
In addition, oxygen when present on the particle surface, acts as an electron
acceptor according to reaction (3.4):
•
2 2e O O ‐ (3.4)
Superoxide ion radicals ( •‐2O ) may give place to hydrogen peroxide that can
react with TiO2 electrons and/or additional •‐2O , forming free hydroxyl radicals
through reactions (3.5) and (3.6):
2 2 22 2O O 2H H O O‐ ‐ (3.5)
2 2 2 2e O H O HO HO O (3.6)
In this complex (but simplified) reaction mechanism, free hydroxyl radicals,
•HO , and/or trapped holes ( s sh HO / O ) may be responsible of the non‐
selectively organic matter oxidation and mineralization (reaction (3.7)), being
this process among the most studied AOPs.
2 2HO (h ) R R ʹ HO (h ) ... CO H O (3.7)
When ozone is present, it can react directly and selectively with some organic
compounds (i.e. aromatic and substituted aromatic compounds, molecules with
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
111
unsaturated bonds e.g. ‐C=C‐, ‐C≡C‐, ‐C=N, ‐C=O, etc.) through different
reaction mechanisms (mainly cycloaddition reactions and/or electrophilic
substitution) included in ‘‘direct ozone reactions’’ terminology (reaction (3.8))
[11].
3 2 2zO R R ʹ H O (3.8)
On the other hand, the presence of radical species (mainly •HO ) coming from
the decomposition of ozone in water gives place to the non‐selectively oxidation
of the organic compounds in water (indirect ozone reactions), process that is
favored in alkaline media [11]:
3 2 2O HO HO O (3.9)
•• ‐3 2 2 3O HO HO O (3.10)
••pKa 4.8 ‐
2 2HO O H (3.11)
• •‐ ‐3 22 3O O O O (3.12)
The generated ozonide radical ( •‐3O ) rapidly reacts with H+ in the solution to
give •
3HO radical, which evolves to give O2 and •HO (reactions (3.13) and (3.14)).
• •‐3 3O H HO (3.13)
•
23HO HO O (3.14)
The benefits of using the combined process, photocatalytic ozonation, are not
only related to the sum of individual processes but also to the fact that the
dissolved ozone can readily react with electrons at the TiO2 surface according to
reaction (3.15) giving place to the ozonide ion radical:
•‐3 3O e O (3.15)
As a consequence, the recombination of electrons and positive holes may be
reduced by this reaction (3.15) and also by reaction (3.6) due to the presence of
CAPÍTULO 3 (CHAPTER 3)
112
higher H2O2 concentration in the reaction medium, eventually formed during
direct ozonation reactions (reaction (3.8)) [12‐14]. With this reaction scheme a
synergistic effect between ozone and semiconductor photocatalysis is expected
due to the larger amount of oxidizing species formed (hydroxyl radicals,
bounded hydroxyl radicals and/or positive holes). This has been observed for
the photocatalytic ozonation of several organic compounds in water [2,5,6,15‐
19].
Some of the target compounds subjected to photocatalytic ozonation were
complex molecules that present high direct ozone‐organic rate constants and/or
also complicated reaction pathways involving several steps with different
intermediate species [5,16,19]. This makes difficult to analyze the reaction
mechanism, the species involved in the oxidation steps and/or the synergistic
effect between ozone and photocatalysis. In that sense, it would be interesting to
test the photocatalytic ozonation process using small and refractory to direct
ozone reactions organic compounds.
In this work methanol has been selected due to its known scavenging
character of hydroxyl radicals (kHO = 9.7x108 M‐1 s‐1, [20]) and because it has been
commonly used to test the photocatalytic efficiency of several TiO2 materials
since it also reacts with photogenerated holes [9,10,21‐25]. In addition,
methanol‐ozone direct reaction takes place at slow reaction rate (kO3 = 0.024 M‐1
s‐1) [26]. Therefore, it is expected that under appropriate experimental
conditions, the formaldehyde evolution (formaldehyde is the first product of
methanol oxidation) gives information about the production rate of oxidizing
species (different from O3). The aim of this work was then to evaluate the
production of photo‐generated oxidizing species (hydroxyl radicals and positive
holes) using methanol as target compound comparing both, photocatalytic
oxidation and photocatalytic ozonation, to determine the true quantum yield of
photocatalytic reactions, and to state the synergy degree between ozone and
irradiated semiconductor during photocatalytic ozonation. To our knowledge
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
113
the determination of the quantum yield of photo‐generated oxidizing species
production in photocatalytic ozonation has not been studied before with any
target compound.
3.2. EXPERIMENTAL SECTION
3.2.1. Experimental set‐up and oxidation/ozonation procedure
Ozonation, photocatalytic oxidation and photocatalytic ozonation
experiments were carried out in a 1 L slurry cylindrical reactor equipped with
magnetic stirring and inlets for measuring temperature, feeding the gas (oxygen
or ozone‐oxygen) through a porous plate situated at the reactor bottom,
sampling port and outlet for the non‐absorbed gas. In photocatalytic
experiments, the reactor was illuminated with 2 black light blue (UVA radiation)
fluorescent lamps (15 W each, from HQPower) placed inside a black box.
The reactor was charged with an aqueous solution containing methanol (2 M,
CH3OH HPLC grade from Panreac) and TiO2 (0.5 g L‐1, Degussa P25, in catalytic
experiments). Initial pH was set to 3 with HClO4, pH = 7 with NaOH or buffered
at pH = 7 with H3PO4 (30 mM) and NaOH (from Panreac). In photocatalytic
experiments, the reactor was then exposed to the radiation (lamps were turned
on 30 min before to stabilize). In ozonation experiments a mixture of ozone‐
oxygen gas (30 L h‐1, 10 ‐ 30 mg L‐1) was also continuously fed to the reactor.
Ozone was generated from pure oxygen in a Sander laboratory ozonator.
Temperature was maintained at 25 °C during the reaction time. Reaction
samples were withdrawn from the reactor at regular intervals for 60 min
reaction time, and then filtered through syringe PET membrane filters
(Chromafil Xtra, 0.20 m). The evolution of the reaction was followed through
the determination of formaldehyde (primary product of methanol oxidation),
hydrogen peroxide, dissolved ozone concentration, ozone concentration in the
outlet gas and pH.
Formaldehyde was determined by the Nash method [27], based on the
CAPÍTULO 3 (CHAPTER 3)
114
Hantzsch reaction. In this assay, 2 mL of reagent (0.2 mL of acetylacetone
(Sigma‐Aldrich), 3 mL of acetic acid (Panreac) and 25 g of ammonium acetate
(Fluka) in 100 mL of water) are mixed with 5 mL of the sample and heated for 30
min at 50 °C in the dark [28]. Spectrophotometric measurements were carried
out at 412 nm (= 7890 M‐1 cm‐1) using a Helios‐ Thermo Spectronic
spectrophotometer. Hydrogen peroxide concentration was determined through
the cobalt/bicarbonate method [29], at 260 nm (= 26645 M‐1 cm‐1) using an
Helios‐ spectrophotometer. Dissolved ozone concentration was measured by
following the method proposed by Bader and Hoigné [30] based on the
discoloration of a 5,5,7 indigotrisulphonate solution (= 600 nm, Helios‐
spectrophotometer, = 20000 M‐1 cm‐1). Ozone in the gas phase was monitored
by means of an Anseros Ozomat ozone analyzer, based on the absorbance at 254
nm.
3.2.2. Photon fluxes determination
Ferrioxalate actinometry [31] was used to determine the incident photon flux,
I0, in the photoreactor, that was found to be 2.66x10‐5 einstein min‐1. In these
experiments the Fe(II) concentration was followed by the ‐fenantroline method
[32] using an Helios‐ spectrophotometer at 510 nm (= 11023 M‐1 cm‐1). The
photon flux absorbed by the catalyst, Ia, was estimated through the
determination of the quantum yield of formaldehyde generation in the
photoreactor. This was obtained by applying the protocol of Serpone and
Salinaro [33] for the photocatalytic oxidation of methanol, analyzing the
formaldehyde evolution under the following operating conditions: methanol
concentration from 0.5 to 2 M, TiO2 concentration from 0.025 to 3 g L‐1, pH0 = 7,
pH = 7 (buffered) and pH0 = 3 (adjusted with NaOH, phosphate buffered or
adjusted with HClO4, respectively) and oxygen saturated solution with 30 L h‐1
gas flow rate. The evolution of the reaction was followed as explained in the
previous section. For comparative purposes, additional photocatalytic oxidation
experiments were carried out using formic acid (0.01 M CH2O2) as target
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
115
compound (0.5 g L‐1 TiO2, pH0 = 3). Formic acid was determined following the
evolution of the CO2 formed by measuring the total organic carbon (TOC)
remaining in the solution using a TOC‐VSCH analyzer from Shimadzu.
3.3. RESULTS AND DISCUSSION
3.3.1. Absorbed photon flux
The determination of the absorbed photon flux by the catalyst is a key issue
in heterogeneous photocatalytic reactions that serves to calculate the quantum
yield of the reaction according to the following equation:
ii
a
rreacted molecules mol i
absorbed photons I einstein
(3.16)
The radiant energy used in a photocatalytic reaction (absorbed, Ia) is
generally lower than that impinging on the reacting system (I0) due to the light
scattered or reflected by the suspended catalyst in the dispersion. The absorbed
light intensity has been usually calculated by applying the radiation transfer
equation (RTE) to the reaction system [34‐38] or experimentally determined by
means of the measurement of the light transmission through a photocatalyst
suspension [39].
In this work, we have indirectly calculated the absorbed photon flux by using
the protocol of Serpone and Salinaro [33] to determine the quantum yield of the
methanol oxidation reaction (formaldehyde formation) under our experimental
conditions.
Methanol photocatalytic oxidation gives place to formaldehyde as primary
oxidation product according to reactions (3.17) and (3.18):
3 2 2CH OH h HO CH OH H H O (3.17)
• •
2 2 2 2 2 22CH OH O O CH OH CH O H O (HO ) (3.18)
CAPÍTULO 3 (CHAPTER 3)
116
where either holes (h+) and free hydroxyl radicals ( •HO ) may participate,
forming also superoxide ion radical ( •‐2O ) or hydroperoxide radical ( •
2HO , acidic
pH). Figure 3.1 shows the time‐evolution of CH2O formation under different
experimental conditions. It was nearly linear throughout the range of conditions
used. The rate of CH2O formation reached a plateau around 2 M methanol, as
can be seen in Figure 3.2. This concentration was then used for the experiments
with different TiO2 concentration. At this CH3OH concentration, formaldehyde
is not expected to be significantly oxidized since oxidizing species will mainly
react with the former. In fact, taking into account the kinetic constants of
methanol and formaldehyde with hydroxyl radicals (9.7x108 and 1x109 M‐1 s‐1
[20], respectively) the formaldehyde formation rate with 2 M methanol was 1940
times higher than formaldehyde oxidation rate (using a generic concentration
about 10‐3 M similar to CH2O concentrations observed in this work) proving that
formaldehyde oxidation could be negligible in the overall reaction rate. The
calculated reaction rates of formaldehyde formation at different TiO2 loading
defines a plateau from 0.5 g L‐1 as depicted in Figure 3.3(A). This behavior can be
well described through the following equation:
2 2
2
lim TiO CH O
lim TiO 0
a C r
a C I
(3.19)
where photonic efficiencies are calculated employing the relationship
= rCH2O/I0. The quantum yield of formaldehyde formation (CH2O) was
determined from the limiting photonic efficiency (lim) at high TiO2 loadings. At
these conditions it has been reported that lim = [33]. Figure 3.3(B) shows fitting
results of linearized Eq. (3.19). The quantum yield calculated for formaldehyde
formation at pH0 = 7 (non‐buffered) was CH2O‐pH7 = 0.48 mol einstein‐1. The pH
value did not change significantly (pHf = 6.7). The corresponding absorbed light
intensity at different TiO2 loadings are summarized in Table 3.1 together with
the integrated absorption fraction over the wavelength used here, FS = Ia/I0. From
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
117
0.5 g L‐1 TiO2 onward, the P25 catalyst presented an absorption fraction higher
than 90 %. On the other hand, in experiments carried out at pH = 7 (phosphate
buffered solution), the reaction rate of formaldehyde formation was lower than
in the previous case at the same operating conditions (catalyst loading, methanol
concentration) as can also be observed in Fig. 3.1 and Table 3.1. This can be due
to phosphate adsorption onto the titania surface [40,41], that can compete for
adsorption sites with methanol and/or react with oxidizing species formed at the
catalyst surface (e.g. with hydroxyl radicals, kHO‐PO43‐ = 1x107 M‐1 s‐1; kHO‐HPO42‐ =
1.5x105 M‐1 s‐1; kHO‐H2PO4‐ = 2x104 M‐1 s‐1; kHO‐H3PO4 = 2.7x106 M‐1 s‐1 [20]), although
this second point could be disregarded due to the great excess of methanol (2 M
CH3OH against 30 mM H3PO4). Since phosphates do not absorb the black light
used, it is expected that the absorbed light intensity, Ia, do not change
significantly compared to the non‐buffered experiments at pH = 7.
0 10 20 30 40 50 600.0
2.0x10-4
4.0x10-4
6.0x10-4
8.0x10-4 2 M 0.025 g L-1 7
2 M 0.5 g L-1 7
0.1 M 0.5 g L-1 7
2 M 0.5 g L-1 7 buffered
2 M 0.5 g L-1 3
CC
H2O
(m
ol L
-1)
TIME (min)
CH3OH TiO2 pH
Figure 3.1. Time‐evolution of formaldehyde formation during methanol photocatalytic
oxidation experiments. Conditions: T = 25 °C, Qg = 30 L h‐1 (O2).
CAPÍTULO 3 (CHAPTER 3)
118
0 1 2 3 40.0
3.0x10-6
6.0x10-6
9.0x10-6
1.2x10-5
1.5x10-5
r CH
2O (
mol
L-1
min
-1)
CCH3OH (mol L-1)
Figure 3.2. Rate of formaldehyde formation vs. methanol concentration in photocatalytic
oxidation experiments. Conditions: pH = 7, T = 25 °C, CTiO2 = 2 g L‐1, Qg = 30 L h‐1 (O2 or
O3/O2).
Figure 3.3. (A): Rate of formaldehyde formation vs. TiO2 loading in methanol
photocatalytic oxidation experiments. (B): Fitting results of linearized Eq. (3.19).
Conditions: pH = 7 and 3, T = 25 °C, Qg = 30 L h‐1 (O2 or O3/O2).
0.0 0.5 1.0 1.5 2.0 2.5 3.00.0
3.0x10-6
6.0x10-6
9.0x10-6
1.2x10-5
pH=7 pH=3
r CH
2O
(m
ol L
-1 m
in-1
)
CTiO2 (g L-1)
0 10 20 30 402
4
6
8
10
1/
(ei
nste
in m
ol-1)
1/CTiO2 (L g-1)
(A)
(B)
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Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
119
Table 3.1. Rate and photonic efficiency of CH2O formation and absorbed and fraction
light intensity.
CTiO2
(g L‐1) pH0
r CH2O /rCO2
(M min‐1)
ξ
(mol einstein‐1)
Ia
(einstein L‐1 min‐1) Fs
0.025 7 6.05x10‐6 0.23 1.27x10‐5 0.47
0.05 7 8.29x10‐6 0.31 1.73x10‐5 0.65
0.1 7 1.07x10‐5 0.40 2.24x10‐5 0.84
0.3 7 1.16x10‐5 0.44 2.43x10‐5 0.91
0.5 7 1.22x10‐5 0.46 2.55x10‐5 0.96
1 7 1.23x10‐5 0.46 2.57x10‐5 0.97
2 7 1.34x10‐5 0.50 2.80x10‐5 ~1.0
3 7 1.17x10‐5 0.44 2.45x10‐5 0.92
0.5 7a 7.51x10‐6 0.28 2.55x10‐5 0.96
0.025 3 2.65x10‐6 0.10 7.81x10‐6 0.29
0.05 3 4.01x10‐6 0.15 1.18x10‐5 0.44
0.1 3 4.62x10‐6 0.20 1.56x10‐5 0.59
0.5 3 8.77x10‐6 0.33 2.60x10‐5 0.97
1 3 8.82x10‐6 0.33 2.61x10‐5 0.98
0.5 3b 6.76x10‐6 0.26 2.60x10‐5 0.97
aBuffered solution (H3PO4 30 mM), bExperiment with CH2O2 (10 mM)
On the other hand, it is known that the size of the TiO2 aggregates in aqueous
solution depends on the pH value and therefore, could affect the intensity of
light absorbed by the photocatalyst [2]. Thus, additional photocatalytic
oxidation experiments with 0.025 ‐ 1 g L‐1 TiO2 were carried out at pH0 = 3 to
calculate both, the quantum yield (CH2O‐pH3) and the absorbed photon flux at this
pH value. Firstly, the pH value did not significantly change during the reaction
time. It was observed that the rate of formaldehyde formation at pH0 = 3 was
quite lower than at pH0 = 7 (non‐buffered) at the same catalyst loading (see Fig.
3.1). These results have also been plotted in Fig. 3.3(A) whereas Fig. 3.3(B) shows
fitting results of linearized Eq. (3.19) for experiments at pH0 = 3. The quantum
CAPÍTULO 3 (CHAPTER 3)
120
yield calculated was CH2O‐pH3 = 0.34 mol einstein‐1. The absorbed light intensity
calculated was somewhat higher than at pH = 7 in experiments at the same
catalyst loading (Table 3.1), according to the lower aggregates size expected at
lower pH. The lower reaction rates and quantum yield observed could be
attributed to the stabilization of the 2 2O CH OH radical at acidic pH [42] and/or
to the presence of ClO4‐ ions (HClO4 used to set pH = 3) that can be absorbed
onto the catalyst surface [43]. Previously, Du and Rabani [42] showed that
quantum yields of formaldehyde formation during photocatalytic oxidation of
methanol with a TiO2 catalyst fell down from 0.2 to 0.05 mol einstein‐1 at
pH = 7 and pH = 3, respectively, whereas when comparing the quantum yield of
CO2 formation during photocatalytic oxidation of formic acid at pH = 3, the
value was quite similar to that found for the methanol system at pH = 7. This
was explained on the basis of the stabilization of the 2 2O CH OH radical at acidic
pH when using methanol. However, our experiments carried out at pH = 3 with
formic acid 0.01 M (high enough to reach a constant reaction rate independent of
formic acid concentration [42]) led to the results summarized in Table 3.1, where
the photonic efficiency calculated for CO2 formation was 0.26 mol einstein‐1,
quite similar (even lower) than the value obtained in the methanol system.
These results seem to be pointed out that the stabilization of the 2 2O CH OH
radical at pH = 3 is not the main responsible of the decrease observed in the
quantum yield at this pH value under the operating conditions used here and,
therefore, we have used also methanol at pH = 3 as model compound.
Finally, under the same experimental conditions, Ia is not expected to
significantly change in presence of ozone during photocatalytic ozonation
experiments since O3 does not absorb black light radiation (= 365 nm). The
calculated Ia values were used to determine the quantum yields for
photocatalytic ozonation processes for comparative purposes.
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Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
121
3.3.2. Photocatalytic oxidation versus photocatalytic ozonation
To compare photocatalytic oxidation behavior versus photocatalytic
ozonation, similar experiments were carried out at the same operating
conditions of methanol concentration, gas flow rate, radiation and pH, but
feeding ozone to the system. At 2 M methanol concentration used here, also in
ozonation and photocatalytic ozonation experiments, formaldehyde ozonation is
not expected. According to the kinetic constants of methanol and formaldehyde
with O3 (0.024 and 0.10 M‐1 s‐1 [26], respectively) and the concentration of both
compounds, the formaldehyde formation rate with 2 M methanol was almost
500 times higher than the formaldehyde ozonation (using a generic
concentration about 10‐3 M similar to CH2O concentrations observed in this
work) proving that formaldehyde ozonation could be neglected in the overall
reaction rate.
The comparison of formaldehyde formation during photocatalytic oxidation
(TiO2/O2/UVA) and photocatalytic ozonation (TiO2/O3/UVA) of methanol is
presented in Figure 3.4(A). Also, for comparative purposes single ozonation (O3)
and photolytic ozonation results (i.e. irradiated O3 without TiO2, O3/UVA) have
been plotted. These experiments were carried out at pH = 3 to minimize indirect
ozone reactions. This pH value did not significantly change throughout the
reaction time. As expected, ozonation and photolytic ozonation gave place to
similar formaldehyde evolution since ozone does not absorb black light
radiation. On the other hand, the results observed for photocatalytic oxidation
are also quite similar to the ozone process. The highest rate of formaldehyde
formation was found during photocatalytic ozonation. Reaction rates were
calculated through the slope of the CH2O concentration‐time evolution and are
displayed in Table 3.2. A synergistic effect between ozone and the irradiated
TiO2 can be observed during TiO2/O3/UVA treatment where the reaction rate is
almost twice higher than the sum of the reaction rates of individual process
(2.91x10‐5 M min‐1 versus 1.89x10‐5 M min‐1, respectively).
CAPÍTULO 3 (CHAPTER 3)
122
Figure 3.4. Time‐evolution of formaldehyde (A) and hydrogen peroxide (B) concentration
during ozonation and photocatalytic oxidation/ozonation experiments. Conditions: pH =
3, T = 25 °C, CTiO2 = 0.5 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 30 L h‐1 (O2 or O3/O2).
The accumulation of important concentrations of hydrogen peroxide was
also observed during ozonation experiments. Results are plotted in Figure
3.4(B). As can be seen, during photocatalytic oxidation of methanol small
amounts of hydrogen peroxide could be detected whereas during single
ozonation and photolytic ozonation, the evolution of hydrogen peroxide is
comparable to the formaldehyde formation. This is explained on the basis of
ozone‐methanol direct reaction which gives place to hydrogen peroxide [44]
(B)
(A)
0 10 20 30 40 50 600.0
1.0x10-4
2.0x10-4
3.0x10-4
4.0x10-4
5.0x10-4
TIME (min)
CH
2O
2 (
mol
L-1
)
Photocatalytic oxidation Ozonation Photolytic ozonation Photocatalytic ozonation
0.0
4.0x10-4
8.0x10-4
1.2x10-3
1.6x10-3
CC
H2O
(mo
l L-1
)
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
123
according to:
3 3 2 2 2O CH OH CH O H O (3.20)
It is noticeable that hydrogen peroxide concentration was much lower in the
photocatalytic ozonation experiment likely due to the consumption of H2O2
through reaction (3.6) (H2O2‐eaq‐ reaction: k = 1.1x1010 M‐1 s‐1 [20]). This reaction
would also enhance the generation of hydroxyl radicals improving methanol
oxidation.
In addition, dissolved ozone was measured during all the ozonation
experiments. The concentration was nearly constant at 7.5x10‐6 M and 4.5x10‐6 M
during ozonation and photocatalytic ozonation, respectively. Also the ozone
concentration in the gas phase in the reactor outlet was lower during
photocatalytic ozonation (5.0 mg L‐1) than during ozonation (6.5 mg L‐1). The
lower values found during photocatalytic ozonation suggests that O3 is also
being consumed through reaction (3.15) (O3‐eaq‐ reaction: k = 3.6x1010 M‐1 s‐1,
which is also higher than the one of O2‐eaq‐ reaction: k = 1.9x1010 M‐1 s‐1 [20]),
improving methanol oxidation rate due to the generation of additional hydroxyl
radicals.
Therefore, taking into account the slow rate of direct methanol‐ozone
reaction and the negligible contribution of ozone decomposition at the pH value
used here, the synergism observed may be related to the reaction of dissolved
ozone and hydrogen peroxide as electron acceptors to produce HO radicals,
thus avoiding to some extent the recombination reactions on the TiO2 catalyst.
CAPÍTULO 3 (CHAPTER 3)
124
Tabl
e 3.
2. E
xper
imen
tal c
ondi
tions
, rea
ctio
n ra
te c
ontr
ibut
ions
and
qua
ntum
yie
ld o
f the
pho
toca
taly
tic re
actio
ns.
Proc
ess
Irra
diat
ion
CTi
O2
(g L
-1)
CO
3,g
inle
t (m
g L-1
) pH
rC
H2O
(M m
in-1
) rO
3b
(M m
in-1
) rH
O-O
3c
(M m
in-1
) rh
νd
(M m
in-1
) Eh
v
(%)
h ν
(mol
ein
stei
n-1)
Phot
ocat
alyt
ic
oxid
atio
n (T
iO2/O
2/UV
A)
on
0.5
0 3
8.77
x10-6
---
---
8.
77x1
0-6
0 0.
34
on
0.5
0 7
1.22
x10-5
---
---
1.
22x1
0-5
0 0.
48
on
0.5
0 7a
7.51
x10-6
---
---
7.
51x1
0-6
0 0.
29
Ozo
natio
n (O
3)
off
0 10
3
1.02
x10-5
1.
02x1
0-5
0 ---
---
---
off
0 20
3
1.55
x10-5
1.
55x1
0-5
0 ---
---
---
off
0 30
3
1.97
x10-5
1.
97x1
0-5
0 ---
---
---
off
0 30
7a
1.55
x10-4
1.
97x1
0-5
1.35
x10-4
---
---
---
Phot
ocat
alyt
ic
ozon
atio
n (T
iO2/O
3/UV
A)
on
0.5
10
3 2.
91x1
0-5
1.02
x10-5
0
1.89
x10-5
53
0.
73
on
0.5
20
3 3.
63x1
0-5
1.55
x10-5
0
2.08
x10-5
57
0.
80
on
0.5
30
3 3.
91x1
0-5
1.97
x10-5
0
1.94
x10-5
55
0.
75
on
0.5
30
7a 2.
38x1
0-4
1.97
x10-5
1.
35x1
0-4
8.36
x10-5
91
3.
27
a Buf
fere
d so
lutio
n (H
3PO
4 30
mM
), b I
t coi
ncid
es w
ith rC
H2O
for s
ingl
e oz
onat
ion
expe
rim
ents
at p
H =
3, c C
alcu
late
d fr
om o
zona
tion
expe
rim
ents
by
mea
ns o
f the
diff
eren
ce b
etw
een
rCH
2O-r
O3,
d Cal
cula
ted
by m
eans
of t
he d
iffer
ence
bet
wee
n rC
H2O
- rO
3-rH
O-O
3.
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
125
3.3.2.1. Influence of ozone concentration
The influence of ozone concentration in the feeding gas during photocatalytic
ozonation of methanol was studied at pH = 3. Results of formaldehyde and
hydrogen peroxide formation with time are depicted in Figure 3.5(A) and Figure
3.5(B), respectively. Also, for the sake of comparison, single ozonation results
varying the ozone gas concentration have been plotted. It can be observed that
ozone concentration exerts a positive effect on methanol oxidation rate
(formaldehyde formation), compared to the ozone free experiments (see also
Fig. 3.4(A)) both in single ozonation and photocatalytic ozonation, although the
effect is less important when increasing the ozone gas concentration. Reaction
rates of formaldehyde formation have been calculated and are shown in Table
3.2. The increase of formaldehyde formation rate is proportional to the
increasing ozone concentration in single ozonation experiments whereas it is
more important from 10 to 20 mg L‐1 of ozone than from 20 to 30 mg L‐1 in
photocatalytic ozonation. This behavior has been observed before during the
photocatalytic ozonation process [19] and has been attributed to complex
Langmuir kinetics for substances that absorb and react on the semiconductor
surface.
Regarding the evolution of H2O2 during the experiments at different O3
concentration (Fig. 3.5(B)), it can be observed that H2O2, formed mainly from
direct ozonation of methanol, is further accumulated during ozonation
treatment while is not in the photocatalytic ozonation process and regardless of
the ozone concentration. The latter is likely due to the consumption of H2O2 in
the photocatalytic process through reaction (3.6). Despite this, a positive effect of
ozone concentration is observed on the evolution of H2O2 in both treatments,
more pronounced during ozonation.
CAPÍTULO 3 (CHAPTER 3)
126
Figure 3.5. Time‐evolution of formaldehyde (A) and hydrogen peroxide (B) concentration
during methanol ozonation and photocatalytic ozonation experiments. Conditions: pH =
3, T = 25 °C, CTiO2 = 0.5 g L‐1, CO3,g inlet = 10, 20, 30 mg L‐1, Qg = 30 L h‐1 (O3/O2).
3.3.2.2. Influence of pH
Previous experiments have been carried out at pH = 3 to minimize ozone
decomposition reaction and, thus, avoiding indirect methanol‐ozone oxidation
reaction. However, pH plays a crucial role during ozonation processes. Results
of photocatalytic oxidation, ozonation and photocatalytic ozonation of CH3OH
at pH = 3 and pH = 7 (buffered solution) are compared in Figure 3.6. Regarding
0 10 20 30 40 50 600.0
2.0x10-4
4.0x10-4
6.0x10-4
8.0x10-4
1.0x10-3
1.2x10-3
TIME (min)
CH
2O2
(mo
l L-1
)
0.0
4.0x10-4
8.0x10-4
1.2x10-3
1.6x10-3
2.0x10-3
s-1
s-1
s-1
10 mgL-1
20 mgL-1
30 mgL-1
CC
H2O
(m
ol L
-1)
Ozonation Photocatalytic ozonation
(A)
(B)
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
127
CH2O formation, slight differences were observed in Fig. 3.6(A) for
photocatalytic oxidation reaction at pH = 3 and pH = 7 as commented in the
previous section. Nevertheless, when ozone is present, a large increase in the
rate of CH2O formation is observed at pH = 7 respect to pH = 3. During single
ozonation experiments the formaldehyde formation rate increased around 8
times (from 1.97x10‐5 to 1.55x10‐4 M min‐1, see Table 3.2). Direct ozone‐methanol
reaction is not affected by pH modifications since CH3OH does not dissociate at
the studied pH range [26]; as a consequence, this behavior is related to ozone
decomposition reactions (3.9) ‐ (3.14) that are produced at neutral pH to
generate HO radicals in the reaction medium. In this line, lower dissolved
ozone concentration has been detected at pH = 7 (8x10‐7 M and 1x10‐6 M in
ozonation and photocatalytic ozonation, respectively) compared to pH = 3
(8.2x10‐6 M and 6.1x10‐6 M in ozonation and photocatalytic ozonation,
respectively). Also, ozone concentration in the gas phase at the reactor outlet has
been observed to decrease with pH increase from 3 to 7 (from 15 at pH 3 to 10
mg L‐1 at pH 7 in ozonation, and from 11 at pH 3 to 6 mg L‐1 at pH 7 in
photocatalytic ozonation runs). This corroborates a higher consumption of
ozone at neutral pH likely due to its faster decomposition.
In addition, hydrogen peroxide formed upon direct ozone‐methanol reaction
can be dissociated (reaction (3.21), pKa = 11.3), promoting reaction (3.10) to some
extent:
pKa 11.3
2 2 2H O HO H (3.21)
In fact, hydrogen peroxide concentrations detected at pH = 7 were fairly
lower than at pH = 3 (see Fig. 3.6(B)), indicating that H2O2 is consumed during
the single ozonation process. Another fact that can affect the ozonation
mechanism due to the pH value is that the hydroperoxide radical formed upon
methanol ‐ HO oxidation ( •
2HO ) will be on its dissociated form ( •‐2O ) at neutral
pH (reaction (3.11), pKa = 4.8), then it also participates in the reaction
mechanism of ozone decomposition.
CAPÍTULO 3 (CHAPTER 3)
128
0 10 20 30 40 50 600.0
2.0x10-4
4.0x10-4
6.0x10-4
8.0x10-4
1.0x10-3
1.2x10-3
TIME (min)
CH
2O2
(mo
l L-1
)
0.0
3.0x10-3
6.0x10-3
9.0x10-3
1.2x10-2
1.5x10-2
Photocatalytic oxidation Ozonation Photocatalytic ozonation
CC
H2O
(m
ol L
-1)
pH=7 pH=3
Figure 3.6. Time‐evolution of formaldehyde (A) and hydrogen peroxide (B) concentration
during methanol ozonation and photocatalytic oxidation/ozonation experiments.
Conditions: pH = 3, 7, T = 25 °C, CTiO2 = 0.5 g L‐1, CO3,g inlet = 30 mg L‐1, Qg = 30 L h‐1 (O2 or
O3/O2).
The increase of formaldehyde formation rate observed was more significant
in photocatalytic ozonation due to the formation of higher concentration of
oxidizing species at the photocatalyst surface according to the reaction scheme
described above. At pH = 7 and 30 mg L‐1 of ozone inlet concentration, a
synergistic effect is clearly observed in the reaction rate (see Table 3.2), being the
sum of ozone and photocatalytic oxidation rates 1.62x10‐4 M min‐1 fairly lower
than that of photocatalytic ozonation (2.38x10‐4 M min‐1). Therefore, an
(A)
(B)
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
129
increment in the pH value exerts a positive effect on the photocatalytic
ozonation reaction due to the formation of species that favor the ozone
decomposition in water.
3.3.2.3. Synergistic effect and quantum yield of photocatalytic induced reactions
From the previous experimental results, different contributions to the rate of
formaldehyde formation have been calculated. On one hand, when single
ozonation experiments are carried out at pH = 3, the contribution of O3 indirect
reactions (rHO‐O3) can be neglected and therefore, the observed reaction rate at
this pH would only be due to the direct ozone‐methanol reaction (rO3). This was
extrapolated to photocatalytic ozonation considering the same contribution of
ozone‐methanol direct reaction than in the absence of light. Additionally, for
photocatalytic treatments, another contribution needs to be considered due to
the photogenerated oxidizing species (i.e. h+ and HO from reactions (3.1) ‐ (3.6),
and (3.15)) apart from the direct or indirect ozonation mechanism. This
contribution, named rhv, is the observed rate of CH2O formation in
photocatalytic oxidation, and was calculated as the difference between the
observed reaction rate rCH2O and the direct‐indirect ozone reactions in
photocatalytic ozonation. On the other hand, when reactions were carried out at
pH = 7, the contribution of indirect ozone reactions become very important. In
this case, this contribution was calculated by subtracting the ozone‐direct
reaction rate (calculated at pH = 3 at the same operating conditions) to the
observed reaction rate of formaldehyde formation during single ozonation
experiments. Also, these results were extrapolated to photocatalytic ozonation at
pH = 7. The different contributions to the global reaction rate of formaldehyde
formation during methanol oxidation have been summarized in Table 3.2.
Regarding the photocatalytic contribution, it is noticeable the increase
underwent in the light induced reaction rate when ozone was present compared
to the one of photocatalytic oxidation at the same operating conditions. The
degree of enhancement produced as a consequence of O3 was calculated as
follows:
CAPÍTULO 3 (CHAPTER 3)
130
h 2 3 h 2 2
h 2 2
r TiO / O / UVA r TiO / O / UVAE ∙100 %
r TiO / O / UVA
(3.22)
Table 3.2 shows the values obtained for all the photocatalytic ozonation
experiments. The enhancement observed in the photocatalytic contribution
increased from 53 % to 57 % when increasing ozone concentration in the feeding
effluent at pH = 3, although slight differences were observed when using 20 or
30 mg L‐1 O3 as commented in the previous section. The highest enhancement
was observed at pH = 7 with 30 mg L‐1 of ozone. This parameter can be
considered as a real evidence of the synergism produced between ozone and
irradiated TiO2.
At the experimental conditions used (high excess of CH3OH), the reaction
rate calculated for the photocatalytic contribution (rhv) can be well considered as
the reaction rate of oxidizing species formation due to light induced reactions.
Therefore, the quantum yield of these reactions (light induced ones) can be
calculated taking into account the absorbed light intensity at the operating
conditions used (see Table 3.1). The calculated quantum yields (hv) for
photocatalytic oxidation and photocatalytic ozonation processes are presented in
Table 3.2. The quantum yield of photogenerated species formation in
photocatalytic ozonation at pH = 3 was found to be increased twice and near
triple times the value observed for photocatalytic oxidation as ozone
concentration increased. Quantum yield reached a plateau at high ozone
concentrations, as can also be observed in Figure 3.7 where the evolution of the
quantum yield with ozone concentration has been depicted. This trend suggests
that a limiting reaction rate has been reached in the system due to the reaction of
hydroxyl radicals with O3. Sun and Bolton [21] also observed this behavior while
studying the photocatalytic oxidation of methanol with the addition of
hydrogen peroxide as an electron acceptor. On the other hand, the highest
improvement of the quantum yield has been observed in the experiment of
photocatalytic ozonation at pH = 7 reaching a value near 3 photogenerated
oxidizing species per absorbed photon.
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
131
0 5 10 15 20 25 300.0
0.2
0.4
0.6
0.8
1.0
hv
(m
ol e
inst
ein
-1)
CO3,g inlet (mg L-1)
Figure 3.7. Evolution of quantum yield of oxidizing species formation through light
induced reactions with ozone inlet concentration. Conditions: pH = 3, T = 25 °C, CTiO2 =
0.5 g L‐1, Qg = 30 L h‐1 (O2 or O3/O2).
The maximum quantum yield of formaldehyde formation expected during
photocatalytic oxidation is CH2O,max = hv,max = 1.5 mol einstein‐1, according to the
reaction mechanism proposed where 2 photons are needed to produce 3
oxidizing species (i.e. holes and/or hydroxyl radicals) taking into account also
the radical •‐2O / •
2HO formed from CH3OH oxidation [6]. This maximum
becomes even greater (hv;max‐O3 = 2 ‐ 3) when ozone is present depending on pH,
due to reactions (3.13) ‐ (3.15) where ozone needs 1 e‐ per HO formed, and the
generation of H2O2 from direct methanol ozonation (reaction (3.20)), that acts as
an electron acceptor (reaction (3.6)) needs also 1 e‐ per HO formed. The
maximum quantum yield at basic‐neutral pH when O3 is present has not been
accurately calculated since ozone‐indirect reactions are closely related to the
reactions involving photogenerated species. Thus, the calculated values seem to
be reasonable although they should be taken with caution, especially that
calculated at pH = 7 where the contribution of the indirect‐ozone reaction has
been taken as the same than in single ozonation and may be underestimated.
CAPÍTULO 3 (CHAPTER 3)
132
Further work will be carried out to quantitatively calculate the different
contributions through a detailed reaction mechanism for the AOPs studied
(photocatalytic oxidation, ozonation and photocatalytic ozonation). Despite this
it is clear that ozone exerts a positive effect on the photocatalytic induced
reactions (not only O3 direct and indirect reactions) due to two main factors: (i)
ozone acts as an electron acceptor, and (ii) hydrogen peroxide usually formed
during direct ozonation reactions also acts as an electron acceptor, both leading
to higher amount of hydroxyl radicals per photon absorbed and also reducing
the recombination process to some extent.
3.3.2.4. Simplified economic considerations
In addition to the benefits observed with the combined treatment, i.e.
photocatalytic ozonation (TiO2/O3/UVA), in the reaction rate of formaldehyde
formation, the synergistic effect between ozone and irradiated TiO2 and the
highest quantum yield of photo‐generated oxidizing species production, one
important issue in the application of combined treatments is the economy of the
process. Accordingly, to compare the different systems, costs associated with the
normal operation have been evaluated by taking into account the oxygen and
electrical energy consumed to generate the oxidants and/or radiation and the
amount of formaldehyde formed. It has to be highlighted that this simplified
economic study aims only the comparison of the systems (i.e. investments, loss
of TiO2 activity and separation, and other factors are not considered) not
providing an actual costs estimation.
According to Rivas et al. [45] for the same ozone generator, the dependence
between O3 production and current consumption can be estimated as:
‐13O % of max. flow rate in gL 186 ∙current consumption in A (3.23)
with maximum mass flow rate from oxygen 12 g h‐1. From the experimental
conditions for 10 and 30 mg L‐1 O3 at 30 L h‐1 those are 2.5 % and 7.5 % of the
maximum amount produced, respectively. In photocatalytic experiments it is
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
133
also necessary to take into account the lamps energy consumption (2 lamps with
an input power of 15 W). The cost associated with electricity has been
considered 0.14 € (kWh)‐1 according to the local supplier. In addition, in all the
experiments oxygen from cylinders was used which has a cost of 0.262 € h‐1 [45]
at a flow rate of 30 L h‐1.
With the above data, Table 3.3 shows the results obtained to produce 0.001
mol of formaldehyde from methanol oxidation taking into account the time
needed. It can be observed that the main contribution (of those considered) is
the oxygen consumption. It has to be pointed out that this contribution would be
far different if air from atmosphere was used both in photocatalytic oxidation
and ozonation runs (in the latter two using an air compressor) since there are
ozone generators able to work with air. The operation costs (as calculated here)
are always higher in photocatalytic experiments or ozonation alone than in the
combined process, especially when working at pH = 7, where the reaction time is
highly reduced. Therefore, photocatalytic ozonation treatment is not only
attractive from the point of view of their performance in terms of reaction rates
but also in terms of economic considerations, where the synergism between
ozone and TiO2 photocatalysis is also clear. However, it has to be emphasized
that this simplified economic study could give different results in every
particular case taking into account the necessities of effluent mineralization,
discharge limits required, investment and replacement costs, etc. as reported
before [45] where ozone treatments involved higher costs at the same
mineralization degree obtained.
CAPÍTULO 3 (CHAPTER 3)
134
Tabl
e 3.
3. E
stim
ated
cos
ts fo
r the
tran
sfor
mat
ion
of 1
0-3 m
ol C
H3O
H (C
H2O
form
ed).
Proc
ess
pH
CO
3g in
let
(mg
L-1)
Tim
e us
ed
(h)
EO3
(kW
h)
E UV
A
(kW
h)
Elec
tric
ity
(€)
O2
(€)
Cos
t/mm
ol
conv
erte
d (€
mm
ol-1
) Ph
otoc
atal
ytic
ox
idat
ion
(TiO
2/O2 /U
VA
)
3 0
1.90
0
5.70
x10-2
7.
98x1
0-3
4.97
x10-1
5.
05x1
0-1
7 0
2.22
0
6.66
x10-2
9.
32x1
0-3
5.81
x10-1
5.
90x1
0-1
Ozo
natio
n (O
3 )
3 10
1.
63
4.83
x10-3
0
6.76
x10-4
4.
27x1
0-1
4.28
x10-1
7 30
0.
11
9.54
x10-4
0
1.34
x10-4
2.
80x1
0-2
2.80
x10-2
Phot
ocat
alyt
ic
ozon
atio
n (T
iO2/O
3 /UV
A)
3 10
0.
57
1.69
x10-3
1.
72x1
0-2
2.64
x10-3
1.
50x1
0-1
1.52
x10-1
7 30
0.
07
6.21
x10-4
2.
10x1
0-3
3.81
x10-4
1.
80x1
0-2
1.90
x10-2
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
135
3.4. CONCLUSIONS
Major conclusions reached in this study are:
Absorbed light intensity during photocatalytic oxidation of methanol was
indirectly calculated at different pH values and was always higher than 90 %
of that impinging the reactor for the catalyst loading used (0.5 g L‐1 of TiO2
P25).
The presence of ozone during photocatalytic ozonation at pH = 3 exerts a
positive effect on the reaction rate of formaldehyde formation respect to
photocatalytic oxidation. This was not only due to the direct ozone‐methanol
reaction, but also due to the reaction of dissolved ozone and hydrogen
peroxide (formed upon CH3OH ozonation), electron acceptors to produce
higher concentration of hydroxyl radicals.
An increase in ozone concentration exerts a positive but low effect during
photocatalytic ozonation likely due to self‐scavenging reactions between
HO and ozone or hydrogen peroxide.
The quantum yield calculated for photo‐generated oxidizing species during
photocatalytic oxidation at pH = 3 was hv = 0.34 mol einstein‐1. This
parameter was increased to around hv = 0.80 mol einstein‐1 when combining
TiO2/O3/UVA. This enhancement was even higher at pH = 7 from hv = 0.29 to
hv = 3.27 mol einstein‐1. Therefore, a clear synergism between ozone and
black light photocatalytic oxidation with TiO2 is pointed out due to the
decrease of the recombination process. The reaction of dissolved ozone and
hydrogen peroxide with photo‐generated electrons is likely the reason of the
reaction rate enhancement.
Simplified economic evaluation taking into account only normal operation
has shown that photocatalytic ozonation could be a cost‐effective treatment
depending on the actual necessities of the effluent to be treated.
CAPÍTULO 3 (CHAPTER 3)
136
AKNOWLEDGEMENTS
This work has been supported by the Spanish Ministerio de Ciencia e
Innovación (MICINN) and European Feder Funds through the Project CTQ2009‐
13459‐C05‐05/PPQ. A. Rey thanks the University of Extremadura for a post‐
doctoral research contract.
REFERENCES
[1] Kabra, K.; Chaudhary, R.; Sawhney, R.L. “Treatment of hazardous organic
and inorganic compounds through aqueous‐phase photocatalysis: A review”.
Ind. Eng. Chem. Res. 43 (2004) 7683‐7696.
[2] Malato, S.; Fernández‐Ibáñez, P.; Maldonado, M.I.; Blanco, J.; Gernjak, W.
“Decontamination and disinfection of water by solar photocatalysis: Recent
overview and trends”. Catal. Today 147 (2009) 1‐59.
[3] Chong, M.N.; Jin, B.; Chow, C.W.K.; Saint, C. “Recent developments in
photocatalytic water treatment technology: A review”. Water Res. 44 (2010)
2997‐3027.
[4] Adán, C.; Bahamonde, A.; Fernández‐García, M.; Martínez‐Arias, A.
“Structure and activity of nanosized iron‐doped anatase TiO2 catalysts for
phenol photocatalytic degradation”. Appl. Catal. B Environ. 72 (2007) 11‐17.
[5] Agustina, T.E.; Ang, H.M.; Vareek, V.K. “A review of synergistic effect of
photocatalysis and ozonation on wastewater treatment”. J. Photochem.
Photobiol. C Photochem. Rev. 6 (2005) 264‐273.
[6] Augugliaro, V.; Litter, M.; Palmisano, L.; Soria, J. “The combination of
heterogeneous photocatalysis with chemical and physical operations: a tool for
improving the photoprocess performance”. J. Photochem. Photobiol. C
Photochem. Rev. 7 (2006) 127‐144.
[7] Salvador, P. “On the nature of photogenerated radical species active in the
oxidative degradation of dissolved pollutants with TiO2 aqueous suspensions: a
revision in the light of the electronic structure of adsorbed water”. J. Phys.
Chem. C 111 (2007) 17038‐17043.
[8] Henderson, M.A. “A surface perspective on TiO2 photocatalysis”. Surface Sci.
Rep. 66 (2011) 185‐297.
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
137
[9] Goldstein, S.; Behar, D.; Rabani, J. “Mechanism of visible light photocatalytic
oxidation of methanol in aerated aqueous suspensions of carbon‐doped TiO2”. J.
Phys. Chem. C. 112 (2008) 15134‐15139.
[10] Goldstein, S.; Behar, D.; Rabani, J. “Nature of the oxidizing species formed
upon UV photolysis of C‐TiO2 aqueous suspensions”. J. Phys. Chem. C
113 (2009) 12489‐12494.
[11] Beltrán, F.J. “Ozone Reaction Kinetics for Water and Wastewater Systems”.
CRC Press, Boca Raton, Florida, USA, 2004.
[12] Mvula, M.; von Sonntag, C. “Ozonolysis of phenols in aqueous solution”.
Org. Biomol. Chem. 1 (2003) 1749‐1756.
[13] Leitzke, A.; von Sonntag, C. “Ozonolysis of unsaturated acids in aqueous
solution: acrylic, methacrylic, maleic, fumaric and muconic acids”. Ozone Sci.
Eng. 31 (2009) 301‐308.
[14] Rakowski, S.; Cherneva, D. “Kinetics and mechanism of the reaction of
ozone with aliphatic alcohols”. Int. J. Chem. Kinetics 22 (1990) 321‐329.
[15] Rajeswari, R.; Kanmani, S. “Degradation of pesticide by photocatalytic
ozonation process and study of synergistic effect by comparison with
photocatalysis and UV/ozonation processes”. J. Adv. Oxid. Technol. 12 (2009)
208‐214.
[16] García‐Araya, J.F.; Beltrán, F.J.; Aguinaco, A. “Diclofenac removal from
water by ozone and photolytic TiO2 catalysed processes”. J. Chem. Technol.
Biotechnol. 85 (2010) 798‐804.
[17] Zou, L.D.; Zhu, B. “The synergistic effect of ozonation and photocatalysis on
color removal from reused water”. J. Photochem. Photobiol. A Chem. 196 (2008)
24‐32.
[18] Beltrán, F.J.; Rivas, F.J.; Gimeno, O.; Carbajo, M. “Photocatalytic enhanced
oxidation of fluorene in water with ozone. Comparison with other chemical
oxidation methods”. Ind. Eng. Chem. Res. 44 (2005) 3419‐3425.
[19] Beltrán, F.J.; Aguinaco, A.; García‐Araya, J.F. “Mechanism and kinetics of
sulfamethoxazole photocatalytic ozonation in water”. Water Res. 43 (2009) 1359‐
1369.
[20] Buxton, G.V.; Greenstock, C.L.; Helman, W.P.; Ross, A.B. “Critical review of
rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl
radicals (∙OH/O‐) in aqueous solution”. J. Phys. Chem. Ref. Data 17 (1988) 513‐
886.
CAPÍTULO 3 (CHAPTER 3)
138
[21] Sun, L.; Bolton, J.R. “Determination of the quantum yield for the
photochemical generation of hydroxyl radicals in TiO2 suspensions”. J. Phys.
Chem. 100 (1996) 4127‐4134.
[22] Tachikawa, T.; Tojo, S.; Kawai, K.; Endo, M.; Fujitsuka, M.; Ohno, T.;
Nishijima, K.; Miyamoto, Z.; Majima, T. “Photocatalytic oxidation reactivity of
holes in the sulfur and carbon‐doped TiO2 powders studied by time‐resolved
diffuse reflectance spectroscopy”. J. Phys. Chem B 108 (2004) 19299‐19306.
[23] Wang, C.; Rabani, J.; Bahnemann, D.W.; Dohrmann, J.K. “Photonic
efficiency and quantum yield of formaldehyde formation from methanol in the
presence of various TiO2 photocatalysts”. J. Photochem. Photobiol. A Chem. 148
(2002) 169‐176.
[24] Gao, R.; Stark, J.; Bahnemann, D.W.; Rabani, J. “Quantum yields of hydroxyl
radicals in illuminated TiO2 nanocrystallite layers”. J. Photochem. Photobiol. A
Chem. 148 (2002) 387‐391.
[25] Marugán, J.; Hufschmidt, D.; López‐Muñoz, M.J.; Selzer, V.; Bahnemann, D.
“Photonic efficiency for methanol photooxidation and hydroxyl radical
generation on silica‐supported TiO2 photocatalysts”. Appl. Catal. B Environ. 62
(2006) 201‐207.
[26] Hoigné, J.; Bader, H. “Rate constant of reactions of ozone with organic and
inorganic compounds in water I. Non‐dissociating organic compounds”. Water
Res. 17 (1983) 173‐183.
[27] Nash, T. “The colorimetric estimation of formaldehyde by means of the
Hantzsch reaction”. Biochemistry 55 (1953) 416‐421.
[28] Flyunt, R.; Leitzke, A.; Mark, G.; Mvula, E.; Reisz, E.; Schick, R.; von
Sonntag, C. “Determination of ∙OH, O2‐, and hydroperoxide yields in ozone
reactions in aqueous solution”. J. Phys. Chem. B 107 (2003) 7242‐7253.
[29] Masschelein, W.; Denis, M.; Ledent, R. “Spectrophotometric determination
of residual hydrogen peroxide”. Water Manag. Water Sewage Works (1977) 69‐
72.
[30] Bader, H.; Hoigné, J. “Determination of ozone in water by the indigo
method”. Water Res. 15 (1981) 449‐456.
[31] Hatchard, C.G.; Parker, C.A. “A new sensitive chemical actinometer. 2.
Potassium ferrioxalate as a standard chemical actinometer”. Proc. Roy. Soc.
Lond. Ser. Math. Phys. Sci. 235 (1956) 518‐536.
PAPER 1: On ozone‐photocatalysis synergism in black‐light induced reactions:
Oxidizing species production in photocatalytic ozonation versus heterogeneous photocatalysis
139
[32] Sandell, E.B. “Colorimetric Determination of Traces of Metals”. Interscience
Pubs., New York, 1959.
[33] Serpone, N.; Salinaro, A. “Terminology, relative photonic efficiencies and
quantum yields in heterogeneous photocatalysis. Part I: suggested protocol”.
Pure Appl. Chem. 71 (1999) 303‐320.
[34] Alfano, O.M.; Cabrera, M.I.; Cassano, A.E. “Photocatalytic reactions
involving hydroxyl radical attack‐I. Reaction kinetics formulation with explicit
photon absorption effects”. J. Catal. 172 (1997) 370‐379.
[35] Cabrera, M.I.; Negro, A.C.; Alfano, O.M.; Cassano, A.E. “Photocatalytic
reactions involving hydroxyl radical attack‐II. Kinetics of the decomposition of
trichloroethylene using titanium dioxide”. J. Catal. 172 (1997) 380‐390.
[36] Brandi, R.J.; Alfano, O.M.; Cassano, A.E. “Rigorous model and experimental
verification of the radiation field in a flat‐plate solar collector simulator
employed for photocatalytic reactions”. Chem. Eng. Sci. 54 (1999) 2817‐2827.
[37] Brucato, A.; Cassano, A.E.; Grisafi, F.; Montante, G.; Rizzuti, L.; Vella, G.
“Estimating radiant fields in flat heterogeneous photoreactors by the six‐flux
model”. AIChE J. 52 (2006) 3882‐3890.
[38] Toepfer, B.; Gora, A.; Li Puma, G. “Photocatalytic oxidation of
multicomponent solutions of herbicides: Reaction kinetics analysis with explicit
photon absorption effects”. Appl. Catal. B Environ. 68 (2006) 171‐180.
[39] Loddo, V.; Addamo, M.; Augugliaro, V.; Palmisano, L.; Schiavello, M.;
Garrone, E. “Optical properties and quantum yield determination in
photocatalytic suspensions”. AIChE J. 52 (2006) 2565‐2574.
[40] Chen, H.Y.; Zahraa, O.; Bouchy, M. “Inhibition of the adsorption and
photocatalytic degradation of an organic contaminant in an aqueous suspension
of TiO2 by inorganic ions”. J. Photochem. Photobiol. A 108 (1997) 37‐44.
[41] Connor, P.A.; McQuillan, A.J. “Phosphate adsorption onto TiO2 from
aqueous solutions: an in situ internal reflection infrared spectroscopic study”.
Langmuir 15 (1999) 2916‐2921.
[42] Du, Y.; Rabani, J. “The measure of TiO2 photocatalytic efficiency and the
comparison of different photocatalytic titania”. J. Phys. Chem. B 107 (2003)
11970‐11978.
[43] Lefevre, G. “In situ Fourier‐transform infrared spectroscopy studies of
inorganic ions adsorption on metal oxides and hydroxides”. Adv. Colloid Interf.
Sci. 107 (2004) 109‐123.
CAPÍTULO 3 (CHAPTER 3)
140
[44] Rakovski, S.; Cherneva, D. “Kinetics and mechanism of the reaction of
ozone with aliphatic alcohols”. Int. J. Chem. Kinetics 22 (1990) 321‐329.
[45] Rivas, F.J.; Encinas, A.; Acedo, B.; Beltrán, F.J. “Mineralization of bisphenol
A by advanced oxidation processes”. J. Chem. Technol. Biotechnol. 84 (2009)
589‐594.
CAPÍTULO 4 (CHAPTER 4) PAPER 2: Influence of structural properties on the activity of WO3 catalysts for visible light photocatalytic ozonation
A. Rey, E. Mena, A.M. Chávez, F.J. Beltrán, S. Malato
Chemical Engineering Science 126 (2015) 80-90
ABSTRACT. This work is focused on the use of WO3 catalysts with different structural
properties for photocatalytic ozonation using visible light as radiation source. Different
WO3 catalysts were prepared by thermodecomposition of tungstite precursor
(H2O·WO3) at 300 and 450 °C and different time of calcination. Photocatalysts were
characterized by means of TGA-DTA, XRD, N2 adsorption-desorption isotherms, pH of
the point of zero charge (pHPZC), XPS and DR-UV-Vis spectroscopy. Photocatalytic
ozonation of ibuprofen under visible light radiation with the catalyst prepared at 450 °C and 5 min of heat-treatment gave place to complete removal of ibuprofen in less than 20 min with TOC removal up to 87 % at 120 min. In addition, this catalyst gave place to a complete removal of a mixture of ten emerging contaminants in a municipal wastewater in less than 60 min with a mineralization up to 40 % at 120 min. The highest catalytic activity of this material in the reaction under study is due to the formation of monoclinic
structure of WO3 together with the presence of higher concentration of oxygen vacancies.
Keywords: Photocatalytic ozonation, tungsten trioxide, crystalline structure, ozone, visible light.
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
143
4.1. INTRODUCTION
The development of efficient water treatment technologies to remove organic
pollutants such as solvents, dyes, pesticides, phenolic compounds or
pharmaceuticals and personal care products, among others, is an important
issue worldwide. In general, these pollutants are persistent, toxic and/or
recalcitrant to conventional wastewater treatments and have shown to be a
potential risk to human health and to the environment. Therefore, they must be
removed in order to prevent their discharge and, of course, to fulfill the required
limits of an increasingly stringent legislation.
Among the alternative technologies, the combination of ozone and
heterogeneous photocatalysis with a semiconductor, i.e. photocatalytic
ozonation, has demonstrated to efficiently remove organic compounds that are
degraded little or slowly by photocatalysis or ozonation alone. The
improvement of the combined process is mainly related to the formation of
higher concentration of oxidizing species such as hydroxyl radicals compared to
the single treatments [1,2]. Usually, this treatment has been applied by using
TiO2 as photocatalyst and UV irradiation. However, from an environmental and
economic point of view, the use of solar light is a need for the practical
deployment of photocatalytic technologies. In this sense, TiO2 presents some
limitations since it only uses around 5 % of solar radiation (UV range) in spite of
being the archetypical photocatalyst due to its relatively high efficiency, low cost
and availability [3].
Different research efforts have been carried out to overcome this limitation
mainly focusing on the development of doped or modified TiO2 or the use of
different semiconductors such as CdS, WO3, SnO2 or ZnO [3,4]. Among them,
bare tungsten trioxide WO3 has been considered unsuitable for efficient
photocatalytic oxidation due to its conduction band level respect to the
reduction band level of O2, whereas it has demonstrated an efficient behavior in
the presence of O3 [5]. WO3 is a wide‐band‐gap semiconductor (2.6 ‐ 3.0 eV) that
CAPÍTULO 4 (CHAPTER 4)
144
can be excited with radiation corresponding to the visible region of the
electromagnetic spectrum [6,7]. This fact together with its relatively low cost, no
toxicity and availability make WO3 an attractive alternative to TiO2 for
photocatalytic ozonation processes under visible or solar light radiation.
The use of WO3 in the photocatalytic ozonation process has been limited to
commercial monoclinic WO3 and phenol as target compound [5,8]. Thus, this
work focuses on the study of WO3 photocatalysts with different structural
properties for the photocatalytic ozonation of a model compound using visible
light radiation. Ibuprofen (IBP), a nonsteroidal anti‐inflammatory drug, has been
selected as target contaminant which is included in the family of emerging
contaminants (ECs) frequently detected in wastewater and different aquatic
environments [9,10]. In addition, a more realistic application of photocatalytic
ozonation with WO3 catalysts has been studied in a mixture of 10 ECs spiked
municipal wastewater treatment plant (MWWTP) effluent using the entire
simulated solar radiation spectrum.
4.2. EXPERIMENTAL SECTION
4.2.1. Catalysts preparation
Photocatalysts were prepared by thermodecomposition of commercial
H2WO4 (99 %, Sigma‐Aldrich) yellow powder precursor. The sample was
calcined in air atmosphere in a muffle furnace at different temperatures (300 and
450 °C) for 1 to 30 min in order to obtain different H2WO4/WO3 crystalline
structures according to the method reported by Cao et al. [11]. The
thermodecomposition reaction proceeds through water loss of the tungstite
structure of H2WO4 (H2O∙WO3) and subsequent crystallization as cubic WO3.
Then, at higher temperature cubic WO3 can be transformed into the monoclinic
WO3 crystalline phase [12]. This process is summarized in Eq. (4.1). Table 4.1
shows the nomenclature of the catalysts and heat‐treatment conditions together
with some characterization results.
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
145
T,t
2 4 2 3 3,cubic 2tungstite
T,t3,cubic 3,monoclinic
H WO H O∙WO WO H O
WO WO
(4.1)
Table 4.1. Nomenclature, calcination conditions and some properties of the
photocatalysts.
CATALYST Calcination
conditions
Tungstite
(%)
Cubic
(%)
Monoclinic
(%)
SBET
(m2 g‐1) pHPZC
Eg
(eV)
W‐0 ‐‐‐ ~100 0 0 27.8 4.71 2.42
W300‐1 300 °C,
1 min 82.0 18.0 0 28.7 4.39 2.43
W300‐3 300 °C,
3 min 0 100 0 40.3 4.13 2.64
W300‐5 300 °C,
5 min 0 85.7 14.3 39.0 4.18 2.65
W300‐60 300 °C,
60 min 0 80.3 19.7 n.m n.m n.m
W450‐5 450 °C,
5 min 0 23.9 76.1 30.2 5.25 2.65
W450‐30 450 °C,
30 min 0 9.5 90.5 21.1 4.98 2.67
n.m.: not measured
4.2.2. Characterization
Thermal gravimetry and differential temperature analysis (TGA‐DTA) was
performed with a SETSYS Evolution‐16 apparatus (Setaram) using the following
conditions: sample loading 20 mg, air flow 50 cm3 min‐1 at a heating rate of
5 °C min‐1 from room temperature to 600 °C.
The crystalline phases present in the photocatalysts were inferred from their
X‐ray diffraction (XRD) patterns recorded using a powder Bruker D8 Advance
XRD diffractometer with a Cu Kα radiation (λ = 0.1541 nm). The data were
collected from 2θ = 10° to 70° at a scan rate of 0.02 s−1 and 1 s per point.
Crystalline composition semi‐quantitative analysis of the samples was
performed on multi‐phase patterns by the Reference Intensity Ratio (RIR)
CAPÍTULO 4 (CHAPTER 4)
146
method using reference diffraction patterns by means of EVA v.14 software
(Bruker‐AXS). Due to the overlapping of the main characteristic peaks of the
cubic and monoclinic WO3 patterns, the analyses for the catalysts with
monoclinic contribution were performed using a calculated value of RIR for the
2θ = 28.6° diffraction peak corresponding to the (112) reflection plane.
BET surface area and pore volume of the photocatalysts were determined
from their nitrogen adsorption‐desorption isotherms obtained at ‐196 °C using
an Autosorb 1 apparatus (Quantachrome). Prior to analysis the samples were
outgassed at 150 °C for 24 h under high vacuum (< 10−4 Pa).
The point of zero charge (pHPZC) of each sample was estimated using the
mass titration method proposed by Subramanian et al. [13]. Suspensions of the
solid in deionized water at 5 % (w/w) were prepared and the pH measured after
24 h of stirring.
X‐ray photoelectron spectra (XPS) were obtained with a Kα Thermo Scientific
apparatus with an Al Kα (h = 1486.68 eV) X‐ray source using a voltage of 12 kV
under vacuum (2x10−7 mbar). Due to the lack of C1s signal, binding energies
were calibrated relative to the O1s peak of stoichiometric WO3 at 530.2 eV
[14,15].
Diffuse reflectance UV‐Vis spectroscopy (DR‐UV‐Vis) measurements, useful
in the determination of the semiconductor band gap, were performed with a
UV‐Vis‐NIR Cary 5000 spectrophotometer (Varian‐Agilent Technologies)
equipped with an integrating sphere device.
4.2.3. Catalytic activity measurements
Ibuprofen sodium salt (IBP), was used as target compound to test the
catalytic activity of the synthesized materials. Photocatalytic experiments were
carried out in semi‐batch mode in a laboratory‐scale system consisting of a 0.5 L
glass‐made spherical reactor, provided with a gas inlet, a gas outlet and a liquid
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
147
sampling port. The reactor was placed in the chamber of a commercial solar
simulator (Suntest CPS, Atlas) provided with a 1500 W air‐cooled Xe arc lamp
with emission restricted to visible light of wavelengths over 390 nm using
quartz, glass and polyester cut‐off filters. The irradiation intensity was kept at
550 W m−2 and the temperature of the system was maintained between 20 and 40
°C throughout the experiments. If required, a laboratory ozone generator
(Anseros Ozomat Com AD‐02) was used to produce a gaseous ozone‐oxygen
stream that was fed to the reactor. In that case, the ozone concentration was
monitored by an Anseros Ozomat GM‐6000Pro gas analyzer.
In a typical photocatalytic ozonation experiment, the reactor was first loaded
with 0.5 L of an aqueous solution containing 10 mg L‐1 of IBP at pH0 = 6.5 (not
buffered). Then, 0.125 g of the catalyst were added and the suspension was
stirred in the darkness for 30 min. After this dark stage, the lamp was switched
on and, simultaneously, a mixture of ozone‐oxygen (10 mg L‐1 ozone
concentration) was fed to the reactor at a flow rate of 20 L h‐1. The irradiation
time for each experiment was 2 h. Samples were withdrawn from the reactor at
intervals and filtered through a 0.2 μm PET membrane to remove the
photocatalyst particles except for dissolved ozone analysis.
Experiments of adsorption (i.e., absence of radiation and ozone), photolysis
(i.e., radiation in absence of catalysts and ozone), photocatalytic oxidation (i.e.,
radiation and catalyst in absence of ozone), ozonation (i.e., absence of radiation
and catalyst), and catalytic ozonation (i.e., absence of radiation) were also
carried out for comparative analysis.
In addition, the effectiveness of the best photocatalytic system was tested in a
more realistic application under complete simulated solar light radiation
(wavelengths from 300 nm). Ten emerging contaminants: acetaminophen (AAP),
metoprolol (MTP), caffeine (CAF), hydrochlorothiazide (HCT), antipyrine
(ANT), sulfamethoxazole (SFM), carbamazepine (CAR), ketorolac (KET),
diclofenac (DCF) and ibuprofen (IBP), frequently found in municipal
CAPÍTULO 4 (CHAPTER 4)
148
wastewater effluents (MWW), were selected. They were added at initial
concentration of 0.5 mg L‐1 each to a MWW taken from Badajoz MWWTP
(Badajoz, Spain) designed for 225,000 equivalent inhabitants with an average
inlet flow of 37,500 m3 day‐1. Effluents were collected downstream of the
MWWTP secondary biological treatment and filtered. The carbonate‐bicarbonate
content was eliminated by air stripping after acidifying the MWW. Then pH was
restored and MWW was stored at ‐20 °C until use.
IBP concentration was analyzed by HPLC‐DAD (Hitachi, Elite LaChrom)
using a Phenomenex C‐18 column (5 μm, 150 mm long, 3 mm diameter) as
stationary phase and 0.7 mL min−1 of acetonitrile‐acidified water (0.1 % H3PO4)
as mobile phase (60 ‐ 40 v/v, isocratic). Identification and quantification was
carried out at 220 nm. ECs concentration in MWW matrix were analyzed by the
same HPLC system with 0.2 mL min‐1 of acetonitrile‐acidified water (0.1 %
formic acid) as mobile phase with a gradient from 10 % to 100 % in acetonitrile
in 40 min and 20 min re‐equilibration time. Identification and quantification was
carried out at the maximum wavelength of each compound. Total organic
carbon (TOC) was measured using a Shimadzu TOC‐VSCH analyzer. Aqueous
ozone was measured by following the indigo method using a UV‐Vis
spectrophotometer (Evolution 201, Thermospectronic) set at 600 nm [16]. Ozone
in the gas phase was continuously monitored by means of an Anseros Ozomat
GM‐6000Pro analyzer. Hydrogen peroxide concentration was determined
photometrically by the cobalt/bicarbonate method at 260 nm using a UV‐Vis
spectrophotometer (Evolution 201, Thermospectronic) [17].
4.3. RESULTS AND DISCUSSION
4.3.1. Characterization of the photocatalysts
Tungstite is a hydrous tungsten oxide, H2O∙WO3 that crystallizes in the
orthorhombic system and can be thermally decomposed into different WO3
structures. WO3 consists of a three‐dimensional array of corner sharing metal‐
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
149
oxygen octahedral in a chess‐board type of array. The ideal structure is a ReO3‐
type cubic form but this is easily distorted away since the relatively small
tungsten atoms tend to be displaced from the center of the octahedral. These
displacements are temperature dependent and also influenced by the presence
of impurities which make WO3 to adopt a variety of symmetry types:
monoclinic, triclinic, orthorhombic and tetragonal. In addition, another two
metastable structures of WO3 have been obtained at low temperatures:
hexagonal and pyrochlore [18].
Thus, knowing the thermal behavior of the commercial tungstite precursor is
mandatory to understand the crystallization of the different WO3 structures.
Figure 4.1 shows the TGA‐DTA profiles for the H2WO4 precursor. The evolution
can be divided in 4 different regions. First of all, from ambient temperature to
150 °C, a small weight loss around 1 % is observed accompanied by an
endothermic contribution which can be assigned to the evaporation of adsorbed
H2O. In region II, from 150 to 325 °C, a noticeable weight loss is obtained
together with an endothermic peak at 225 °C which corresponds with the loss of
coordinated H2O from the tungstite structure to form cubic WO3 [11,12]. The
cubic WO3 structure is metastable and it can be stabilized by the presence of Na+
impurities [19]. However, at this temperature, formation of different structures
such as monoclinic WO3 cannot be disregarded as it is expected from the
absence of impurities in the commercial precursor used here. On the other hand,
in the range 325 ‐ 500 °C (region III), it can be observed a slight endothermic
contribution around 371 °C followed by an exothermic peak with a maximum
centered at 447 °C together with a slight weight loss (around 0.2 %). In this
region, the complete transformation of cubic WO3 into monoclinic WO3 occurs
[11,12]. No significant changes are observed in region IV in the range 500 ‐ 600
°C. From these results, the heat‐treatment temperature of the precursor was set
at 300 and 450 °C to obtain different crystalline phases composition of the
catalysts. The time of the heat‐treatment was relatively short (1 ‐ 30 min)
following the work of Cao et al. [11]. One sample was also heat‐treated during 60
CAPÍTULO 4 (CHAPTER 4)
150
min for comparative purposes.
0 100 200 300 400 500 6000.93
0.94
0.95
0.96
0.97
0.98
0.99
1.00
III
M/M
0
TEMPERATURE (oC)
TGA
I
IVII-10
-8
-6
-4
-2
0
2
4
6
8
10
DT
A DTA
Figure 4.1. TGA‐DTA profiles of H2WO4 precursor.
The structural characterization of the catalysts was performed by XRD as
shown in Figure 4.2 where XRD patterns of the tungstite H2O∙WO3
orthorhombic phase (JCPDS 018‐1418), cubic WO3 (JCPDS 41‐0905) and
monoclinic WO3 (JCPDS 043‐1035) have been plotted for comparison. It can be
noticed that the main peaks in the diffraction pattern of the precursor, W‐0, are
attributable to the orthorhombic phase of tungstite, although a small shoulder
around 24.2° can also indicate a low contribution of the cubic WO3 phase. When
the precursor is heat‐treated at 300 °C the cubic structure of WO3 is
progressively formed as the time of calcination increases. Thus, for 1 min at 300
°C (W300‐1) it can be observed the coexistence of tungstite orthorhombic phase
together with cubic WO3, as the main peaks of tungstite at 16.5° and 25.6°
decreased in favor of the increasing cubic diffraction peaks. The transformation
of tungstite into cubic WO3 has been completed in the W300‐3 catalyst, where the
main tungstite diffraction peaks have completely disappeared and all the
diffraction peaks are attributable to cubic crystalline structure. After that, an
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
151
increase in the time of calcination at 300 °C provokes the formation of
monoclinic WO3 to some extent. This can be noticed by the presence of the
contributions centered at 26.5° and 28.6°. These are characteristics of the
monoclinic crystalline structure, together with the widening of the peaks at 23.9
and 34.0° that are affected by the overlapped cubic‐monoclinic contributions,
with the formation of the triple peaks in monoclinic phase at these 2 values
corresponding to reflections from (001), (020) and (200) planes [14]. It is also
noticeable that an increase in the calcination time from 5 to 60 min did not
produce any significant change in the XRD patterns (W300‐5 and W300‐60) which
is indicative of the stabilization of the cubic structure at this temperature.
Finally, the temperature increase from 300 to 450 °C provoked the
transformation of the cubic phase of WO3 into monoclinic to a greater extent
according to previous TGA‐DTA results. Table 4.1 shows the semiquantitative
analysis of crystalline phases in the catalysts.
Table 4.1 also summarizes the BET surface area values of the catalysts from
N2 adsorption‐desorption isotherms. In general, small surface areas are obtained
in this type of materials unless different strategies are used for this purpose. As
can be observed, the surface areas undergo a slight increase (from c.a. 30 to 40
m2 g‐1) when the dehydration of tungstite occurs and cubic crystalline structure
is formed. On the other hand, at higher calcination temperature (450 °C), a
decrease of the surface area is observed in the catalyst, being higher as the time
of the heat treatment increases.
CAPÍTULO 4 (CHAPTER 4)
152
10 20 30 40 50 60 70
W300
-60
W450
-30
W450
-5
W300
-5
W300
-3
W300
-1
2 (º)
INT
EN
SIT
Y (
a.u.
)
W-0
H2WO
4, tungstite (orthorhombic)
018-1418
WO3, cubic
41-0905
WO3, monoclinic
043-1035
Figure 4.2. XRD patterns of the catalysts.
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
153
The pH of the point of zero charge was also determined for the catalysts
studied and the results are summarized in Table 4.1. The acidity of the surface
can play a significant role in the behavior of WO3 based materials in wastewater
treatments related to adsorption processes of organic compounds and species
with unpaired electrons [20]. It can be noticed a slight decrease in the value of
pHPZC together with the transformation of tungstite into WO3 cubic structure so
that the catalysts with this majority crystalline phase have a pHPZC value around
4. On the other hand, an increase in the calcination temperature up to 450 °C led
to pHPZC values around 5 for W450‐5 and W450‐30 catalysts. The decrease of the
surface acidity with increasing the calcination temperature has been previously
reported in monoclinic WO3 [21] and WO3‐ZrO2 catalysts [22], and has been
attributed to the creation of vacancies in the WO3 structure from the extraction
of lattice oxygen, which leads to a decrease in the number of Lewis acid sites
[21]. In addition, a higher calcination temperature could also produce the
dehydration/dehydroxylation of the WO3 surface to some extent lowering the
number of Brönsted acid sites [21].
Surface characterization of the catalysts was performed by XPS to analyze the
stoichiometry and chemical binding states of W. XPS spectra of W4f region are
depicted in Figure 4.S1 of the supplementary information. In all cases, the
spectra obtained showed a doublet with binding energies around 35.5 and 37.6
eV for W 4f7/2 and W 4f5/2, respectively, with the W 4f7/2‐W4f5/2 binding energy
separation being 2.1 eV. The energy position of this doublet corresponds to the
W6+ oxidation state in WO3 [14,15,23]. On the other hand, XPS spectra of O1s
spectral region for all the catalysts are shown in Figure 4.S2. The main
contribution at 530.2 eV has been assigned to oxygen in the WO3 lattice
[14,15,24]. Moreover, an additional contribution centered at 532.5 eV has been
observed for W‐0 and W300‐1 catalysts, which has been assigned to oxygen in
water molecules bound in the tungstite structure (H2O∙WO3) or absorbed in the
catalyst surface [14]. This is consistent with TGA‐DTA and XRD results having
CAPÍTULO 4 (CHAPTER 4)
154
these two catalysts more than 80 % of tungstite in their structure.
The diffuse reflectance UV‐Vis spectra of the photocatalysts (Figure 4.3)
showed a higher optical absorbance in the visible region up to ca. 530 nm for
tungstite composed catalysts (W‐0 and W300‐1), whereas decreased to ca. 480 nm
for cubic and monoclinic WO3 composed catalysts (W300‐3, W300‐5, W450‐5, W450‐
30). The optical energy band gap (Eg) was calculated by means of Tauc’s
expression in Figure 4.S3 (supplementary information), and results are
summarized in Table 4.1. These values are approximate due to the need of
extrapolation of the resulting curve as can be observed in Fig. 4.S3 for all the
catalysts. The Eg value for tungstite H2WO4 (major component of W‐0 and W300‐
1) was found to be 2.42 eV which is increased up to 2.64 when the precursor is
completely transformed into cubic WO3 structure (in W300‐3 catalyst). Then, the
Eg value progressively increased up to 2.67 eV as monoclinic WO3 was formed.
These results are quite similar to previously reported values [7,11].
200 300 400 500 600 700 800 9000.00
0.25
0.50
0.75
1.00
1.25
1.50
AB
SO
RB
AN
CE
(u
.a.)
WAVELENGTH (nm)
W-0 W
300-1
W300
-3
W300
-5
W450
-5
W450
-30
Figure 4.3. DR‐UV‐Vis spectra of the catalysts.
4.3.2. Comparison of processes
The effectiveness of the photocatalysts in the use of visible light radiation
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
155
combined with ozone was tested using IBP as target compound by cutting off all
the wavelengths lower than 390 nm in the solar simulator. First, the performance
of the combined process (photocatalytic ozonation) respect to the single
treatments has been checked with all the catalysts. Results obtained with W450‐5
catalyst are depicted in Figure 4.4 and Figure 4.5, being this catalyst
representative of the behavior of the most active ones. Fig. 4.4(A) shows the time
evolution of IBP dimensionless concentration upon the different treatments
applied. First of all, the absence of IBP depletion due to direct photolysis with
visible light radiation was confirmed which is consistent with the UV‐Vis
absorption spectrum of this compound (supplementary Figure 4.S4). On the
other hand, IBP adsorption capacity of the W450‐5 catalyst at the conditions
studied here was somewhat noticeable, reaching around 10 % of IBP removal
after 2 h contact time. The depletion rate of IBP is increased in photocatalytic
oxidation reaching 25 % of IBP removal after 2 h of treatment although it can be
noticed the low efficiency of the process as a consequence of the high
recombination rate in WO3 materials used in the presence of oxygen [5]. When
O3 is present, all the treatments led to complete IBP removal in less than 1 h of
reaction time (with conversion higher than 99 %) regardless of the
presence/absence of catalyst and/or radiation in contrast to O3‐free treatments.
Although the evolution of IBP is quite similar for all the O3‐treatments, some
slight differences can be observed in the depletion rate. Single ozonation and
photolytic ozonation presented quite similar IBP depletion rate. On the other
hand, the presence of the W450‐5 catalysts and ozone in dark conditions (catalytic
ozonation) did not improve the degradation of IBP thus indicating that this
catalyst does not exert any catalytic effect in ozone decomposition into reactive
species such as hydroxyl radicals or produces an inefficient decomposition for
IBP degradation. Photocatalytic ozonation gave place to the highest IBP
depletion rate reaching complete IBP removal in 20 min compared to 40 ‐ 50 min
necessary for the other O3 treatments.
CAPÍTULO 4 (CHAPTER 4)
156
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Dark
Adsorption Photolysis Photocatalytic oxidation Ozonation Catalytic ozonation Photolytic ozonation Photocatalytic ozonation
CIB
P/C
IBP
,0
TIME (min)
Figure 4.4. (A) IBP and (B) TOC dimensionless concentration evolution during all the
treatments applied with W450‐5 catalyst (lines show trends). Conditions: pH0 = 6.5, T = 20 ‐
‐ 40 °C, CIBP,0 = 10 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 20 L h‐1 (O2 or O3/O2).
On the other hand, it is known that the oxidation of IBP proceeds through
different steps giving place to different intermediate compounds which are
eventually transformed into CO2 and H2O (complete mineralization). Thus, the
(A)
(B)
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Dark
TO
C/T
OC
0
TIME (min)
0
0
CT
OC
/CT
OC
,0
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
157
mineralization of IBP has been followed and presented in Fig. 4.4(B) as
normalized TOC concentration. As expected from previous IBP results, neither
adsorption nor direct photolysis led to significant mineralization percentages.
Photocatalytic oxidation also gave place to negligible mineralization according
to the IBP conversion reached, lower than 25 %. Regarding the O3‐treatments,
noticeable differences have been found depending on the presence/absence of
radiation and/or photocatalyst. Single ozonation led to 25 % of mineralization at
60 min and then stopped as a consequence of the formation of some
intermediate compounds that are refractory to ozone direct reaction (mainly
short‐chain organic acids) [25], and/or due to the absence of IBP in the reaction
medium which eventually produces hydroxyl radicals through IBP‐O3 direct
reaction as discussed below. Fairly similar results were observed during
photolytic ozonation under visible light radiation. On the other hand, the
presence of the catalyst combined with ozone in dark conditions (catalytic
ozonation) did not improve the previous results. This process showed the
poorest behavior among the O3‐treatments in terms of mineralization, with
negligible TOC conversion during the reaction time, confirming the absence of
any positive catalytic effect in the reaction. In fact, the low mineralization
reached compared to ozone alone treatment is likely indicative of an inefficient
consumption of O3 in the catalyst surface in the darkness. These results are
contrary to those found for the mineralization of phenol with a commercial WO3
(monoclinic) catalyst [5], which points out that the target compound nature can
play a key role in the interactions compound‐catalyst surface during catalytic
ozonation. On the contrary, photocatalytic ozonation showed the highest
mineralization rate, leading to 87 % mineralization at 120 min reaction time.
These results also point out the synergism produced between ozone and visible
light irradiated WO3 since final TOC removal of the combined process is much
higher than the sum of these values for the individual processes (87 % in
O3/WO3/Vis vs. 5 % in O2/WO3/Vis and 25 % in O3), confirming also the ability of
O3 to capture the electrons on the WO3 surface compared to O2 [5].
CAPÍTULO 4 (CHAPTER 4)
158
0 20 40 60 80 100 1200
1x10-5
2x10-5
3x10-5
4x10-5
5x10-5
6x10-5
CO
3 (M
)
TIME (min)
0 20 40 60 80 100 1200
1x10-5
2x10-5
3x10-5
4x10-5
5x10-5 Photolysis Photocatalytic oxidation Ozonation Catalytic ozonation Photolytic ozonation Photocatalytic ozonation
CH
2O2 (M
)
TIME (min)
Figure 4.5. (A) Dissolved O3 concentration and (B) H2O2 concentration during all the
treatments applied with W450‐5 catalyst (lines show trends). Conditions: pH0 = 6.5, T = 20 ‐
‐ 40 °C, CIBP,0 = 10 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 20 L h‐1 (O2 or O3/O2).
In addition, time profiles of dissolved ozone and hydrogen peroxide
concentration during the treatments have been plotted in Figure 4.5. Regarding
(A)
(B)
PAPER 2: Influence of structural properties on the
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159
dissolved ozone, in Fig. 4.5(A) it is seen that ozone accumulated in solution
during the single ozonation experiment reaching a maximum concentration at
about 60 min reaction time. Different profiles were observed in the other O3‐
treatments in which ozone concentration increased in the first minutes of the
experiment and then remained almost constant. The ozone concentration was
always lower in catalytic ozonation than during single ozonation. This
observation together with the low mineralization reached, is an additional
evidence of an inefficient decomposition of O3 on the catalyst surface in the
darkness. On the other hand, in the photocatalytic ozonation process, the
maximum concentration of dissolved ozone was noticeably lower, leading to a
nearly steady concentration of about 1x10‐5 M in contrast to 5x10‐5 M during
single ozonation. The lowest values of dissolved ozone together with the highest
mineralization rate observed in this treatment suggest that O3 has been
consumed on the WO3 catalyst surface through its reaction with electrons
generated in the photocatalytic process, thus enhancing the production of
oxidizing species such as hydroxyl radicals, as with TiO2 catalysts [1,2,26].
On the other hand, hydrogen peroxide is commonly formed through direct
ozone reactions [27‐29]. This has been experimentally confirmed in this work for
IBP, as shown in Fig. 4.5(B). Thus, during single ozonation an increase in H2O2
concentration up to 45 min, when IBP was completely removed, can be
observed. Then, H2O2 concentration remained almost constant until the end of
the experiment. The concentration of H2O2 formed during catalytic and
photolytic ozonation was somewhat lower. On the other hand, during
photocatalytic ozonation the formation of H2O2 takes place at higher rate
reaching a maximum concentration at about 20 ‐ 30 min when IBP was
completely removed. Then the consumption of H2O2 was faster and its
concentration negligible at the end of the treatment. These results point out that
H2O2 is being consumed through photocatalytic reactions probably acting as an
electron acceptor in the WO3 surface thus increasing the production of hydroxyl
CAPÍTULO 4 (CHAPTER 4)
160
radicals and avoiding recombination to some extent, as also occurs with TiO2
[26]. Finally, the formation of H2O2 during photocatalytic oxidation was much
lower than in ozone processes confirming the production of H2O2 mainly
through direct IBP‐O3 reaction.
Therefore, compared to individual processes, photocatalytic ozonation
treatment with WO3 and visible‐light radiation led to a faster IBP depletion with
the highest mineralization rate due to the synergistic effect between O3 and
irradiated WO3.
4.3.2.1. Simplified mechanistic approach of IBP photocatalytic ozonation
Direct ozone reactions in water follow second order kinetics and the kinetic
regime can be established through the determination of the Hatta number
according to [25]:
3 3IBP O O IBP
L
k D CHa
k
(4.2)
where DO3 = 1.3x10‐9 m2 s‐1 is the ozone diffusivity in water [30], and kL = 5x10‐5
m s‐1 is the individual liquid phase mass transfer coefficient for a bubbled stirred
tank [31], being CIBP the molar IBP concentration and kIBP‐O3 = 9.1 M‐1 s‐1 the direct
reaction rate constant at pH = 7 [32]. The highest calculated Ha value was 0.023
for the initial conditions thus confirming slow kinetic regime of ozone
absorption. Thus, the depletion rate of IBP through direct O3 reaction is
described as follows:
3 3
3
IBPIBP O O IBP
O
dC 1k C C
dt z
(4.3)
where z = 1 is the stoichiometric coefficient of the reaction in mol O3/mol IBP
and CO3 is the dissolved ozone concentration. The differential equation (4.3) was
solved using the value of the kinetic constant and the experimental values of CO3
PAPER 2: Influence of structural properties on the
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161
by means of 4th order Runge‐Kutta. The results of simulated dimensionless CIBP
together with the experimental values during single ozonation are depicted in
Figure 4.6. It can be clearly noticed that the experimental rate of IBP depletion is
higher than the expected results (simulated), thus indicating that additional
reactions may develop and contribute to IBP removal. Indirect reactions due to
O3 decomposition are initiated through the following steps:
3 2 2O HO HO O (4.4)
pKa 11.3
2 2 2H O HO H (4.5)
‐2 3 23HO O O HO (4.6)
pKa 4.8 ‐
2 2HO O H
(4.7)
‐ ‐32 3O O O (4.8)
‐3O H ... HO
(4.9)
These reactions lead to the formation of hydroxyl radicals ( HO) capable of
oxidizing IBP in water. However, the pH of the reaction varied from 6.5 to 4.5
and therefore, reaction (4.4) between 3O / HO seems not to play an important
role on the IBP disappearance. In addition, an accumulation of H2O2 was
observed during the single ozonation experiment (Fig. 4.5(B)), thus indicating
that reaction (4.6) between 3 2O / HO can be also disregarded. These evidences
together with the evolution of mineralization whereas IBP is still present in the
reaction medium lead to hypothesize the formation of ozonide radical through
IBP‐O3 direct reaction, as it has been previously reported for phenol or
diclofenac [28,33,34]:
‐33 OINTOIBP (4.10)
which immediately results in the formation of HOthrough reaction (4.9), being
CAPÍTULO 4 (CHAPTER 4)
162
0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
CIB
P/C
IBP
,0
TIME (min)
Experimental Simulated O
3
Simulated O3+HO·
INT any intermediate organic compound.
Figure 4.6. Experimental and simulated IBP dimensionless concentration evolution
during single ozonation. Conditions: pH0 = 6.5, T = 20 ‐ 40 °C, CIBP,0 = 10 mg L‐1, CO3,g inlet
= 10 mg L‐1, Qg = 20 L h‐1 (O3/O2).
With these considerations, the depletion rate of IBP through direct O3
reaction taking into account the formation of HO during this reaction is
described as follows:
3 3
3
IBPIBP O O IBP IBPIBP HO HO
O ‐HO
dCk C C k C C
dt
(4.11)
3 3
3
HOIBP O O IBP IBPIBP HO HO
IBP‐O
dCk C C k C C
dt
(4.12)
where kIBP‐HO∙ is the rate constant between hydroxyl radical and IBP and CHO∙ is
the hydroxyl radical concentration. The differential equations (4.11) and (4.12)
were simultaneously solved using the value of the kinetic constants and the
experimental values of CO3 by means of 4th order Runge‐Kutta (kIBP‐HO∙ = 7.4x109
PAPER 2: Influence of structural properties on the
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163
M‐1 s‐1 [32]). These results are also depicted in Figure 4.6, where it can be
observed that the experimental IBP depletion rate is still higher than the
simulated values. This could be related to the accuracy of the rate constants used
for the simulation and/or the development of additional IBP degradation routes.
The confirmation of the different degradation ways needs extra experiments out
of the scope of this work.
According to this and the previous findings, it can be postulated that
photocatalytic ozonation of IBP with WO3 catalyst takes place through reactions
(4.13) ‐ (4.20), where both O3 and HO are responsible for IBP disappearance and
photogenerated oxidizing species ( HO and/or positive holes) are the main
responsible for mineralization.
3 2 2IBP O ... INT H O HO (4.13)
h3 3 3WO WO e WO h
(4.14)
3 3WO e O ... HO (4.15)
3 2 2WO e H O ... HO (4.16)
3WO h IBP INT (4.17)
HO IBP INT (4.18)
3 2 2WO h INT ... CO H O (4.19)
2 2HO INT ... CO H O (4.20)
Reaction (4.13) summarizes the direct IBP‐O3 steps that would lead to some
large organic intermediates together with the H2O2 detected and accumulated
during single ozonation experiment, and eventually with the formation of HO .
On the other hand, visible light radiation provokes the creation of electron‐hole
pairs (reaction (4.14)) and electrons are proposed to react with dissolved O3 and
CAPÍTULO 4 (CHAPTER 4)
164
H2O2 through reactions (4.15) and (4.16), according to the profiles of these
species during photocatalytic ozonation versus single ozonation. Finally,
photogenerated holes and hydroxyl radicals can oxidize both IBP and any INT
to achieve the complete conversion into CO2 and H2O according to reactions
(4.17) ‐ (4.20).
The relative contribution of each step to the overall process would depend on
the experimental conditions (e.g. pH, ozone dose or catalyst loading), and the
confirmation of the intermediate species formed requires additional
experimental work that will be the subject of a future work.
4.3.3. Catalysts screening for photocatalytic ozonation under visible‐light
radiation
All the catalysts prepared were tested in the photocatalytic ozonation of IBP
under visible‐light radiation. The results are depicted in Figure 4.7 and Figure
4.8 together with single ozonation results for the sake of comparison. Regarding
the evolution of IBP (Fig. 4.7(A)) results are quite similar although some trends
can be extracted. First of all, slight differences were observed in the IBP
adsorption capacity of the catalysts during the dark stage that is generally
consistent with the values of pHPZC obtained where a higher surface acid
character led to a higher amount of initial IBP adsorbed. However, these
differences seem not to play a crucial role in the behavior of the catalysts. On the
other hand, it can be noticed that W‐0 and W300‐1 catalysts gave place to similar
IBP depletion rate than ozone alone whereas the rest of the catalysts gave place
to a slight higher depletion rate. In general, the catalytic activity is increased as
the heat‐treatment in time and/or temperature was increased. Main differences
between ozone and O3‐photocatalytic processes were found in terms of
mineralization (see Fig. 4.7(B)). Ozone alone gave place to a 25 % mineralization
at about 60 min reaction time. On the other hand, the tungstite precursor, W‐0,
and the subsequent catalysts W300‐1 and W300‐3 showed a very different
normalized TOC profile. It can be noticed a first step in which mineralization is
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
165
slow, with a TOC removal rate fairly lower than during ozone alone reaction up
to 60 min. After this time, the mineralization rate increased achieving around 35
‐ 40 % of mineralization at 120 min with the three catalysts. It can be noticed that
in addition to the differences in the properties of these three catalysts (Eg, SBET,
pHPZC, etc.), they have in common that no monoclinic WO3 crystallization has
occurred during the heat‐treatment applied. A significant increase in the
catalytic activity is observed with the W300‐5 catalyst in terms of mineralization,
reaching around 70 % TOC removal at 120 min. The main difference between
this and the previous catalysts is the presence of monoclinic crystalline phase to
some extent in its structure according to XRD results. The best results were
obtained with the 450 °C heat‐treated catalysts reaching around 85 ‐ 87 %
mineralization at 120 min reaction time. In these catalysts, a contribution of
monoclinic WO3 higher than 75 % was observed together with a higher pHPZC
which could indicate the presence of oxygen vacancies to some extent, although
this could not be confirmed by XPS. The differences found in the mineralization
profile of W450‐5 and W450‐30 catalysts, the first one leading to the highest
mineralization rate, could be related to the higher SBET found in this catalyst.
In addition, the evolution of dissolved ozone and hydrogen peroxide
concentration formed during the reaction has been plotted in Figure 4.8.
Regarding dissolved ozone (Fig. 4.8(A)), it is noticeable the lowest values of the
steady concentration with the most active catalysts, W450‐5 and W450‐30,
compared to the single ozonation reaction. These results suggest an efficient O3
consumption on the catalysts surface through its reaction with the electrons
generated in the photocatalytic process. Fig. 4.8(B) shows the hydrogen peroxide
concentration evolution for all the catalysts where a faster decomposition rate is
also observed for the most active ones (W450‐5 and W450‐30), thus being likely
indicative of a higher consumption of H2O2 in surface reactions as electron
acceptor.
Therefore, at similar textural and optical properties, the monoclinic structure
CAPÍTULO 4 (CHAPTER 4)
166
-20 0 20 40 600.0
0.2
0.4
0.6
0.8
1.0
Dark
W-0 W300-1
W300-3
W300-5
W450-5
W450-30
O3
CIB
P/C
IBP
,0
TIME (min)
in WO3 catalysts seems to be beneficial for the photocatalytic ozonation process
under visible light.
Figure 4.7. (A) IBP and (B) TOC dimensionless concentration evolution during
photocatalytic ozonation with all the catalysts studied (lines show trends). Conditions:
pH0 = 6.5, T = 20 ‐ 40 °C, CIBP,0 = 10 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 20 L
h‐1 (O3/O2).
(A)
(B)
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Dark
TO
C/T
OC
0
TIME (min)
0
0
CT
OC
/CT
OC
,0
PAPER 2: Influence of structural properties on the
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167
0 20 40 60 80 100 1200
1x10-5
2x10-5
3x10-5
4x10-5
5x10-5
6x10-5
CO
3 (M
)
TIME (min)
0 20 40 60 80 100 1200
1x10-5
2x10-5
3x10-5
4x10-5
5x10-5
6x10-5
W-0 W300-1
W300-3
W300-5
W450-5
W450-30
O3
CH
2O2 (M
)
TIME (min)
Figure 4.8. (A) Dissolved O3 concentration and (B) H2O2 concentration during
photocatalytic ozonation with all the catalysts studied (lines show trends). Conditions:
pH0 = 6.5, T = 20 ‐ 40 °C, CIBP,0 = 10 mg L‐1, CWO3 = 0.25 g L‐1, CO3 g inlet = 10 mg L‐1, Qg = 20 L
h‐1 (O3/O2).
(A)
(B)
CAPÍTULO 4 (CHAPTER 4)
168
4.3.4. Simulated solar light photocatalytic ozonation of ECs in MWW
The selected catalyst W450‐5 which presented the highest catalytic activity in
the photocatalytic ozonation of IBP with visible light was tested in the
photocatalytic ozonation of a mixture of ten ECs (AAP, MTP, CAF, HCT, ANT,
SFM, CAR, KET, DCF and IBP) in a real municipal wastewater effluent. Main
parameters of the MWW effluent are summarized in Table 4.2. The
photocatalytic ozonation treatment was carried out using the entire simulated
solar spectrum (λ > 300 nm). For comparative purposes also single ozonation
treatment was carried out. The evolution of the ECs is depicted in Figure 4.9. It
can be noticed that many of these compounds (AAP, ANT, SFM, CAR, KET and
DCF) presented similar depletion rate both in single ozonation and
photocatalytic ozonation treatments as a consequence of their high reactivity
towards O3. Thus, O3 itself is highly effective to remove these compounds.
Direct rate constants of ECs with O3 are summarized in Table 4.S1. On the other
hand, the ECs whose rate constant is low such as MTP, CAF, HCT and IBP, were
faster eliminated through photocatalytic ozonation (in less than 60 min), being
the differences found between ozone alone and the combined process higher the
lower the direct rate constant of O3‐EC reaction was. Again in the MWW matrix
main differences between ozone and photocatalytic ozonation were found in
terms of TOC removal as shown in Figure 4.10. As it can be seen, ozone alone
gave place to a lower degree of mineralization reaching less than 10 % TOC
removal after 120 min in contrast to more than 40 % for photocatalytic
ozonation. In addition, Table 4.2 also shows the characterization of the
wastewater after the treatments being noticeable the decrease in COD and
aromaticity when the combined treatment was applied. Finally, Figure 4.11
shows the evolution of dissolved O3 concentration together with the formation
of H2O2 during the treatments. The obtained profiles are consistent with the
previous results thus indicating the consumption of both O3 and H2O2 (formed
through O3‐ECs direct reaction) at the WO3 surface leading to the formation of
additional oxidizing species such as hydroxyl radicals, which improve the
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
169
mineralization reached. Thus, intermediate compounds formed from ECs
oxidation are expected to be more easily eliminated through photocatalytic
ozonation compared to single ozonation.
Therefore, the use of WO3 as catalyst for photocatalytic ozonation of ECs
containing MWW is an attractive alternative that improves the use of solar
radiation leading to a fast elimination of organic contaminants (mainly those
refractory to ozone reactions) and achieving an important mineralization degree.
Table 4.2. Characterization of MWW effluent.
PARAMETER Before treatment Ozonation Photocatalytic ozonation
TOC (mg C L‐1) 14.2 13.1 8.2
IC (mg C L‐1) 0.19 0.15 0.20
Turbidity (NTU) 3.0 n.m. n.m.
pH 6.5 5.9 5.7
Absorbance 254 nm 0.268 0.102 0.015
COD (mg O2 L‐1) 44 33 13
BOD5 (mg O2 L‐1) 28 n.m. n.m.
Phosphate (mg L‐1) 5 5 5
n.m.: not measured
CAPÍTULO 4 (CHAPTER 4)
170
-30 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
C/C
0
TIME (min)
Ozonation Photocatalytic ozonation
AAP
-30 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0MTP
C/C
0
TIME (min)
-30 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0CAF
C/C
0
TIME (min)-30 0 20 40 60 80 100 120
0.0
0.2
0.4
0.6
0.8
1.0
1.2
HCT
C/C
0
TIME (min)
-30 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0ANT
C/C
0
TIME (min)-30 0 20 40 60 80 100 120
0.0
0.2
0.4
0.6
0.8
1.0SFM
C/C
0
TIME (min)
-30 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0CAR
C/C
0
TIME (min)-30 0 20 40 60 80 100 120
0.0
0.2
0.4
0.6
0.8
1.0KET
C/C
0
TIME (min)
-30 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0DCF
C/C
0
TIME (min)-30 0 20 40 60 80 100 120
0.0
0.2
0.4
0.6
0.8
1.0IBP
C/C
0
TIME (min)
Figure 4.9. ECs dimensionless concentration during ozonation and photocatalytic
ozonation with W450‐5 catalyst in a MWW effluent (lines show trends). Conditions: pH0 =
6.5, T = 20 ‐ 40 °C, CECs,0 = 0.5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 20 L h‐1
(O3/O2).
PAPER 2: Influence of structural properties on the
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171
0 20 40 60 80 100 1200.0
4.0x10-6
8.0x10-6
1.2x10-5
1.6x10-5
2.0x10-5
2.4x10-5
CO
3 (M
)
TIME (min)
0 20 40 60 80 100 1200
1x10-5
2x10-5
3x10-5
4x10-5
5x10-5
6x10-5
7x10-5
Ozonation Photocatalytic ozonation
CH
2O2(
M)
TIME (min)
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
CT
OC
/CT
OC
,0
TIME (min)
Ozonation Photocatalytic ozonation
Dark
Figure 4.10. TOC dimensionless concentration evolution during ozonation and
photocatalytic ozonation with W450‐5 catalyst in a MWW effluent (lines show trends).
Conditions: pH0 = 6.5, T = 20 ‐ 40 °C, CECs,0 = 0.5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg
L‐1, Qg = 20 L h‐1 (O3/O2).
Figure 4.11. (A) Dissolved O3 concentration and (B) H2O2 concentration during ozonation
and photocatalytic ozonation with W450‐5 catalyst in MWW effluent (lines show trends).
Conditions: pH0 = 6.5, T = 20 ‐ 40 °C, CECs,0 = 0.5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet =
10 mg L‐1, Qg = 20 L h‐1 (O3/O2).
(A) (B)
CAPÍTULO 4 (CHAPTER 4)
172
4.4. CONCLUSIONS
WO3 catalysts with different structural and surface properties where
prepared by thermodecomposition of commercial tungstite (H2O∙WO3) at
different temperatures, 300 and 450 °C, and short times, from 1 to 30 min. This
heat‐treatment allows obtaining tungstite, cubic and monoclinic crystalline
phases of tungsten trioxide. In general, as time or temperature increased,
monoclinic structure is formed which is also accompanied by a reduction of
specific surface area and an increase in the presence of oxygen vacancies. These
properties, monoclinic structure together with the presence of oxygen vacancies
played a key role in the behavior of the WO3 catalysts in the photocatalytic
ozonation of IBP under visible light radiation, favoring the electron transport in
the catalyst surface to some extent, thus improving the efficiency of the process.
Among the treatments tested with the best catalytic system, photocatalytic
ozonation showed a synergistic effect between ozone and irradiated WO3
leading to a complete IBP depletion in less than 20 min with a degree of
mineralization of about 87 % at 120 min. Finally, the use of WO3 as catalyst for
photocatalytic ozonation of ECs in a MWW effluent with simulated solar
radiation led to a fast elimination of the contaminants in less than 60 min and a
mineralization degree higher than 40 % at 120 min. The mechanism and kinetics
of photocatalytic ozonation with WO3 catalysts will be the subject of further
work.
AKNOWLEDGEMENTS
This work has been supported by the Spanish Ministerio de Economía,
Industria y Competitividad (MINECO) and European Feder Funds through the
project CTQ2012‐35789‐C02‐01. Authors acknowledge the SACSS‐SAIUEX and
UAI‐ICP for the characterization analyses. E. Mena thanks the Consejería de
Empleo, Empresa e Innovación (Gobierno de Extremadura) and European Social
Fund for providing her a predoctoral FPI grant (Ref. PD12059).
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
173
REFERENCES
[1] Agustina, T.E.; Ang, H.M.; Vareek, V.K. “A review of synergistic effect of
photocatalysis and ozonation on wastewater treatment”. J. Photochem.
Photobiol. C: Photochem. Rev. 6 (2005) 264‐273.
[2] Augugliaro, V.; Litter, M.; Palmisano, L.; Soria, J. “The combination of
heterogeneous photocatalysis with chemical and physical operations: A tool for
improving the photoprocess performance”. J. Photochem. Photobiol. C
Photochem. Rev. 7 (2006) 127‐144.
[3] Hernández‐Alonso, M.D.; Fresno, F.; Suárez, S.; Coronado, J.M.
“Development of alternative photocatalysts to TiO2: Challenges and
opportunities”. Energ. Environ. Sci. 2 (2009) 1231‐1257.
[4] Malato, S.; Fernández‐Ibáñez, P.; Maldonado, M.I.; Blanco, J.; Gernjak, W.
“Decontamination and disinfection of water by solar photocatalysis: Recent
overview and trends”. Catal. Today 147 (2009) 1‐59.
[5] Nishimoto, S.; Mano, T.; Kameshima, Y.; Miyake, M. “Photocatalytic water
treatment over WO3 under visible light irradiation combined with ozonation”.
Chem. Phys. Letters 500 (2010) 86‐89.
[6] Abe, R.; Takami, H.; Murakami, N.; Ohtani, B. “Pristine simple oxides as
visible light driven photocatalysts: Highly efficient decomposition of organic
compounds over platinum‐loaded tungsten oxideʺ. J. Am. Chem. Soc. 130 (2008)
7780‐7781.
[7] González‐Borrero, P.P.; Sato, F.; Medina, A.N.; Baesso, M.L.; Bento, A.C.;
Baldissera, G.; Persson, C.; Niklasson, G.A.; Granqvist, C.G.; Ferreira da Silva, A.
“Optical band‐gap determination of nanostructured WO3 film”. Appl. Phys.
Letters 96 (2010) 061909‐1‐061909‐3.
[8] Mano, T.; Nishimoto, S.; Kameshima, Y.; Miyake, M. “Investigation of
photocatalytic ozonation treatment of water over WO3 under visible light
irradiation”. J. Ceramic. Soc. Japan 119 (2011) 822‐827.
[9] Barceló, D.; Petrovic, M. (Eds.). “Emerging contaminants from industrial and
municipal wastes. Occurrence, analysis and effects”. The Handbook of
Environmental Chemistry 5‐S1. Springer, Berlin (Germany), 2008.
[10] Santos, J.L.; Aparicio, I.; Callejón, M.; Alonso, E. “Occurrence of
pharmaceutically active compounds during 1‐year period in wastewaters from
four wastewater treatment plants in Seville (Spain)”. J. Hazard. Mater. 164 (2009)
CAPÍTULO 4 (CHAPTER 4)
174
1509‐1516.
[11] Cao, J.; Luo, B. Lin, H.; Xu, B.; Chen, S. “Thermodecomposition synthesis of
WO3/H2WO4 heterostructures with enhanced visible light photocatalytic
properties”. Appl. Catal. B Environ. 111‐112 (2012) 288‐296.
[12] Guery, C.; Choquet, C.; Dujeancourt, F.; Tarascon, J.M.; Lassegues, J.C.
“Infrared and X‐ray studies of hydrogen intercalation in different tungsten
trioxides and tungsten trioxide hydrates”. J. Solid State Electrochem. 1 (1997)
199‐207.
[13] Subramanian, S.; Noh, J.S.; Schwarz, J.A. “Determination of the point of
zero‐charge of composite oxides”. J. Catal. 114 (1988) 433‐439.
[14] Senthil, K.; Yong, K. “Growth and characterization of stoichiometric
tungsten oxide nanorods by thermal evaporation and subsequent annealing”.
Nanotechnology 18 (2007) 395604‐1‐395604‐7.
[15] Pang, H.F.; Li, Z.J.; Xiang, X.; Fu, Y.Q.; Placido, F.; Zu, X.T. “Hierarchical
structured tungsten oxide nanocrystals via hydrothermal route: microstructure,
formation mechanism and humidity sensing”. Appl. Phys. A 112 (2013) 1033‐
1042.
[16] Bader, H.; Hoigné, J. “Determination of ozone in water by the indigo
method”. Water Res. 15 (1981) 449‐456.
[17] Masschelein, W.; Denis, M.; Ledent, R. “Spectrophotometric determination
of residual hydrogen peroxide”. Water Management. Water Sewage Works,
August (1977) 69‐72.
[18] Tilley, R.J.D. “The crystal chemistry of the higher tungsten oxides”. Int. J.
Refractory Metals Hard Mater. 13 (1995) 93‐109.
[19] Szilágyi, I.M.; Fórizs, B.; Rosseler, O.; Szegedi, A.; Németh, P.; Király, P.;
Tárkányi, G.; Vajna, B.; Varga‐Josepovits, K.; László, K.; Tóth, A.L.; Baranyai, P.;
Leskelä, M. “WO3 photocatalysts: Influence of structure and composition”. J.
Catal. 294 (2012) 119‐127.
[20] Tomova, D.; Iliev, V.; Rakovsky, S.; Anachkov, M.; Eliyas, A.; Li Puma, G.
“Photocatalytic oxidation of 2,4,6‐trinitrotoluene in the presence of ozone under
irradiation with UV and visible light”. J. Photochem. Photobiol. A Chem. 231
(2012) 1‐8.
[21] Kanan, S.M.; Lu, Z.; Cox, J.K.; Bernhardt, G.; Tripp, C.P. “Identification of
surface sites on monoclinic WO3 powders by infrared spectroscopy”. Langmuir
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
175
18 (2002) 1707‐1712.
[22] Sun, W.D.; Zhao, Z.B.; Guo, C.; Ye, X.K.; Wu, Y. ʺStudy of the alkylation of
isobutene with n‐butene over WO3/ZrO2 srong solid acid. 1. Effect of the
preparation method, WO3 loading, and calcination temperature”. Ind. Eng.
Chem. Res. 39 (2000) 3717‐3725.
[23] Lee, J.S.; Jang, I.H.; Park, N.G. “Effects of oxidation state and crystallinity of
tungsten oxide interlayer on photovoltaic property in bulk hetero‐junction solar
cell”. J. Phys. Chem. C 116 (2012) 13480‐13487.
[24] Azimirad, R.; Naseri, N.; Akhavan, O.; Moshfegh, A.Z. “Hydrophilicity
variation of WO3 thin films with annealing temperature”. J. Phys. D Appl. Phys.
40 (2007) 1134‐1137.
[25] Beltrán, F.J. “Ozone reaction kinetics for water and wastewater systemsʺ.
Boca Raton, CRC Press, Florida (USA), 2004.
[26] Mena, E.; Rey, A.; Acedo, B.; Beltrán, F.J.; Malato, S. “On ozone‐
photocatalysis synergism in black‐light induced reactions: Oxidizing species
production in photocatalytic ozonation versus heterogeneous photocatalysis”.
Chem. Eng. J. 204‐206 (2012) 131‐140.
[27] Rakowski, S.; Cherneva, D. “Kinetics and mechanism of the reaction of
ozone with aliphatic alcohols”. Int. J. Chem. Kinetics 22 (1990) 321‐329.
[28] Mvula, M.; Von Sonntag, C. “Ozonolysis of phenols in aqueous solution”.
Org. Biomol. Chem. 1 (2003) 1749‐1756.
[29] Leitzke, A.; Von Sonntag, C. “Ozonolysis of unsaturated acids in aqueous
solution: acrylic, methacrylic, maleic, fumaric and muconic acids”. Ozone Sci.
Eng. 31 (2009) 301‐308.
[30] Johnson, P.N.; Davis, R.A. “Diffusivity of ozone in water”. J. Chem. Eng.
Data 41 (1996) 1485‐1487.
[31] Froment, G.F.; Bishoff, K.B. Chemical Reactor Analysis and Design. John
Willey & Sons. New York (USA), 1979.
[32] Huber, M.M.; Gobel, A.; Joss, A.; Hermann, N.; Loffler, D.; McArdell, C.;
Ried, A.; Siegrist, H.; Ternes, T.A.; Von Gunten, U. “Oxidation of
pharmaceuticals during ozonation of municipal wastewaters effluents: A pilot
study”. Environ. Sci. Technol. 39 (2005) 4290‐4299.
[33] Buffle, M.; von Gunten, U. “Phenols and amine induced HO∙ generation
during the initial phase of natural water ozonation”. Environ. Sci. Technol. 40
CAPÍTULO 4 (CHAPTER 4)
176
(2006) 3057‐3063.
[34] Sein, M.M.; Zedda, M.; Tuerk, J.; Schmidt, T.C.; Golloch, A., Von Sonntag, C.
“Oxidation of diclofenac with ozone in aqueous solution”. Environ. Sci. Technol.
42 (2008) 6656‐6662.
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
177
SUPPLEMENTARY INFORMATION OF CHAPTER 4
Figure 4.S1. High resolution XPS spectra of W4f spectral region for all the catalysts.
40 39 38 37 36 35 34 33 32
W 4f5/2
37.6
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W450
-30 W 4f7/2
35.5
2.1 eV
40 39 38 37 36 35 34 33 32
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W450
-5
W 4f5/2
37.6
W 4f7/2
35.5
2.1 eV
40 39 38 37 36 35 34 33 32
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W300
-5
W 4f5/2
37.5
W 4f7/2
35.4
2.1 eV
40 39 38 37 36 35 34 33 32
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W300
-3
W 4f5/2
37.6
W 4f7/2
35.5
2.1 eV
40 39 38 37 36 35 34 33 32
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W300
-1
W 4f5/2
37.4
W 4f7/2
35.3
2.1 eV
40 39 38 37 36 35 34 33 32
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W-0
W 4f5/2
37.4
W 4f7/2
35.2
2.2 eV
CAPÍTULO 4 (CHAPTER 4)
178
Figure 4.S2. High resolution XPS spectra of O1s spectral region for all the catalysts.
536 534 532 530 528 526
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W450
-30 O 1s530.2
536 534 532 530 528 526
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W450
-5 O 1s530.2
536 534 532 530 528 526
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W300
-5 O 1s530.2
536 534 532 530 528 526
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W300
-3 O 1s530.2
536 534 532 530 528 526
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W300
-1
H2O
O 1s530.2
536 534 532 530 528 526
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W-0
H2O
O 1s530.2
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
179
Figure 4.S3. Optical band gap energy determination according to Tauc’s expression.
2.0 2.2 2.4 2.6 2.8 3.00
1
2
3
4
5
(h)
2 (
eV
)2
h(eV)
2.67 eV
W450
-30
2.0 2.2 2.4 2.6 2.8 3.00
1
2
3
4
5
(h)
2 (
eV
)2
h (eV)
2.65 eV
W450
-5
2.0 2.2 2.4 2.6 2.8 3.00
1
2
3
4
5
(h)
2 (
eV)2
h (eV)
2.65 eV
W300
-5
2.0 2.2 2.4 2.6 2.8 3.00
1
2
3
4
5
(h)
2 (
eV)2
h (eV)
2.64 eV
W300
-3
2.0 2.2 2.4 2.6 2.8 3.00
1
2
3
4
5
(h)
2 (eV
)2
h (eV)
2.43 eV
W300
-1
2.0 2.2 2.4 2.6 2.8 3.00
1
2
3
4
5
(h)
2 (eV
)2
h (eV)
2.42 eV
W-0
CAPÍTULO 4 (CHAPTER 4)
180
200 300 400 500 600
0.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7S
IGN
AL
(u.a
.)
W AVELENGTH (nm )
Figure 4.S4. UV‐Vis spectrum of IBP.
Table 4.S1. Rate constants of direct O3‐ECs reactions (pH = 7).
EMERGING
CONTAMINANT ABBREVIATION
kO3
(M‐1 s‐1) Reference
Acetaminophen AAP 4.1x106 [1S]
Metoprolol MTP 1.4x103 [2S]
Caffeine CAF 6.5x102 [3S]
Hydrochlorothiazide HCT 5.1x103 [4S]
Antipyrine ANT 2.5x104 [5S]
Sulfamethoxazole SFM 4.2x105 [6S]
Carbamazepine CAR 3.0x105 [7S]
Ketorolac KET 3.4x105 [5S]
Diclofenac DCF 6.8x105 [8S]
Ibuprofen IBP 9.1 [9S]
PAPER 2: Influence of structural properties on the
activity of WO3 catalysts for visible light ozonation
181
References of Supplementary Information
[1S] Andreozzi, R.; Caprio, V.; Marotta, R.; Vogna, D.
“Paracetamol oxidation from aqueous solutions by means of ozonation and
H2O2/UV system”. Water Res. 37 (2003) 993‐1004.
[2S] Benitez, F.J.; Acero, J.L.; Real, F.J.; Roldán, G. “Ozonation of pharmaceutical
compounds: Rate constants and elimination in various water matrices”.
Chemosphere 77 (2009) 53‐59.
[3S] Broséus, R.; Vincent, S.; Aboulfadl, K.; Daneshvar, A.; Sauvé, S.; Barbeau, B.;
Prévost, M. “Ozone oxidation of pharmaceuticals, endocrine disruptors and
pesticides during drinking water treatment”. Water Res. 43 (2009) 4707‐4717.
[4S] Real, F.J.; Acero, J.L.; Benítez, F.J.; Roldán, G.; Fernández, L.C. “Oxidation of
hydrochlorothiazide by UV radiation, hydroxyl radicals and ozone: Kinetics and
elimination from water systems”. Chem. Eng. J. 160 (2010) 72‐78.
[5S] Rivas, F.J.; Sagasti, J.; Encinas, A.; Gimeno, O. “Contaminants abatement by
ozone in secondary effluents. Evaluation of second‐order rate constants”. J.
Chem. Technol. Biotechnol. 86 (2011) 1058‐1066.
[6S] Beltrán, F.J.; Aguinaco, A.; García‐Araya, J.F. “Mechanism and kinetics of
sulfamethoxazole photocatalytic ozonation in water”. Water Res. 43 (2009) 1359‐
1369.
[7S] Huber, M.M.; Canonica, S.; Park, G.Y.; Von Gunten, U. “Oxidation of
pharmaceuticals during ozonation and advanced oxidation processes”. Environ.
Sci. Technol. 37 (2003) 1016‐1024.
[8S] Sein, M.M.; Zedda, M.; Tuerk, J.; Schmidt, T.C.; Golloch, A.; von Sonntag, C.
“Oxidation of diclofenac with ozone in aqueous solution”. Environ. Sci. Technol.
42 (2008) 6656‐6662.
[9S] Huber, M.M.; Gobel, A.; Joss, A.; Hermann, N.; Loffler, D.; McArdell, C.;
Ried, A.; Siegrist, H.; Ternes, T.A.; Von Gunten, U. “Oxidation of
pharmaceuticals during ozonation of municipal wastewater effluents: A pilot
study”. Environ. Sci. Technol. 39 (2005) 4290‐4299.
CAPÍTULO 5 (CHAPTER 5) PAPER 3: Visible light photocatalytic ozonation of DEET in the presence of different forms of WO3
E. Mena, A. Rey, S. Contreras, F.J. Beltrán
Catalysis Today 252 (2015) 100-106
ABSTRACT. This work deals with the use of WO3 materials for the removal of DEET, an
emerging contaminant, through photocatalytic ozonation using visible light as radiation
source. A commercial WO3 and own-made WO3 catalysts with different forms and
crystalline structure were synthetized by sol-gel and hydrothermal methods. After catalyst characterization with different techniques, photocatalytic ozonation under visible light radiation was found an efficient process to remove DEET. Monoclinic and/or
orthorhombic structure of WO3 seems to favor the catalytic activity more than a
developed surface area together with the presence of W reduced states in the presence of ozone. The best catalysts lead to a complete removal of DEET in less than 20 min with mineralization up to 70 % in 2 h.
Keywords: Photocatalytic ozonation, tungsten trioxide, DEET, ozone, visible light.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
185
5.1. INTRODUCTION
N,N‐diethyl‐meta‐toluamide (DEET) has been widely used as the active
compound in insect repellents for protection against insect bites at 4 ‐ 100 %
concentration in several formulations (lotions, gels, aerosols and sticks) [1].
Thus, DEET contamination has been detected in groundwater, rivers, seawater,
wastewater treatment plants effluents and even in drinking water treated by
conventional systems. This highlights its persistence and degree of recalcitrance
to conventional treatments [2,3]. Toxic effects of DEET to aquatic organisms are
relatively low (e.g. LC50 is in the range 75 ‐ 230 mg L‐1 for Daphnia magna and
Gambusia affinis) [3]. However, it has been reported to have potential
carcinogenic properties in human nasal mucosal cells [4], and additional
research is necessary to examine its fate, transport and ecotoxicity.
Different technologies have been applied for DEET removal such as
ozonation [4]; chlorination and bromination [5]; or AOPs (advanced oxidation
processes) such as heterogeneous photocatalysis with TiO2 [1,3,6]; Fenton and
photo‐Fenton [6,7]; UV/H2O2, O3/H2O2 and UV/O3/H2O2 [7]. The best
performance was obtained using photocatalytic treatments or O3/H2O2 with and
without UVC radiation.
Photocatalytic ozonation, i.e. the combination of ozone and heterogeneous
photocatalysis, has proved to be an efficient AOP improving the efficiency of the
single processes [8,9], and has been also successfully applied for DEET removal
using TiO2 as photocatalyst and UVC as radiation source [6]. However, the use
of UV lamps makes the process expensive for commercial applications so as the
use of solar light is a need for the practical deployment of photocatalytic
technologies. Also, TiO2 presents some limitations since it only uses around 5 %
of solar radiation (UV range) in spite of being the archetypical photocatalyst due
to its relatively high efficiency, low cost and availability [10].
As an attractive alternative to TiO2 for the photocatalytic ozonation process
CAPÍTULO 5 (CHAPTER 5)
186
under solar radiation, WO3 is a visible‐light‐responsive semiconductor of
relatively low cost, with no toxicity and availability, which has demonstrated its
catalytic activity in the degradation and mineralization of phenol through this
process [11]. Up to now, only monoclinic WO3 has been used as catalyst in
photocatalytic ozonation. Thus, this work is focused on the use of different
forms of WO3 as photocatalysts for the photocatalytic ozonation of DEET using
visible light radiation.
5.2. EXPERIMENTAL SECTION
5.2.1. Catalysts preparation
Ten different catalysts were tested, a WO3 commercial catalyst and nine own‐
made WO3 catalysts prepared by sol‐gel and hydrothermal methods. Sol‐gel
procedure was used to prepare WO3 microspheres as described elsewhere [12].
The method is based in the preparation of a CaWO4 precursor with CaCl2 and
Na2WO4 in the presence of citric acid at pH = 12. Once the precursor was
obtained, it was collected by centrifugation, washed with ultrapure water and
ethanol, and dried in air. Then, this precursor was soaked in HNO3 for 24 h at
room temperature followed by calcination at 500, 600 or 700 °C for 2 h. WO3
flocky‐microspheres were prepared by hydrothermal synthesis following a
modified procedure described elsewhere [13]. Typically, Na2WO4 was dissolved
in ultrapure water and acidified (pH = 1 ‐ 1.2) with HCl solution. After stirring
for 240 min, the system was mixed with ethanol 40 % (v/v) and then transferred
to a 125 mL autoclave heated at 120 °C for 48 h. Finally, the precipitates were
separated by centrifugation, washed with water and ethanol, and dried at 100 °C
for 2 h. Then this sample was calcined at 500, 600 or 700 °C for 2 h. Finally, all
fractions were treated at the reaction conditions (0.25 g L‐1 catalysts, 0.5 L
ultrapure water, 10 mg L‐1 O3 with 15 L h‐1 flow rate under visible radiation).
Catalysts nomenclature and synthesis conditions are summarized in Table 5.1.
PAPER 3: Visible light photocatalytic ozonation of DEET in the presence of different forms of WO3
187
Tabl
e 5.
1. P
rope
rtie
s of
WO
3 cat
alys
ts.
CA
TALY
ST
Synt
hesi
s W
O3
crys
talli
ne
phas
e L
(nm
) (I
808/I
716)R
AM
AN
SB
ET
(m2 g
-1)
Na/
WXP
S
(at/a
t) C
a/W
XPS
(at/a
t) C
/WXP
S
(at/a
t)
W-c
C
omm
erci
al
Mon
oclin
ic
28.2
7 1.
97
9.8
---
---
0.42
W1-
500
Sol-g
el
T =
500
°C
Ort
horh
ombi
c 21
.99
2.19
9.
0 0.
16
0.39
0.
36
W1-
600
Sol-g
el
T =
600
°C
Mon
oclin
ic
20.8
2 1.
98
8.1
0.10
0.
42
0.27
W1-
700
Sol-g
el
T =
700
°C
Mon
oclin
ic
32.9
8 1.
96
5.3
0.08
0.
35
0.35
W2
Hyd
roth
erm
al
Hex
agon
al
13.1
6 ---
82
.0
0.38
---
0.
12
W2-
t H
ydro
ther
mal
Tr
eate
d (O
3+hν
) H
exag
onal
10
.99
---
93.7
0.
22
---
0.06
W2-
500-
t H
ydro
ther
mal
T
= 50
0 °C
Tr
eate
d (O
3+hν
) M
onoc
linic
28
.27
2.01
8.
2 0.
17
---
0.10
W2-
600
Hyd
roth
erm
al
T =
600
°C
Mon
oclin
ic
35.9
6 1.
99
5.8
0.34
---
0.
15
W2-
600-
t H
ydro
ther
mal
T
= 60
0 °C
Tr
eate
d (O
3+hν
) M
onoc
linic
35
.96
1.98
9.
6 0.
13
---
0.09
W2-
700-
t H
ydro
ther
mal
T
= 70
0 °C
Tr
eate
d (O
3+hν
) M
onoc
linic
32
.97
1.95
9.
4 0.
21
---
0.06
CAPÍTULO 5 (CHAPTER 5)
188
5.2.2. Characterization
X‐ray diffraction (XRD) patterns were recorded using a powder Bruker D8
Advance XRD diffractometer with a Cu Kα radiation (λ = 0.1541 nm). The data
were collected from 2θ = 10° ‐ 70° at a scan rate of 0.02 s−1. Raman spectra were
obtained using a Nicolet Almega XR Dispersive micro‐Raman (Thermo
Scientific) with a spectral resolution of 2 cm‐1. The samples were excited with a
633 nm laser. BET surface area was determined from nitrogen adsorption‐
desorption isotherms obtained at ‐196 °C using an Autosorb 1 apparatus
(Quantachrome). The samples were outgassed at 250 °C for 12 h under high
vacuum (< 10−4 Pa). A Hitachi S‐48000 scanning electron microscope with 20 ‐ 30
kV accelerating voltage and 500 ‐ 2000 magnification was used to study the
morphology of the samples. X‐ray photoelectron spectra (XPS) were obtained
with a Kα Thermo Scientific apparatus with an Al Kα (h = 1486.68 eV) X‐ray
source using a voltage of 12 kV under vacuum (2x10−7 mbar). Binding energies
were calibrated relative to the C1s peak at 284.6 eV.
5.2.3. Catalytic activity measurements
DEET was used as target compound to test the catalytic activity of the
photocatalysts. The chemical structure and UV‐Vis spectrum of DEET are shown
in Figure 5.S1 of the supplementary information. Photocatalytic experiments
were carried out in semi‐batch mode in a 0.5 L glass‐made spherical reactor,
provided with a gas inlet, a gas outlet and a liquid sampling port. The reactor
was placed in a solar simulator (Suntest CPS, Atlas) provided with a 1500 W Xe
lamp with emission restricted to visible light (λ > 390 nm) using different cut‐off
filters. The irradiation intensity was 550 W m−2, incident photon flux in the
reactor was 3.51x10‐4 einstein min‐1 (actinometrical measurements), and the
temperature maintained between 20 ‐ 40 °C. If required, an ozone generator
(Anseros Ozomat Com AD‐02) was used to produce a gaseous ozone‐oxygen
stream that was fed to the reactor.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
189
In a typical photocatalytic ozonation experiment, the reactor was loaded with
0.5 L of an aqueous solution containing 5 mg L‐1 of DEET (pH0 = 6). Then, 0.125 g
of catalyst were added and the suspension was stirred in the darkness for 30
min. After that, the lamp was switched on and a mixture of ozone‐oxygen (10
mg L‐1 ozone concentration) was fed to the reactor at 15 L h−1 flow rate. The
irradiation time for each experiment was 2 h. Samples were withdrawn from the
reactor and filtered through a 0.2 μm PET membrane except for dissolved ozone
analysis. Experiments of adsorption, photolysis, photocatalytic oxidation,
ozonation and catalytic ozonation were also carried out for comparative reasons.
DEET concentration was analyzed by HPLC‐DAD (Hitachi, Elite LaChrom)
using a Phenomenex C‐18 column (5 μm, 150 mm long, 3 mm diameter) and 0.6
mL min−1 of acetonitrile‐acidified water (0.1 % formic acid) as mobile phase (30 ‐
70 v/v, isocratic). Identification and quantification was carried out at 220 nm.
Total organic carbon (TOC) was measured using a Shimadzu TOC‐VSCH
analyzer. Aqueous ozone was photometrically determined by the indigo method
at 600 nm [14]. Hydrogen peroxide concentration was analyzed photometrically
by the cobalt/bicarbonate method, at 260 nm [15]. Photometric analyses were
performed in a UV‐Vis spectrophotometer Evolution 201 (Thermospectronic).
Ozone in the gas phase was continuously monitored by means of an Anseros
Ozomat GM‐6000Pro analyzer. Short‐chain organic acids were analyzed by ion
chromatography with chemical suppression (Metrohm 881 Compact Pro) with a
conductivity detector using a MetroSep A sup 5 column (250 mm long, 4 mm
diameter) at 45 °C and 0.7 mL min‐1 of Na2CO3 from 0.6 ‐ 14.6 mM in 50 min (10
min post‐time for equilibration) as mobile phase.
5.3. RESULTS AND DISCUSSION
5.3.1. Photocatalysts characterization
Figure 5.1(A) shows the XRD patterns of the WO3 photocatalysts. All the
diffraction peaks of commercial WO3, W‐c, are indexed to monoclinic phase.
CAPÍTULO 5 (CHAPTER 5)
190
W1‐500 catalyst presented mainly WO3 orthorhombic structure but this phase
was transformed to monoclinic at calcination temperatures of 600 and 700 °C
(W1‐600 and W1‐700). Monoclinic and orthorhombic phases differ only in the
extent of the W atoms displacement and can transform reversibly into each other
[16]. Also, the two peaks located at 18.5° and 28.7° are indicative of the
tetragonal phase of CaWO4 [12]. On the other hand, W2 synthetized by
hydrothermal treatment was crystalized as hexagonal WO3 structure. After
being washed at the reaction conditions, no modifications were observed in the
XRD patterns of non‐calcined samples (Figure 5.S2). However, calcination
produced the transformation into monoclinic WO3 as can be observed in Figure
5.1 for treated samples. In addition, Na phases as Na2W4O13 are detected in these
calcined catalysts with main diffraction peaks at 10.8°, 21.8° and 32.9° [17],
indicating an unsatisfactory washing procedure. After being treated at reaction
conditions, the relative intensity of these peaks significantly decreased as can be
noticed in Fig. 5.S2 for W2‐600. The crystal size of the WO3 phases was
determined by the Scherrer’s equation (Table 5.1). In general, a growing in the
crystal size is observed as the calcination temperature increased for catalysts
with the same structure. In addition, no significant changes were observed due
to the washing procedure in W2 catalysts.
Raman spectra depicted in Fig. 5.1(B) and Figure 5.S3 (untreated W2
samples) confirmed the XRD results. The bands located at 808 and 716 cm‐1
correspond to O‐W‐O stretching vibrations whereas the peaks at 328 and 270
cm‐1 were identified as O‐W‐O deformation vibrations [16,18]. Since the main
bands of monoclinic and orthorhombic structures are similar, the intensity ratio
of the main Raman peaks allows to distinguish them being around 2.2 for
orthorhombic WO3 and 1.97 for monoclinic [18]. The intensity ratio of the peaks
located at 808 and 716 cm‐1 is presented in Table 5.1 confirming the
orthorhombic structure of W1‐500 catalyst and its transformation into
monoclinic WO3 at higher calcination temperatures. W2 and W2‐t catalysts
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
191
presented a wide multiple band with a maximum at 799 cm‐1 which can be
composed by the main bands of hexagonal O‐W‐O stretching vibrations at 789,
693 and 651 cm‐1 [16]. The calcination of this catalyst gave place to the
transformation of hexagonal WO3 into monoclinic phase according to the main
characteristic bands and their intensity ratio shown in Table 5.1 (W2 calcined
samples). Finally, the washing procedure of the treated W2 samples did not
produce any significant modification in the WO3 structure (Fig. 5.S3).
Figure 5.1. Structural characterization of WO3 catalysts. (A) XRD patterns (M: Monoclinic
WO3, O: Orthorhombic WO3, H: Hexagonal WO3, C: CaWO4, N: Na2W4O13). (B) Raman
spectra.
(A) (B)
200 300 400 500 600 700 800 900
W2-700-t
INT
EN
SIT
Y (
a.u.
)
RAMAN SHIFT (cm-1)
W2-t
W2-500-t
W2-600-t
710
799
803
805
808
717273328
W-c
W1-500
W1-600
W1-700
807
716
712
330275
269 324
10 20 30 40 50 60 70
W2-700-t
W2-500-t
W1-700
W1-600
W2-600-t
W2-t
W1-500
INT
EN
SIT
Y (
a.u.
)
2
W-c
M
M M
M
M M M
C C O
H H H
H H H
N N N
CAPÍTULO 5 (CHAPTER 5)
192
The morphology of the samples was evaluated by SEM (Figure 5.2).
Commercial WO3 presented an irregular morphology whereas the formation of
spherical particles was confirmed in sol‐gel synthetized catalyst W1‐500 with a
heterogeneous distribution in the range 1 ‐ 5 μm. An increase in the calcination
temperature led to the progressive distortion‐breakup of the microspheres as
confirmed for W1‐600 and W1‐700 in Fig. 5.2(C) and Fig. 5.2(D), respectively. On
the other hand, hydrothermal synthetized catalyst presented mainly irregular
morphology with some isolated spherical particles (Fig. 5.2(E)). An image
magnification in Fig. 5.2(F) revealed a dense packing of small WO3 needles
formed on the surface of WO3 particles [13]. Calcination of this sample gave
place to a more irregular morphology with an enlargement of the needles
transforming them into small WO3 particles (Fig. 5.2(G) and Fig. 5.2(H)). Finally,
the washing treatment did not apparently change the morphology of the W2
type samples.
Specific surface areas (SBET) were calculated from N2 adsorption‐desorption
isotherms and are summarized in Table 5.1. The W1 series presented low SBET
values that decreased with the increasing calcination temperature. For similar
crystalline structures this is in agreement with the growing of the crystal size.
These values are quite similar to that of commercial W‐c. On the other hand, the
hydrothermal treatment gave place to catalysts with higher SBET (82.0 m2 g‐1) due
to the morphology and the lower crystal size of W2 samples. However,
calcination provoked not only a crystalline phase transformation but also a very
sharp decrease of the SBET. In addition, it can be noticed that the washing
procedure in W2 samples led to an increase of the SBET to some extent. Taking
into account that the crystal size is not modified, this effect can be related to the
elimination of some impurities blocking the surface of the samples.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
193
Figure 5.2. SEM images of WO3 catalysts.
5 um
W-c W1-500
W1-600 W1-700
W2-600 W2-600-t
W2
(A)
(H)(G)
(F)(E)
(D)
(C)
(B)
W2
CAPÍTULO 5 (CHAPTER 5)
194
Surface chemical composition and oxidation state of W in the catalysts were
analyzed by XPS. The full spectra of the catalysts (not shown) presented the
main contributions of W4f and O1s. The presence of Na1s was detected in all the
synthetized catalysts and Ca2p in W1 series. In addition C1s contribution was
detected in all samples. The atomic surface ratios of Na, Ca and C to W are
summarized in Table 5.1. For the W1 series, similar values of Ca/W and C/W are
obtained whereas Na/W was lower for the W1‐700 catalyst. These contributions
are due to the precursors used during the synthesis. The most noticeable results
of W2 series are related to the amount of Na in the catalyst surface. This ratio
highly decreased when the catalyst was submitted to the washing procedure
thus indicating that some Na species were released to water. These results are in
agreement to XRD patterns of treated samples. It can be also observed a
decrease on the surface carbon content after the treatment in all the W2 samples.
On the other hand, W4f high resolution spectral region for W‐c, W1 and W2
series are depicted in Figure 5.3. For commercial W‐c and W2 series, two
different peaks corresponding to W4f7/2 and W4f5/2 appear at 35.9 and 38.0 eV,
respectively, which are characteristic of W6+ [16,19]. No significant differences
were found in W2 and W2 treated catalysts (Figure 5.S4). However, the
broadening of these peaks together with their shift to lower binding energies are
indicative of the presence of reduced W5+ to some extent in W1 catalysts [16,19].
This effect is more pronounced in the W1‐700 catalyst as observed in Fig. 5.3.
The partial reduced state of W1 catalysts could be related to the presence of
citric acid from the synthesis procedure that was not completely eliminated
during the washing step, thus acting as a reducing agent. The presence of W5+ in
W1 catalysts could introduce some differences in the electron transport and
charge recombination during photocatalytic processes.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
195
41 40 39 38 37 36 35 34 33 32
W2-700-t
W2-600-t
W2-500-t
W2-t
W1-700
W1-600
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W-c
W1-500
38.0 35.9
37.635.4
37.2 35.3
W 4f
Figure 5.3. High resolution XPS spectra of W4f spectral region of WO3 catalysts.
5.3.2. Catalytic activity
The effectiveness of the catalysts in the photocatalytic ozonation process was
tested using DEET under visible light radiation. First, the performance of this
combined process respect to the single treatments has been checked with all the
catalysts. Results for W‐c catalyst are shown in Figure 5.4. In general, the same
behavior was observed for all the synthetized catalysts. Results for W1‐600 and
W2‐600‐t as an example can be found in the supplementary information (Figures
CAPÍTULO 5 (CHAPTER 5)
196
5.S5‐5.S8). Fig. 5.4(A) shows the time evolution of DEET dimensionless
concentration upon the different treatments applied. Adsorption and photolysis
showed negligible effect on DEET concentration (lower than 5 % removal). The
absence of DEET depletion due to direct photolysis is consistent with the UV‐Vis
absorption spectrum of this compound (Fig. 5.S1). Photocatalytic oxidation gave
place to a slow degradation of DEET, reaching 22 % depletion after 2 h.
Although WO3 is a visible light responsive semiconductor, in photocatalytic
oxidation it presents a low efficiency as a consequence of the high recombination
rate in the presence of oxygen, an unsuitable oxidant for WO3. This drawback is
overcome in the presence of O3 which can easily react with photoexcited
electrons in the conduction band of WO3 [11]. All the ozone treatments showed
the highest degradation rate of DEET. On the one hand, single ozonation and
catalytic ozonation gave place to similar DEET evolution thus indicating that in
dark conditions W‐c did not present any catalytic effect in ozone decomposition
into more reactive species. On the other hand, DEET degradation rate was
highly increased during photocatalytic ozonation reaching more than 99 %
removal in less than 20 min compared to 60 min necessary for the other O3
treatments. In general, pH0 = 6 evolved to pH ~ 4 during the reaction time in all
the treatments (except for direct photolysis), so that in ozone systems, it is
expected a negligible contribution of indirect reactions due to O3 decomposition
at these pH values [20].
On the other hand, it is known that the oxidation of DEET proceeds through
different steps giving place to intermediate compounds which are eventually
transformed into CO2 and H2O (complete mineralization) [4]. DEET
mineralization has been followed as normalized TOC concentration in Fig.
5.4(B). Adsorption, direct photolysis and photocatalytic oxidation led to
negligible TOC removal according to previous DEET evolution. Regarding the
O3‐treatments, single ozonation led to only 19 % mineralization at 120 min, as a
consequence of the formation of some intermediate compounds that are
refractory to ozone direct reaction (mainly short‐chain organic acids) [20].
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
197
Oxalic, formic and acetic acids were detected at concentrations of 2.1, 0.6 and 2.7
mg L‐1, respectively, at the end of the reaction. The presence of the catalyst
combined with ozone in dark conditions did not improve the previous results
confirming the absence of any positive catalytic effect in the reaction. On the
contrary, photocatalytic ozonation showed the highest mineralization rate,
leading to 60 % mineralization in 120 min reaction time. These results also point
out the synergism produced between ozone and visible light irradiated WO3
since final TOC removal of the combined process is much higher than the sum
of these values for the individual processes (60 % in O3/WO3/Vis vs. 4 % in
O2/WO3/Vis and 19 % in O3), also confirming the ability of O3 to capture the
electrons on the WO3 surface compared to O2 [11], and the formation of higher
concentration of oxidizing species such as •HO radicals as occurred in TiO2 [8,9].
Dissolved ozone (Fig. 5.4(C)) and hydrogen peroxide concentration (Fig.
5.4(D)) during the treatments confirmed these results. H2O2 is commonly formed
through direct ozone reactions [21], and was experimentally confirmed in this
work for DEET. The profiles obtained for dissolved O3 and H2O2 during the
photocatalytic ozonation showed a faster decomposition of both compounds
with respect to ozonation and catalytic ozonation. This is indicative of their
consumption on the irradiated WO3 surface through photocatalytic reactions,
probably acting as electron acceptors and thus enhancing the production of
oxidizing species as also occurs in TiO2 [8]. The mechanism and kinetics of the
reaction will be the subject of further work.
CAPÍTULO 5 (CHAPTER 5)
198
Figure 5.4. (A
) DEET dim
ensionless concentration, (B) TOC dim
ensionless concentration, (C) dissolved
O3 concentration and (D) H
2O2 concentration evolution during all the treatm
ents applied w
ith W‐c
catalyst. E
xperim
ental conditions: pH
0 = 6, T = 20 ‐ 40 °C, C
DEET,0 = 5 m
g L‐1, CWO3 = 0.25 g L‐1, C
O3ginlet = 10
mg L‐1, Q
g = 15 L h‐1 (O
2 or O
3/O
2).
(A)
(B)
(C)
(D)
-20
020
40
6080
100
120
0.0
1.0x
10-5
2.0x
10-5
3.0x
10-5
CO3 (M)
Dar
k
TIM
E (
min
)
-20
02
040
6080
100
120
CH2O2 (M)
Dar
k
TIM
E (
min
)
0.0
1.0x
10-5
2.0x
10-5
3.0x
10-5
4.0x
10-5
TOC/TOC0
Dar
k
0.0
0.2
0.4
0.6
0.8
1.0
0.0
0.2
0.4
0.6
0.8
1.0
Dar
k
Ad
sorp
tion
Pho
toly
sis
Pho
toca
taly
tic o
xid
atio
n O
zon
atio
n C
ata
lytic
ozo
nat
ion
Pho
toca
taly
tic o
zona
tion
CDEET/CDEET,0
CTOC/CTOC,0
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
199
The comparison of the activity and efficiency of the different catalysts during
photocatalytic ozonation is depicted in Figure 5.5. For the W2 series, only
treated catalysts have been selected for this study since the catalysts tested
before washing presented a poor activity as can be observed in Figure 5.S9 of the
supplementary material. This could be due to a negative effect of the precursors
that still remained in the catalysts surface and were reduced after the treatment
(Na and C according to XRD and XPS results).
Regarding the evolution of DEET in Fig. 5.5(A) all the catalysts gave place to
higher DEET depletion rate compared with single ozonation. The highest DEET
degradation rate was observed for W1‐600 and W1‐700 catalysts, reaching a
complete removal in 10 min.
Main differences in the activity of the catalysts were found in terms of
mineralization as shown in Fig. 5.5(B). All the catalysts gave place to higher
mineralization degree than ozone alone. Commercial W‐c catalyst led to a 60 %
mineralization at 120 min reaction time similarly to W1‐500, showing an
intermediate activity compared to the synthetized catalysts. Both catalysts
presented similar textural properties but different crystalline structure although
orthorhombic WO3 is a distortion of the monoclinic structure, being both very
similar [16,22]. On the other hand, W1‐600 and W1‐700 catalysts, with
monoclinic structure, were the most active ones but TOC profiles were very
different. W1‐600 catalyst led to a progressive mineralization reaching a 70 % in
2 h. On the contrary, with W1‐700 catalyst, it can be noticed a first step in which
mineralization is slow up to 45 min and, after this time, the mineralization rate
increased also achieving around 70 % at 120 min. This behavior can be due to
the most reduced state of W1‐700 catalyst. The highest catalytic activity could be
related to the presence of reduced W5+ states according to XPS results that can
favor the electron transport in the WO3 structure [19]. These results are contrary
to those previously found for photocatalytic oxidation where a more oxidized
material had a higher catalytic activity in the presence of oxygen and reported
that W5+/W4+ states could act as recombination centers [16]. However, the
CAPÍTULO 5 (CHAPTER 5)
200
presence of O3 could take the advantage of avoiding the recombination to some
extent and thus benefits from the higher electron mobility. For W2 series,
mineralization rate of W2‐t catalyst was noticeably slower regardless of its high
SBET, indicating that the hexagonal structure of this catalyst could not be
appropriated for the process. This is evidenced by the performance of calcined
catalysts (with monoclinic structure and very low SBET) similar or even better
than that of hexagonal W2‐t. It can be noticed the highest catalytic activity of
W2‐600‐t catalyst in terms of mineralization compared to W2‐500‐t and W2‐700‐t
catalysts whose main difference is the higher crystal size of monoclinic phase in
W2‐600‐t.
Finally, the best performance of the W1 series can be observed in the
evolution of short‐chain organic acids formed during the process. Fig. 5.5(C)
represents the sum of the concentrations of main organic acids detected (oxalic,
acetic and formic) in terms of mg C L‐1. The degradation of these intermediates
is only observed for W1 series catalyst.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
201
-20 0 20 40 600.0
0.2
0.4
0.6
0.8
1.0 W-c W1-500 W1-600 W1-700 W2-t W2-500-t W2-600-t W2-700-tDark
CD
EE
T/C
DE
ET
,0
TIME (min)
Figure 5.5. (A) DEET dimesionless concentration, (B) TOC dimensionless concentration
and (C) sum of the TOC corresponding to short‐chain organic acids (ΣCTOC,a) evolution
during photocatalytic ozonation with the best WO3 catalysts (dotted lines show single
ozonation results). Experimental conditions: pH0 = 6, T = 20 ‐ 40 °C, CDEET,0 = 5 mg L‐1,
CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 15 L h‐1 (O3/O2).
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Dark
CT
OC
/CT
OC
,0
TIME (min)
(A)
(B)
0 20 40 60 80 100 1200.0
0.4
0.8
1.2
1.6
2.0
CT
OC
,a (
mg
L-1)
TIME (min)
(C)
CAPÍTULO 5 (CHAPTER 5)
202
5.4. CONCLUSIONS
WO3 catalysts were synthetized with different morphology (spherical and
irregular) and structure (orthorhombic, monoclinic and hexagonal) through sol‐
gel and hydrothermal procedures. An increase in the calcination temperature
produced the transformation of orthorhombic or hexagonal into monoclinic
structure leading to higher crystal sizes and a less developed surface area. The
presence of citric acid in the sol‐gel synthetized catalysts led to the presence of
W reduced states to some extent. The catalysts were active in the photocatalytic
ozonation of DEET with visible light radiation. A synergistic effect between
ozone and visible light irradiated WO3 was produced leading to a higher
efficiency in the combined process. Monoclinic and/or orthorhombic structure of
WO3 favored the catalytic activity with respect to hexagonal structure, more
than the developed surface area. Also the presence of W reduced states seems to
be beneficial for the process in the presence of ozone. The best catalysts gave
place to a complete removal of DEET in less than 20 min with mineralization up
to 70 % in 2 h.
AKNOWLEDGEMENTS
Authors thank the Spanish MINECO and European Feder Funds (CTQ2012‐
35789‐C02‐01) for economic support. E. Mena thanks the Consejería de Empleo,
Empresa e Innovación (Gobierno de Extremadura) and European Social Fund
for her predoctoral FPI grant (Ref. PD12059).
REFERENCES
[1] Antonopoulou, M.; Giannakas, A.; Deligiannakis, Y.; Konstantinou, I.
“Kinetic and mechanistic investigation of photocatalytic degradation of the N,N‐
diethyl‐m‐toluamide”. Chem. Eng. J. 231 (2013) 314‐325.
[2] Costanzo, S.D.; Watkinson, A.J.; Murby, E.J.; Kolpin, D.W.; Sandstrom, M.W.
“Is there a risk associated with the insect repellent DEET (N,N‐diethyl‐m‐
toluamide) commonly found in aquatic environments?”. Sci. Total Environ. 384
(2007) 214‐220.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
203
[3] Adams, W.A.; Impellitteri, C.A. “The photocatalysis of N,N‐diethyl‐m‐
toluamide (DEET) using dispersions of Degussa P‐25 TiO2 particles”. J.
Photochem. Photobiol. A Chem. 202 (2009) 28‐32.
[4] Benítez, F.J.; Acero, J.L.; García‐Reyes, J.F.; Real, F.J.; Roldán, G.; Rodríguez,
E.; Molina‐Díaz, A. “Determination of the reaction rate constants and
decomposition mechanisms of ozone with two model emerging contaminants;
DEET and Nortriptyline”. Ind. Eng. Chem. Res. 52 (2013) 17054‐17073.
[5] Acero, J.L.; Benítez, F.J.; Real, F.J.; Roldán, G.; Rodríguez, E. “Chlorination
and bromination kinetics of emerging contaminants in aqueous systems”. Chem.
Eng. J. 219 (2013) 43‐50.
[6] Benítez, F.J.; Acero, J.L.; Real, F.J.; Roldán, G.; Rodríguez, E. “The
effectiveness of single oxidants and AOPs in the degradation of emerging
contaminants in waters: A comparison study”. Ozone Sci. Eng. 35 (2013), 263‐
272.
[7] Li, W.; Nanaboina, V.; Zhou, Q.; Korshin, G.V. “Effects of Fenton treatment
on the properties of effluent organic matter and their relationships with the
degradation of pharmaceuticals and personal care products”. Water Res. 46
(2012) 403‐412.
[8] Mena, E.; Rey, A.; Acedo, B.; Beltrán, F.J.; Malato, S. “On ozone‐
photocatalysis synergism in black‐light induced reactions: Oxidizing species
production in photocatalytic ozonation versus heterogeneous photocatalysis”.
Chem. Eng. J. 204‐206 (2012) 131‐140.
[9] Agustina, T.E.; Ang, H.M.; Vareek, V.K. “A review of synergistic effect of
photocatalysis and ozonation on wastewater treatment”. J. Photochem.
Photobiol. C Photochem. Rev. 6 (2005) 264‐273.
[10] Hernández‐Alonso, M.D., Fresno, F., Suárez, S., Coronado, J.M.
“Development of alternative photocatalysts to TiO2: Challenges and
opportunities”. Energ. Environ. Sci. 2 (2009) 1231‐1257.
[11] Nishimoto, S., Mano, T., Kameshima, Y., Miyake, M. “Photocatalytic water
treatment over WO3 under visible light irradiation combined with ozonation”.
Chem. Phys. Letters 500 (2010) 86‐89.
[12] Zhang, L.; Tang, X.; Lu, Z.; Wang, Z.; Li, L.; Xiao, Y. “Facile synthesis and
photocatalytic activity of hierarchical WO3 core‐shell microspheres”. Appl. Surf.
Sci. 258 (2011) 1719‐1724.
[13] Shen, Y.; Ding, D.; Deng, Y. “Fabrication and characterization of WO3 flocky
microspheres induced by ethanol”. Powder Technol. 211 (2011) 114‐119.
CAPÍTULO 5 (CHAPTER 5)
204
[14] Bader, H.; Hoigné, J. “Determination of ozone in water by the indigo
method”. Water Res. 15 (1981) 449‐456.
[15] Masschelein, W.; Denis, M.; Ledent, R. “Spectrophotometric determination
of residual hydrogen peroxide”. Water & Sewage Works (1977) 69‐72.
[16] Szilágyi, I.M.; Fórizs, B.; Rosseler, O.; Szegedi, A.; Németh, P.; Király, P.;
Tárkányi, G.; Vajna, B.; Varga‐Josepovits, K.; László, K.; Tóth, A.L.; Baranyai, P.;
Leskelä, M. “WO3 photocatalysts: Influence of structure and composition”. J.
Catal. 294 (2012) 119‐127.
[17] Sheng, T.; Chavvakula, P.; Cao, B.; Yue, N.; Zhang, Y.; Zhang, H. “Growth
of ultra‐long sodium tungsten oxide and tungsten oxide nanowires: Effects of
impurity and residue deposition”. J. Crystal Growth 395 (2014) 61‐67.
[18] Kustova, G.N.; Chesalov, Y.A.; Plyasova, L.M.; Molina, I.Y.; Nizovskii, A.I.
“Vibrational spectra of WO3∙nH2O and WO3 polymorphs”. Vibrational Spect. 55
(2011) 235‐240.
[19] Lee, J.S.; Jang, I.H.; Park, N.G. “Effects of oxidation state and crystallinity of
tungsten oxide interlayer on photovoltaic property in bulk hetero‐junction solar
cell”. J. Phys. Chem. C 116 (2012) 13480‐13487.
[20] F.J. Beltrán. Ozone reaction kinetics for water and wastewater systems. Boca
Raton, CRC Press, 2004, Florida (USA).
[21] Leitzke, A.; Von Sonntag, C. “Ozonolysis of unsaturated acids in aqueous
solution: acrylic, methacrylic, maleic, fumaric and muconic acids”. Ozone Sci.
Eng. 31 (2009) 301‐308.
[22] Pecquenard, B.; Lecacheux, H.; Livage J.; Julien, C. “Orthorhombic WO3
Formed via a Ti‐Stabilized WO3∙13H2O Phase”. J. Solid State Chem. 135 (1998)
159‐168.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
205
200 300 400 500 600
0.0
0.5
1.0
1.5
2.0
2.5
3.0
SIG
NA
L (
u.a
.)
WAVELENGTH (nm)
10 20 30 40 50 60 70
W2-600-t
W2-600
W2-t
W2
INT
EN
SIT
Y (
a.u.
)
2
SUPPLEMENTARY INFORMATION OF CHAPTER 5
Figure 5.S1. UV‐Vis spectrum and chemical structure of DEET.
Figure 5.S2. XRD patterns of non‐treated and treated WO3 catalysts.
CAPÍTULO 5 (CHAPTER 5)
206
.
200 300 400 500 600 700 800 900
INT
EN
SIT
Y (
a.u.
)
RAMAN SHIFT (cm-1)
W2
W2-t
W2-600
W2-600-t
805
273
328
716
799
Figure 5.S3. Raman spectra of non‐treated and treated WO3 catalysts.
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
207
41 40 39 38 37 36 35 34 33 32
W2-600-t
W2
W2-t
INT
EN
SIT
Y (
a.u.
)
BINDING ENERGY (eV)
W2-60038.0 35.9
W 4f
Figure 5.S4. High resolution XPS spectra of W4f spectral region of non‐treated and
treated WO3 catalysts.
CAPÍTULO 5 (CHAPTER 5)
208
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Adsorption Photolysis Photocatalytic oxidation Ozonation Catalytic ozonation Photocatalytic ozonation
Dark
CD
EE
T/C
DE
ET
,0
TIME (min)
Figure 5.S5. (A) DEET and (B) TOC dimensionless concentration evolution during all the
treatments applied with W1‐600 catalyst. Experimental conditions: pH0 = 6, T = 20 ‐ 40 °C,
CDEET,0 = 5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 15 L h‐1 (O2 or O3/O2).
(A)
(B)
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Dark
TO
C/T
OC
0
TIME (min)
0
0
CT
OC
/CT
OC
,0
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
209
0 20 40 60 80 100 1200.0
1.0x10-5
2.0x10-5
3.0x10-5
CO
3 (M
)
TIME (min)
0 20 40 60 80 100 1200.0
1.0x10-5
2.0x10-5
3.0x10-5
4.0x10-5
5.0x10-5
6.0x10-5 Photocatalytic oxidation Ozonation Catalytic ozonation Photocatalytic ozonation
CH
2O2
(M)
TIME (min)
Figure 5.S6. (A) Dissolved O3 concentration and (B) H2O2 concentration during all the
treatments applied with W1‐600 catalyst. Experimental conditions: pH0 = 6, T = 20 ‐ 40 °C,
CDEET,0 = 5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 15 L h‐1 (O2 or O3/O2).
(A)
(B)
CAPÍTULO 5 (CHAPTER 5)
210
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Adsorption Photolysis Photocatalytic oxidation Ozonation Catalytic ozonation Photocatalytic ozonationDark
CD
EE
T/C
DE
ET
,0
TIME (min)
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Dark
CT
OC
/CT
OC
,0
TIME (min)
Figure 5.S7. (A) DEET and (B) TOC dimensionless concentration evolution during all the
treatments applied with W2‐600‐t catalyst. Experimental conditions: pH0 = 6, T = 20 ‐ 40
°C, CDEET,0 = 5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 15 L h‐1 (O2 or O3/O2).
(A)
(B)
PAPER 3: Visible light photocatalytic ozonation
of DEET in the presence of different forms of WO3
211
0 20 40 60 80 100 1200.0
1.0x10-5
2.0x10-5
3.0x10-5
CO
3 (M
)
TIME (min)
0 20 40 60 80 100 1200.0
1.0x10-5
2.0x10-5
3.0x10-5
4.0x10-5
5.0x10-5
6.0x10-5 Photocatalytic oxidation Ozonation Catalytic ozonation Photocatalytic ozonation
CH
2O2
(M)
TIME (min)
Figure 5.S8. (A) Dissolved O3 concentration and (B) H2O2 concentration during all the
treatments applied with W2‐600‐t catalyst. Experimental conditions: pH0 = 6, T = 20 ‐ 40
°C, CDEET,0 = 5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 15 L h‐1 (O2 or O3/O2).
(A)
(B)
CAPÍTULO 5 (CHAPTER 5)
212
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Dark
W2 W2-t W2-600 W2-600t
CT
OC
/CT
OC
,0
TIME (min)
Figure 5.S9. TOC dimensionless concentration evolution during photocatalytic ozonation
with untreated and treated W2 catalysts. Experimental conditions: pH0 = 6, T = 20 ‐ 40 °C,
CDEET,0 = 5 mg L‐1, CWO3 = 0.25 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 15 L h‐1 (O3/O2).
CAPÍTULO 6 (CHAPTER 6) PAPER 4: Nanostructured CeO2 as catalysts for different AOPs based in the application of ozone and simulated solar radiation
E. Mena, A. Rey, E.M. Rodríguez, F.J. Beltrán
Catalysis Today 280 (2017) 74-79
ABSTRACT. Two CeO2 catalysts with different morphology (nanocubes and nanorods)
were synthetized by hydrothermal treatment, characterized by TEM, XRD, N2
adsorption-desorption isotherms, XPS and DR-UV-Vis spectroscopy and examined for different AOPs (photocatalytic oxidation, catalytic and photocatalytic ozonation) using simulated solar radiation (λ > 300 nm) or visible radiation (λ > 390 nm) from a Xe lamp, and DEET as target compound. Neither significant DEET removal nor mineralization was achieved by solar photocatalysis and the catalysts did not show an important beneficial effect during catalytic ozonation. The best results were obtained in photocatalytic ozonation process being nanorods more active under visible radiation according to its lower band gap energy whereas nanocubes presented the highest intrinsic photocatalytic
activity under visible and solar radiation. Photocatalytic ozonation of DEET with CeO2
catalysts proceeds through a complex reaction mechanism that involves at least ozone-direct reaction that triggers a radical chain mechanism, indirect ozone reactions, ozone photolytic and photocatalytic decomposition and hydrogen peroxide formation-decomposition.
Keywords: Photocatalytic ozonation, tungsten trioxide, DEET, ozone, visible light.
PAPER 4: Nanostructured CeO2 as catalysts for different
AOPs based in the application of ozone and simulated solar radiation
215
6.1. INTRODUCTION
Water pollution due to specific organic contaminants (phenolic compounds,
aromatic hydrocarbons, pesticides, pharmaceuticals, etc.) is a growing problem
that requires the implementation of advanced technologies capable of removing
not only the initial compounds but also their metabolites. Among the available
technologies, advanced oxidation processes (AOPs) have demonstrated to
efficiently remove a wide variety of organic pollutants due to the generation of
oxidizing species, mainly hydroxyl radicals, which non‐selectively attack the
organic matter [1]. Photocatalytic treatments are promising alternatives due to
the possibility of using natural solar light as a radiation source to minimize
treatment costs. Besides, the combination of photocatalysis with ozone
(photocatalytic ozonation) increases the generation of oxidizing species leading
to higher degrees of mineralization than individual treatments of ozonation or
photocatalysis [2‐4].
For the photocatalytic processes, much of the research effort is focused on the
development of photocatalysts capable of absorbing visible light effectively in
order to increase the fraction of usable solar radiation compared to the
commonly used TiO2 [5,6]. In this sense, cerium oxide CeO2 is a wide band‐gap
semiconducting material usually in the range of 2.9 and 3.2 eV which absorbs
light in the near UV and slightly in the visible region depending on the
morphology and particle size, being its properties as catalyst and catalytic
support highly shape‐size dependent [7‐9]. In addition, CeO2 possesses a
sufficiently long lifetime of photo‐generated electron‐hole pairs to trigger
photocatalytic reactions in liquid and gas phase [10]. Bare CeO2 with different
morphologies/properties (microspheres, nanospheres, nanosheets, etc.) has been
used as photocatalyst in the degradation of organic pollutants in water, most of
them belonging to the dyes family [11‐15] or phenol and phenol derivatives
[8,16]. These works pointed out that the morphology and crystal size of CeO2 are
highly important in the amount and nature of surface defects, oxygen vacancies,
CeO2‐light interaction and, definitely, in the behavior of CeO2 as photocatalyst.
CAPÍTULO 6 (CHAPTER 6)
216
On the other hand, CeO2 has demonstrated to be catalytically active during
ozonation of organic pollutants of different nature such as phenol [17], oxalic
and oxamic acids [18] or bezafibrate [19]. Main reaction pathways proposed are
related to the adsorption and subsequent decomposition of ozone into hydroxyl
radicals, being the presence of Ce(III)/oxygen vacancies in the surface the main
responsible of the catalytic activity [17,18].
As far as we know, the combination of photocatalytic oxidation with ozone
using CeO2 as photocatalyst has not been previously reported but it is expected
to take benefits of both processes (photocatalysis and catalytic ozonation)
together with the synergism observed during photocatalytic ozonation
treatment [2]. Therefore, the aim of this work focuses on the application of two
nanostructured CeO2 catalysts with different morphology (nanocubes and
nanorods) in ozone and photocatalysis based AOPs (photocatalysis, catalytic
and photocatalytic ozonation) using simulated solar radiation (300 ‐ 800 nm) and
visible radiation (390 ‐ 800 nm) from a Xe lamp, and N,N‐diethyl‐meta‐
toluamide (DEET) as target compound. DEET is a common insect repellent
found in different aquatic environments [20] that presents a low reactivity with
ozone [21], thus facilitating the comparison between the different processes.
6.2. EXPERIMENTAL SECTION
6.2.1. Catalysts preparation
Two CeO2 catalysts were synthetized by hydrothermal treatment according
to a previous work [22]. Briefly, Ce(NO3)3∙6H2O was dissolved in 6 M NaOH
solution and then transferred to a 125 mL autoclave (filled at 75 %) and heated
during 24 h at 100 °C for nanorods (NR) or 180 °C for nanocubes (NC). After the
hydrothermal treatment, the autoclave was cooling down to room temperature
and then, the precipitates were separated by centrifugation, washed sequentially
with water and ethanol and dried at 100 °C overnight.
PAPER 4: Nanostructured CeO2 as catalysts for different
AOPs based in the application of ozone and simulated solar radiation
217
6.2.2. Characterization
CeO2 catalysts were characterized by different techniques. X‐ray diffraction
(XRD) patterns were recorded using a powder Bruker D8 Advance XRD
diffractometer with a Cu Kα radiation (λ = 0.1541 nm). The data were collected
from 2θ = 20° ‐ 80° at a scan rate of 0.02 s−1. A Tecnai G2 20 (FEI Company)
transmission electron microscope at 200 kV was used to study the crystallinity
and morphology of the samples. Textural properties were analyzed by nitrogen
adsorption‐desorption isotherms at ‐196 °C using an Autosorb‐1 apparatus
(Quantachrome). The samples were outgassed at 250 °C for 12 h under vacuum
(< 10−2 mbar). Surface characterization was performed by X‐ray photoelectron
spectroscopy. XPS spectra were obtained with a Kα Thermo Scientific apparatus
with Al Kα (h = 1486.68 eV) X‐ray source using a voltage of 12 kV under
vacuum (2x10−7 mbar). Binding energies were calibrated relative to the C1s peak
at 284.6 eV and spectra deconvolution was accomplished using XPS‐Peak 4.1
software. Diffuse reflectance UV‐Vis spectroscopy (DR‐UV‐Vis) measurements
were performed with a UV‐Vis‐NIR Cary‐5000 spectrophotometer (Varian‐
Agilent Technologies) equipped with an integrating sphere device. Band gap
values of the samples were obtained using Tauc’s equation, assuming an
indirect band gap semiconductor behavior [23].
6.2.3. Catalytic activity measurements
Photocatalytic experiments were carried out in a solar simulator (Suntest
CPS, Atlas) provided with a 1500 W Xe lamp operated at 550 W m−2. In some
cases light transmission was restricted to λ > 390 nm by means of a modified
polyester cut‐off filter (Edmund Optics) for visible experiments, whereas a
window glass filter was used for simulating the spectrum of solar radiation
reaching the Earth’s surface (λ > 300 nm). The spectral irradiance of Xe lamp
using the window glass filter, and the transmittance of the polyester filter are
represented in Figure 6.1. The experiments were carried out in semi‐batch mode
(batch with respect to the liquid phase and continuous with respect to the gas
CAPÍTULO 6 (CHAPTER 6)
218
phase) using a borosilicate glass‐made round flask provided with gas inlet, gas
outlet and a liquid sampling port. In a typical photocatalytic ozonation
experiment, the reactor was loaded with 500 mL of 5 mg L‐1 of DEET aqueous
solution (pH0 = 6). The catalyst was then added (0.25 g L‐1) and the suspension
was magnetically stirred in the dark for 30 min before switching on the lamp
and feeding ozone to the reactor (15 L h−1 gas flow rate, 10 mg L−1 O3, with an
ozone generator Labor‐Ozonisator from Sander fed with pure oxygen from a
cylinder). The temperature in all the experiments was maintained at 35 ‐ 40 °C.
Besides photocatalytic ozonation (CeO2/O3/h), blank experiments of photolysis
(h), single ozonation (O3), photolytic ozonation (O3/h), adsorption (CeO2),
photocatalysis (CeO2/h) and catalytic ozonation (CeO2/O3) were carried out for
comparative reasons. Most of the experiments were triplicated. Samples were
withdrawn from the reactor at different interval times, filtered with 0.2 μm PET
membranes and analyzed.
300 400 500 600 700 8000
1
2
3
4
5
SP
EC
TR
AL
IRR
AD
IAN
CE
(W
m-2
nm
-1)
AB
SO
RB
AN
CE
(u
.a.)
WAVELENGTH (nm)
NC
NR
0
20
40
60
80
100F
ILT
ER
TR
AN
SM
ITT
AN
CE
(%
)
Figure 6.1. UV‐Vis characteristics of the photocatalytic system applied (spectral
irradiance of the Xe lamp in the solar simulator, transmittance of the UV filter and UV‐
Vis absorbance of the CeO2 catalysts).
DEET concentration was analyzed by HPLC‐DAD (Hitachi, Elite LaChrom)
PAPER 4: Nanostructured CeO2 as catalysts for different
AOPs based in the application of ozone and simulated solar radiation
219
using a Phenomenex C‐18 column (5 μm, 150 mm long, 3 mm diameter) and 0.6
mL min−1 of acetonitrile‐acidified water (0.1 % formic acid) as mobile phase (30 ‐
70 v/v, isocratic). Identification and quantification was carried out at 220 nm.
Total organic carbon (TOC) was measured using a Shimadzu TOC‐VSCH
analyzer. Aqueous ozone was photometrically determined by the indigo method
at 600 nm [24], in a UV‐Vis spectrophotometer Evolution 201
(Thermospectronic) and ozone in the gas phase was continuously monitored by
an online analyzer (Anseros Ozomat GM‐6000Pro). Hydrogen peroxide
concentration was photometrically determined by the cobalt/bicarbonate
method at 260 nm using the same spectrophotometer [25]. Short‐chain organic
acids were analyzed by ion chromatography with chemical suppression
(Metrohm 881 Compact Pro) and a conductivity detector, using a MetroSep A
sup 5 column (250 mm long, 4 mm diameter) at 45 °C and 0.7 mL min‐1 of
Na2CO3 from 0.6 ‐ 14.6 mM in 50 min (10 min post‐time for equilibration) as
mobile phase.
6.3. RESULTS AND DISCUSSION
6.3.1. Photocatalysts characterization
The formation of nanostructured CeO2 with NC or NR morphology was
confirmed by TEM as can be observed in Figure 6.2(A). From the NC images,
also some rectangular prism particles have been detected (a minority in the
images analyzed). The size of NC and NR, determined from several TEM
representative images (not shown), is given in Table 6.1. A size distribution
between 25 and 100 nm was obtained for NC whereas the NR presented a
thickness of ~ 7.2 nm and lengths between 40 and 200 nm. NC and NR showed
{100} and {110} + {100} exposed facets, respectively, which is in accordance with
previous reports [22]. On the other hand, XRD patterns shown in Fig. 6.2(B)
confirmed the formation of pure CeO2 cubic phase (fluorite structure, JCPDS 34‐
0394) in both NC and NR samples [22].
CAPÍTULO 6 (CHAPTER 6)
220
Figure 6.2. Structural characterization of the CeO2 nanostructured catalysts: (A) TEM
images and (B) XRD patterns.
The porosity of the catalysts was evaluated by means of N2 adsorption‐
desorption isotherms. BET surface area and total pore volumes are summarized
in Table 6.1. Previous reports have shown that these materials are free of a
developed pore structure except general random stacking of nanoparticles
[22,26]. Due to its morphology and lower crystal size, NR presented higher SBET
and pore volume values than NC. With these characteristics, it is reasonable to
assume that the powder dispersion of these materials (NR and NC) can be
illuminated without internal shading from pores. Thus, the surface area exposed
to light (SBET in this case) should be taken into account since it can affect the
photocatalytic behavior to some extent.
NC
NR
(A) (B)
20 30 40 50 60 70 80
INT
EN
SIT
Y (
a.u.
)
2 (º)
NR-CeO2
NC-CeO2
PAPER 4: Nanostructured CeO2 as catalysts for different AOPs based in the application of ozone and simulated solar radiation
221
Tabl
e 6.
1. P
rope
rtie
s of
the
CeO
2 nan
ostr
uctu
red
cata
lyst
s.
CA
TALY
ST
Synt
hesi
s a S
ize
(nm
) Ex
pose
d fa
cets
CeO
2 cr
ysta
lline
ph
ase
SBET
(m
2 g-1
) V
p
(cm
3 g-1
) b C
e(II
I)
(%, a
t.)
b Ce(
IV)
(%, a
t.)
Eg
(eV
)
NC
H
ydro
ther
mal
NaO
H
6 M
180
°C, 2
4h
25 -
100
{100
} C
ubic
14
0.
185
12.6
87
.4
3.32
NR
Hyd
roth
erm
al N
aOH
6
M 1
00 °C
, 24h
(7
.2 ±
1.1
) x
(40
- 200
) {1
10} +
{100
} C
ubic
10
9 0.
522
23.6
76
.4
3.07
a S
ize o
f the
mai
n di
men
sions
by
TEM
ana
lyse
s; b C
alcu
late
d fro
m C
e 3d
XPS
dec
onvo
lutio
n
CAPÍTULO 6 (CHAPTER 6)
222
One of the main characteristics of CeO2 catalysts is their oxygen and electron
mobility due to the presence of surface defects, reduced states of Ce(III) or
oxygen vacancies in the CeO2 structure. High‐resolution XPS spectra
corresponding to Ce3d spectral region of NC and NR samples have been plotted
in Figure 6.3. From this figure, a slight shift of Ce3d peaks to lower binding
energies in NR with respect to NC sample is noticed. This has been attributed to
an increase of the electron density around Ce nucleus, suggesting the existence
of more Ce(III) ions on the surface of NR sample [26]. In addition, this has been
confirmed by the highest intensity of the peaks corresponding to Ce(III) species
(v0, v’, u0, u’) in NR catalyst [18,26,27], leading to a higher percentage of surface
Ce(III) in NR compared to NC sample (see Table 6.1).
920 910 900 890 880 870
NR 23.6% 76.4%NC 12.6% 87.4%
CP
S (
a.u.
)
BINDING ENERGY (eV)
NR
NC
v0
v'
u0u'
Ce(III) Ce(IV)
Figure 6.3. High‐resolution XPS spectra and deconvolution of the Ce3d spectral region of
the CeO2 nanostructured catalysts.
PAPER 4: Nanostructured CeO2 as catalysts for different
AOPs based in the application of ozone and simulated solar radiation
223
Finally, optical properties of the catalysts have been studied by DR‐UV‐Vis
spectroscopy. Fig. 6.1 shows the UV‐Vis absorption spectra of NC and NR where
a noticeable shift to higher wavelength for NR sample is observed, which
suggests this catalyst can use a higher fraction of visible radiation to trigger
photocatalytic reactions. The optical energy band gap values (Eg) calculated
assuming indirect band gap behavior for both catalysts are also summarized in
Table 6.1. A value of 3.32 eV was obtained for NC, somewhat higher than that
reported for bulk CeO2 (around 3.19 eV) due to smaller particle sizes [28]. On the
contrary, a noticeable lower Eg value was calculated for NR indicating a clearly
shift to higher absorption wavelengths as deduced from Fig. 6.1. This effect may
be attributed to the lattice defects and oxygen vacancies present in the ceria NR
[28], in agreement with the results obtained by XPS analyses.
6.3.2. Catalytic activity
The activity of the CeO2 catalysts was evaluated in DEET degradation
through AOPs of photocatalytic oxidation, catalytic ozonation and
photocatalytic ozonation in aqueous solution using visible and simulated solar
radiation. The results obtained for DEET degradation are shown in Figure 6.4.
Some blank experiments are also included for comparative purposes. As
reported in a previous work, at the conditions applied no DEET degradation by
direct photolysis was observed [29]. The adsorption capacity of both catalysts
was also negligible being DEET removal lower than 10 %. In addition, in spite of
the optical properties of the catalysts, no significant DEET removal was achieved
by photocatalysis regardless of the radiation used (visible or solar). These results
differ from those previously reported using different CeO2 catalysts, although in
those cases photosensitizing dyes were used as target compounds [11‐15]. On
the other hand, the highest DEET degradation rate was obtained for ozone‐
based systems. Thus, single ozonation led to a complete removal of DEET in less
than 60 min. A similar DEET evolution was observed during catalytic ozonation
with respect to ozonation regardless of the catalyst used. This points out that
these materials do not show any beneficial effect on DEET removal when
CAPÍTULO 6 (CHAPTER 6)
224
combined with ozone in the dark, at the conditions used in this work.
-30 0 30 60 90 1200.0
0.2
0.4
0.6
0.8
1.0
NR NR/h NR/O
3
NR/O3/h
NC NC/h NC/O
3
NC/O3/h
h O
3
O3/h
Dark
CD
EE
T/C
DE
ET
,0
TIME (min)
-30 0 30 60 90 1200.0
0.2
0.4
0.6
0.8
1.0
NR NR/h NR/O
3
NR/O3/h
NC NC/h NC/O
3
NC/O3/h
h O
3
O3/h
Dark
CD
EE
T/C
DE
ET
,0
TIME (min)
Figure 6.4. Evolution of normalized DEET concentration during different treatments
applied under visible (A) and solar (B) simulated radiation (maximum standard
deviation of the DEET concentrations in 3 repeated runs 0.03, not shown). Experimental
conditions: pH0 = 6, T = 35 ‐ 40 °C, CDEET,0 = 5 mg L‐1, CCeO2 = 0.25 g L‐1, CO3,g inlet = 10
mg L‐1, Qg = 15 L h‐1 (O2 or O3/O2).
However, the impact of the radiation (visible and solar) on the efficacy of
ozone‐based processes was significant and, in general, higher degradation rates
(A) VISIBLE
(B) SOLAR
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225
were observed when solar light was used (Fig. 6.4(B)). In this sense, DEET was
completely removed after 45 and 30 min when ozone was combined with visible
and solar radiation, respectively, being these periods clearly lower than that
required by single ozonation (60 min). It has been previously demonstrated that
ozone can undergo direct photolysis under solar radiation promoting the
formation of some reactive oxygen species capable of oxidizing the organic
matter in aqueous solution [30,31]. Nevertheless, the interaction between visible
radiation and ozone is unclear and will be investigated in future works. Finally,
according to Fig. 6.4(A), the addition of NC and NR catalysts to the ozone +
radiation system had almost no effect on DEET degradation rate. Only a slight
beneficial effect was observed when visible light was used, regardless of the
type of catalyst.
Main differences between ozone‐based treatments were observed in terms of
DEET mineralization (i.e. transformation into CO2, H2O and inorganic
compounds). Figure 6.5 shows the percentage of mineralization during ozone‐
based treatments at 120 min reaction time. As observed, ozone alone gave rise to
only a 19 % mineralization due to the formation of intermediate compounds
refractory to ozone direct reaction (oxalic, formic and acetic acids were detected
in the reaction medium) [32]. The presence of CeO2 catalysts slightly increased
(by 5 %) the final mineralization, which could be attributable to a positive effect
of cerium oxide on the elimination of some intermediate compounds formed
during DEET ozonation. The combination of ozone and radiation gave rise to
29 % and 47 % of TOC removal under visible and solar light, respectively.
Therefore, compared to visible light the effect of the entire solar spectrum
(which also includes wavelengths between 300 ‐ 390 nm) is much higher.
However, the best results were obtained when photocatalytic ozonation was
applied, being NR the most active under visible radiation (68 % TOC removal),
likely due to its lower band gap energy as a consequence of higher Ce(III)
surface ratio, whereas NC presented the highest photocatalytic activity under
solar radiation (80 % TOC removal). This is in agreement with the highest
energy of {100} facets (thermodynamically more unstable) predominant in NC
CAPÍTULO 6 (CHAPTER 6)
226
with respect to NR, allowing higher photocatalytic activity for this material
because once excited the photoinduced holes exhibit superior oxidizing ability,
according to previously reported results [26].
0
20
40
60
80
100
O3/NRO3/NCO3
TO
C R
EM
OV
AL
(%)
Dark Visible Solar
Figure 6.5. TOC removal with the CeO2 catalysts and O3 processes. Experimental
conditions: pH0 = 6, T = 35 ‐ 40 °C, CDEET,0 = 5 mg L‐1, CCeO2 = 0.25 g L‐1, CO3,g inlet = 10
mg L‐1, Qg = 15 L h‐1 (O3/O2).
These results have also been confirmed by calculating the pseudo‐first order
apparent rate constant of TOC removal, kTOC, during photocatalytic ozonation
treatment, according to Eq. (6.1):
TOCTOC TOC
dC‐ k C
dt (6.1)
where CTOC is expressed in mg L‐1, t in min and kTOC in min‐1. Results are
summarized in Table 6.2 together with the correlation coefficient. Taking into
account the different irradiated surface area of the two catalysts (see Table 6.1),
values of normalized to SBET apparent rate constants (kTOC/SBET, min‐1 m‐2 g1) have
been calculated. The values obtained (Table 6.2) confirm the highest
photocatalytic activity of ceria NC. It is also noticeable that solar UV radiation
did not improve the mineralization rate when NR is used probably due to an
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227
increase of the recombination rate in the semiconductor.
Table 6.2. Pseudo‐first order apparent rate constant for TOC elimination during
photocatalytic ozonation runs.
CATALYST RADIATION kTOCx103
(min‐1)
kTOC/SBETx104
(min‐1 m‐2 g) R2
NC
Visible 7.2 5.1 0.98
Solar 11.8 8.4 0.96
NR
Visible 11.2 1.1 0.95
Solar 11.5 1.1 0.94
6.3.3. Considerations on the reaction mechanism of DEET photocatalytic
ozonation with CeO2 catalysts and solar radiation
According to the above results, photocatalytic ozonation of DEET using the
CeO2 catalysts studied here, may proceed through a complex mechanism with
the expected participation of ozone (direct reaction), and/or •HO radicals formed
from ozone decomposition (indirect reactions). Ozone can decompose into •HO
by several pathways: (1) in the dark, (2) by interaction with light, (3) by
interaction with the catalyst, and (4) by interaction with both light and the
catalyst. To discuss the possible contributions of each pathway Figure 6.6 shows
the evolution of DEET eliminated, dissolved ozone and hydrogen peroxide
formed during ozone‐based treatments using NC CeO2 and simulated solar
radiation.
Regarding to ozone‐direct reactions, a low reactivity of DEET towards ozone
has been previously reported (kO3‐DEET = 0.123 M‐1 s‐1 [21]). Taking into account
this value, at the conditions used in this work, the reaction between ozone and
CAPÍTULO 6 (CHAPTER 6)
228
DEET progress into the slow kinetic regime [32,33], being DEET degradation
rate by direct ozone reaction given by Eq. (6.2):
3 3
DEETO ‐DEET DEET O
dC‐ k C C
dt (6.2)
where CDEET is DEET molar concentration and CO3 the dissolved ozone molar
concentration. The expected evolution of DEET eliminated during ozonation in
the dark considering only its reaction with ozone is shown in Fig. 6.6. As it is
observed, the experimental conversion rate was much higher than that predicted
by Eq. (6.2), indicating that species different from O3 are the main responsible
for the degradation of the target compound. However, it is important to notice
that kO3‐DEET value was reported at 20 °C whereas the temperature during the
experiments was maintained at 35 ‐ 40 °C. Also, due to the low value of kO3‐DEET
some uncertainty on its determination should not be disregarded.
During all ozone‐based treatment pH evolved from 6 to ~ 4. At these pH
values, at the beginning of the experiment some ozone decomposition can be
promoted by hydroxide anion although it likely becomes negligible when pH
decreases [32]. Therefore, during DEET degradation by single ozonation in the
dark it can be hypothesized that direct interaction between DEET‐O3 acts as a
trigger of a chain mechanism that leads to the formation of ozonide and/or •HO
radicals, as has been previously reported for different organic compounds [33].
The key role of •HO during DEET single ozonation has been previously reported
[34].
The evolution of dissolved O3, also depicted in Fig. 6.6, shows a higher
accumulation during ozonation compared to treatments where a catalyst and/or
radiation were also present. This is indicative of the decomposition of ozone
over the catalyst surface (although almost inefficient in the dark) and/or through
its photolysis. According to the literature [30,35], both mechanisms could
promote the formation of hydroxyl radicals, thus increasing the mineralization
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229
rate of DEET.
Figure 6.6. Evolution of DEET converted, H2O2 formed and O3 dissolved during the O3
treatments applied using CeO2‐NC catalyst and simulated solar light. Experimental
conditions: pH0 = 6, T = 35 ‐ 40 °C, CDEET,0 = 5 mg L‐1, CCeO2 = 0.25 g L‐1, CO3,g inlet = 10
mg L‐1, Qg = 15 L h‐1 (O3/O2).
In addition, it is well known that hydrogen peroxide is commonly formed
through direct ozone and hydroxyl radical reactions. Accordingly, H2O2 was
detected during DEET ozonation. As observed in Fig. 6.6, during single
ozonation about the same concentration of H2O2 is formed respect to DEET
depleted. When light and/or NC catalyst were also present the ratio between
H2O2 in solution and DEET eliminated was much lower, and more markedly in
the presence of NC thus indicating the decomposition of H2O2 is accelerated by
CeO2 catalyst surface. It is important to notice that H2O2 decomposition seems to
be inefficient towards DEET degradation and mineralization according to the
evolution of DEET and TOC during catalytic ozonation (dark). However, the
0 20 40 60 80 100 120
TIME (min)
NC/O3/h
0 20 40 60 80 100 1200.0
1.0x10-5
2.0x10-5
3.0x10-5O
3/h
C (
M)
TIME (min)
NC/O3
0.0
1.0x10-5
2.0x10-5
3.0x10-5
O3
C (
M)
DEET (converted) H
2O
2 (formed)
DEET (simulated) O
3 (dissolved)
CAPÍTULO 6 (CHAPTER 6)
230
role of H2O2 as electron acceptor during solar light photocatalytic ozonation
using CeO2 cannot be disregarded [35]. The role of hydrogen peroxide as
electron acceptor as well as the contribution of the different reaction pathways
to DEET degradation and mineralization by photocatalytic ozonation using solar
light and CeO2 NC as catalyst will be the subject of future work.
6.4. CONCLUSIONS
Nanostructured CeO2 catalysts with different morphology (nanorods and
nanocubes) were active in photocatalytic ozonation treatment of DEET under
visible and simulated solar radiation. CeO2 with nanorod‐like structure
presented higher irradiated surface area due to its morphology and also higher
amount of surface defects and oxygen vacancies, leading to smaller band gap
energy and thus being apparently more active than CeO2 nanocubes under
visible light radiation. However, the presence of the exposed facets {100} in CeO2
nanocubes provokes an increased intrinsic photocatalytic activity in this
material with respect to nanorods under both visible and solar radiation.
Photocatalytic ozonation of DEET with CeO2 catalysts proceeds through a
complex reaction mechanism that involves at least an ozone‐direct reaction that
triggers a radical chain mechanism, indirect ozone reactions, ozone photolytic
and photocatalytic decomposition and hydrogen peroxide formation‐
decomposition. The importance of the different processes and the effect of
visible radiation will be the subject of a future work.
AKNOWLEDGEMENTS
Authors thank the Spanish MINECO/European Feder Funds (CTQ2012‐
35789‐C02‐01) and Junta de Extremadura (Ayuda Exp. GR15‐033) for economic
support. E. Mena thanks the Consejería de Empleo, Empresa e Innovación (Junta
de Extremadura) and European Social Fund for her FPI grant (Ref. PD12059).
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REFERENCES
[1] Oturan, M.A.; Aaron, J.J. “Advanced oxidation processes in
water/wastewater treatment: Principles and applications. A review”. Crit. Rev.
Environ. Sci. Technol. 44 (2014) 2577‐2641.
[2] Agustina, T.E.; Ang, H.M.; Vareek, V.K. “A review of synergistic effect of
photocatalysis and ozonation on wastewater treatment”. J. Photochem.
Photobiol. C Photochem. Rev. 6 (2005) 264‐273.
[3] Mehrjouei, M.; Müller, S.; Möller, D. “A review on photocatalytic ozonation
used for the treatment of water and wastewater”. Chem. Eng. J. 263 (2015) 209‐
219.
[4] Xiao, J.; Xie, Y.; Cao, H. “Organic pollutants removal in wastewater by
heterogeneous photocatalytic ozonation”. Chemosphere 121 (2015) 1‐17.
[5] Malato, S.; Fernández‐Ibáñez, P.; Maldonado, M.I.; Blanco, J.; Gernjak, W.
“Decontamination and disinfection of water by solar photocatalysis: Recent
overview and trends”. Catal. Today 147 (2009) 1‐59.
[6] Dong, S.Y.; Feng, J.L.; Fan, M.H.; Pi, Y.Q.; Hu, L.M.; Han, X.; Liu, M.L.; Sun,
J.Y.; Sun, J.H. “Recent developments in heterogeneous photocatalytic water
treatment using visible light‐responsive photocatalysts: a review”. RSC Adv. 5
(2015) 14610‐14630.
[7] Huang, W.; Gao, Y. “Morphology‐dependent surface chemistry and catalysis
of CeO2 nanocrystals”. Catal. Sci. Technol. 4 (2014) 3772‐3784.
[8] Aslam, M.; Qamar, M.T.; Tahir‐Soomro, M.; Ismail, I.M.I.; Salah, N.;
Almeelbi, T.; Gondal, M.A.; Hameed, A. “The effect of sunlight induced surface
defects on the photocatalytic activity of nanosized CeO2 for the degradation of
phenol and its derivatives”. Appl. Catal. B Environ. 180 (2016) 391‐402.
[9] Huang, X.; Beck, M.J. “Size‐dependent appearance of intrinsic qxO “activated
oxygen” molecules on ceria nanoparticles”. Chem. Mater. 27 (2015) 5840‐5844.
[10] Hernández‐Alonso, M.D.; Hungría, A.B.; Martínez‐Arias, A.; Fernández‐
García, M.; Coronado, J.M.; Conesa, J.C.; Soria, J. “EPR study of the
photoassisted formation of radicals on CeO2 nanoparticles employed for toluene
photooxidation”. Appl. Catal. B Environ. 50 (2004) 167‐175.
[11] Deng, W.; Chenn, D.; Chenn, L. “Synthesis of monodisperse CeO2 hollow
spheres with enhanced photocatalytic activity”. Ceram. Int. 41 (2015) 11570‐
11575.
CAPÍTULO 6 (CHAPTER 6)
232
[12] Muduli, S.K.; Wang, S.; Chen, S.; Fan‐Ng, C.; Huan, C.H.A.; Sum, T.C.; Soo,
H.S. “Mesoporous cerium oxide nanospheres for the visible‐light driven
photocatalytic degradation of dyes”. Beilstein J. Nanotechnol. 5 (2014) 517‐523.
[13] Yu, Y.; Zhu, Y.; Meng, M. “Preparation, formation mechanism and
photocatalysis of ultrathin mesoporous single‐crystal‐like CeO2 nanosheets”.
Dalton Trans. 42 (2013) 12087‐12092.
[14] Feng, T.; Wang, X.; Feng, G. “Synthesis of novel CeO2 microspheres with
enhanced solar light photocatalyic properties”. Mater. Lett. 100 (2013) 36‐39.
[15] Sifontes, A.B.; Rosales, M.; Méndez, F.J.; Oviedo, O.; Zoltan, T. “Effect of
calcination temperature on structural properties and photocatalytic activity of
ceria nanoparticles synthesized employing chitosan as template”. J. Nanomater.
Article ID 265797 (2013) 1‐9.
[16] Karunakaran, C.; Dhanalakshmi, R. “Semiconductor‐catalyzed degradation
of phenols with sunlight”. Sol. Energy Mater. Sol. Cells 92 (2008) 1315‐1321.
[17] da Silva, M.F.P.; Soeira, L.S.; Daghastanli, K.R.P.: Martins, T.S.; Cuccovia,
I.M.; Freire, R.S.; Isolani, P.C. “CeO2‐catalyzed ozonation of phenol: The role of
cerium citrate as precursor of CeO2”. J. Therm. Anal. Calorim. 102 (2010) 907‐
913.
[18] Faria, P.C.C.; Orfao, J.J.M.; Pereira, M.F.R. “A novel ceria‐activated carbon
composite for the catalytic ozonation of carboxylic acids”. Catal. Comm, 9 (2008)
2121‐2126.
[19] Gonçalves, A.G.; Orfao, J.J.M.; Pereira, M.F.R. “Ozonation of bezafibrate
over ceria and ceria supported on carbon materials”. Environ. Technol. 36 (2015)
776‐785.
[20] Adams, W.A.; Impellitteri, C.A. “The photocatalysis of N,N‐diethyl‐m‐
toluamide (DEET) using dispersions of Degussa P‐25 TiO2 particles”. J.
Photochem. Photobiol. A Chem. 202 (2009) 28‐32.
[21] Benítez, F.J.; Acero, J.L.; García‐Reyes, J.F.; Real, F.J.; Roldán, G.; Rodríguez,
E.; Molina‐Díaz, A. “Determination of the reaction rate constants and
decomposition mechanisms of ozone with two model emerging contaminants:
DEET and Nortriptyline”. Ind. Eng. Chem. Res. 52 (2013) 17054‐17073.
[22] Mai, H.X.; Sun, L.D.; Zhang, Y.W.; Si, R.; Feng, W.; Zhang, H.P.; Liu, H.C.;
Yan, C.H. “Shape‐selective synthesis and oxygen storage behavior of ceria
nanopolyhedra, nanorods, and nanocubes”. J. Phys. Chem. B 109 (2005) 24380‐
24385.
[23] Tauc, J. “Absorption edge and internal electric field in amorphous
PAPER 4: Nanostructured CeO2 as catalysts for different
AOPs based in the application of ozone and simulated solar radiation
233
semiconductors”. Mater. Res. Bull. 5 (1970) 721‐729.
[24] Bader, H.; Hoigné, J. “Determination of ozone in water by the indigo
method”. Water Res. 15 (1981) 449‐456.
[25] Masschelein, W.; Denis, M.; Ledent, R. “Spectrophotometric determination
of residual hydrogen peroxide”. Water & Sewage Works (1977) 69‐72.
[26] Jiang, D.; Wang, W.; Zhang, L.; Zheng, Y.; Wang, Z. “Insights into the
surface‐defect dependence of photoreactivity over CeO2 nanocrystals with well‐
defined crystal facets”. ACS Catal. 5 (2015) 4851‐4858.
[27] Muñoz‐Batista, M.J.; Gómez‐Cerezo, M.N.; Kubacka, A.; Tudela, D.;
Fernández‐García, M. “Role of interface contact in CeO2‐TiO2 photocatalytic
composite materials”. ACS Catal. 4 (2014) 63‐72.
[28] Zdravkovic, J.; Simovic, B.; Golubovic, A.; Poleti, D.; Veljkovic, I.;
Scepanovic, M.; Brankovic, G. “Study of CeO2 nanopowders obtained by
hydrothermal method from various precursors”. Ceramics Int., 141 (2015) 1970‐
1979.
[29] Mena, E.; Rey, A.; Contreras, S.; Beltrán, F.J. “Visible light photocatalytic
ozonation of DEET in the presence of different forms of WO3”. Catal. Today 252
(2015) 100‐106.
[30] Sánchez, L.; Domènech, X.; Casado, J.; Peral, J. “Solar activated ozonation of
phenol and malic acid”. Chemosphere 50 (2003) 1085‐1093.
[31] Quiñones, D.H.; Rey, A.; Álvarez, P.M.; Beltrán, F.J.; Plucinski, P.K.
“Enhanced activity and reusability of TiO2 loaded magnetic activated carbon for
solar photocatalytic ozonation”. Appl. Catal. B Environ. 144 (2014) 96‐106.
[32] Beltran, F.J. “Ozone reaction kinetics for water and wastewater systems”.
Boca Raton, CRC Press, 2004, Florida (USA).
[33] Rey, A.; Mena, E.; Chávez, A.M.; Beltrán, F.J.; Medina, F. “Influence of
structural properties on the activity of WO3 catalysts for visible light
photocatalytic ozonation”. Chem. Eng. Sci. 126 (2015) 80‐90.
[34] Acero, J.L.; Benítez, F.J.; Real, F.J.; Rodríguez, E. “Elimination of selected
emerging contaminants by the combination of membrane filtration and chemical
oxidation processes”. Water Air Soil Pollut. 226 (2015) 139‐1‐139‐14.
[35] Mena, E.; Rey, A.; Acedo, B.; Beltrán, F.J.; Malato, S. “On ozone‐
photocatalysis synergism in black‐light induced reactions: Oxidizing species
production in photocatalytic ozonation versus heterogeneous photocatalysis”.
Chem. Eng. J. 204‐206 (2012) 131‐140.
CAPÍTULO 7 (CHAPTER 7) PAPER 5: WO3-TiO2 based catalysts for the simulated solar radiation assisted photocatalytic ozonation of emerging contaminants in a municipal wastewater treatment plant effluent
A. Rey, P. García-Muñoz, M.D. Hernández-Alonso, E. Mena, S. García-Rodríguez, F.J. Beltrán
Applied Catalysis B: Environmental 154-155 (2014) 274-284
ABSTRACT. This work is focused on the use of TiO2-WO3 photocatalysts for the
removal of a mixture of emerging contaminants through photocatalytic ozonation using
simulated solar light as radiation source. Own-made TiO2-WO3 catalysts with WO3
content around 4 wt. % were synthetized using P25 and titanate nanotubes as TiO2
starting materials. Photocatalysts were thoroughly characterized by means of ICP-OES,
N2 adsorption-desorption isotherms, XRD, TEM, Raman, XPS and DR-UV-Vis
spectroscopy. Photocatalytic ozonation using WO3 based titanate nanotubes composite
gave place to complete removal of emerging contaminants (caffeine, metoprolol and ibuprofen in a Municipal Wastewater Treatment Plant effluent aqueous matrix) in less than 40 min with TOC removal up to 64 % after 2 h. The highest catalytic activity of this
material in the reaction under study compared to bare TiO2 is due to several effects such
as higher activity under visible light radiation joined to an increase in adsorption capacity
of organic compounds and catalytic ozonation ability caused by the presence of WO3.
Keywords: Photocatalytic ozonation, tungsten oxide-titanium dioxide, titanate nanotubes, ozone, emerging contaminants.
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
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7.1. INTRODUCTION
Emerging contaminants (ECs), such as pharmaceuticals and personal care
products, which usually present endocrine disrupting activity, are frequently
detected in wastewater and aquatic environments [1,2]. These contaminants can
cause severe adverse effects in human and wildlife and their removal is of a
great concern on environmental and health risk management [2,3]. However
these compounds are hardly biodegradable so they are not removed by
conventional treatment at municipal wastewater treatment plants (MWWTPs),
and therefore advanced treatment technologies are required [1].
Advanced oxidation processes (AOPs) which involve the formation of highly
reactive species such as hydroxyl radical ( HO ), have demonstrated their
efficiency to degrade many ECs transforming them into harmless products [4].
Among these AOPs, photocatalytic ozonation, i.e. the combination of
photocatalytic oxidation and ozone processes, can greatly enhance the rates of
ECs degradation and, specially, the mineralization achieved by the single
processes [5‐7]. In fact, a synergistic effect has been observed between ozone and
irradiated photocatalyst due to both the strong electron trapping effect of ozone,
that avoids, to some extent, the electron‐hole recombination process, and to the
reaction of ozone with the superoxide ion radical ( 2O ), intermediate species in
photocatalytic reactions. In these two steps the ozonide ion radical ( 3O ) is
generated and further transformed intoHO , thus increasing the concentration
of this highly reactive species in the aqueous medium [5,8,9].
One of the most important aspects in the application of photocatalytic
wastewater treatment technologies is the cost derived from the radiation source
operation and maintenance. The use of solar energy as radiation source avoids
this drawback leading to more economic technologies [10]. However, most of
the photocatalytic ozonation studies conducted to date make use of TiO2 as
photocatalyst that only uses around 5 % of solar radiation (UV range). Different
methods to solve this limitation, mainly applied to ozone‐free photocatalytic
CAPÍTULO 7 (CHAPTER 7)
238
oxidation processes, have been developed, such as ion doping, modification of
TiO2 surface with noble metals, metal ion implantation or coupling
semiconductors [10,11]. In this line, composite materials of coupled WO3 and
TiO2 semiconductors can extend the absorption of radiation to the visible region
as a result of the lower band gap energy of WO3 (around 2.7 eV), and also allow
a wide electron‐hole separation avoiding recombination to some extent [11].
TiO2‐WO3 coupled semiconductors have been successfully applied to the
degradation of different organic contaminants in water through photocatalytic
oxidation under visible light radiation [12‐16]. In addition, these materials
combined with gold nanoparticles have also been tested for photocatalytic
ozonation of model compounds such as oxalic acid and 2,4,6‐trinitrotoluene
(TNT) giving place to promising results [17,18].
This work has been focused on the use of different TiO2‐WO3 composite
catalysts to extend the radiation absorption to visible light and make better use
of the solar emission spectrum in photocatalytic ozonation process applied to
wastewater detoxification. Two different TiO2 supports, commercial TiO2 P25
and own‐made titanate nanotubes, which present very different textural and
structural properties, have been used. The synthetized materials were tested in
photocatalytic ozonation of ibuprofen (IBP), metoprolol (MTP) and caffeine
(CAF) spiked in a municipal wastewater effluent from a secondary treatment
and using simulated solar light as radiation. These contaminants, IBP, a
nonsteroidal anti‐inflammatory drug for pain relief and fever reduction; MTP, a
β‐blocker used for several cardiovascular diseases; and CAF, a stimulant drug
mainly from coffee consumption, have been selected because of their frequent
presence in municipal wastewater [2].
7.2. EXPERIMENTAL SECTION
7.2.1. Catalysts preparation
Aeroxide TiO2 P25 (anatase/rutile 80/20, 21 nm crystal size) was used as the
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
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starting photocatalytic material. The procedure to obtain titanate nanotubes
catalyst (NT) has been reported in previous works [19,20]. Briefly, 1 g of the
precursor P25 was hydrothermally treated at 130 °C in 70 mL of 10 M NaOH in a
Teflon‐lined autoclave, during 48 h. The mixture was stirred for 30 min before
and after the thermal treatment. In a second stage, the obtained powders were
thoroughly washed using diluted HCl. The powders were recovered from the
solution by centrifugation, dried at 100 °C overnight and then calcined in air
atmosphere for 3 h at 350 °C.
P25 and NT photocatalysts were coated with nanosized WO3 particles (P25‐
WO3 and NT‐WO3 samples) as described elsewhere [18]. First 0.216 g H2WO4
were added to 100 mL of ultrapure water and then aqueous ammonia solution
was added drop wise until the tungstic acid was completely dissolved. Further 5
g of P25 or NT were added under continuous stirring and the obtained mixture
was stirred for 30 min and then acidified to pH = 4 with 0.5 M HCl. Then 10 mL
of an aqueous solution of oxalic acid 0.1 M was added to the mixture and stirred
for 1 h at 40 °C to prevent the aggregation of WOX particles in the precipitate.
Finally, the solid was recovered by filtration, dried at 110 °C for 2 h and then
calcined in air atmosphere for 2 h at 420 °C. The TiO2‐WO3 thus prepared had a
theoretical amount of WO3 of 3.8 wt. %. For comparative purposes, NT was also
calcined in air atmosphere for 2 h at 420 °C (sample NT‐T).
7.2.2. Characterization
Total tungsten content of the catalysts was analyzed by inductively coupled
plasma with an ICP‐OES Optima 3300DV (Perkin‐Elmer) after acidic microwave
digestion of the samples.
BET surface area and pore volume of the photocatalysts were determined
from their nitrogen adsorption–desorption isotherms obtained at ‐196 °C using
an Autosorb 1 apparatus (Quantachrome). Prior to analysis the samples were
outgassed at 250 °C for 12 h under high vacuum (< 10−4 Pa).
CAPÍTULO 7 (CHAPTER 7)
240
The crystalline phases present in the photocatalysts were inferred from their
X‐ray diffraction (XRD) patterns recorded using a powder Bruker D8 Advance
XRD diffractometer with a Cu Kα radiation (λ = 0.1541 nm). The data were
collected from 2θ = 20° to 80° at a scan rate of 0.02 s−1 and 1 s per point.
A JEM‐2100F 200 kV transmission electron microscope (JEOL Ltd.) and a
TEM Tecnai G20 Twin 200 kV transmission electron microscope (FEI Company)
were used to study the crystallinity and morphology of the samples.
Raman spectra were obtained using a Nicolet Almega XR Dispersive micro‐
Raman (Thermo Scientific) with a spectral resolution of 2 cm‐1. The samples
were excited at 633 nm with the laser beam power at 100 % and 200 scans
accumulation.
X‐ray photoelectron spectra (XPS) were obtained with a Kα Thermo Scientific
apparatus with an Al Kα (h = 1486.68 eV) X‐ray source using a voltage of 12 kV
under vacuum (2x10−7 mbar). Binding energies were calibrated relative to the
C1s peak at 284.6 eV. The resulting XPS peaks were curve‐fitted to a
combination of Gaussian and Lorentzian functions using a Shirley type
background for peak analysis.
Diffuse reflectance UV‐Vis spectroscopy (DR‐UV‐Vis) measurements, useful
for the determination of the semiconductor band gap, were performed with a
UV‐Vis‐NIR Cary 5000 spectrophotometer (Varian‐Agilent Technologies)
equipped with an integrating sphere device.
7.2.3. Catalytic activity measurements
Ibuprofen sodium salt (IBP), metoprolol tartrate (MTP) and caffeine (CAF)
were used as target contaminants to test the catalytic activity of the synthesized
materials. They were added to a real municipal wastewater effluent (MWW)
taken from Badajoz MWWTP (Badajoz, Spain) designed for 225,000 inhabitants
(population equivalent) with an average inlet flow of 37,500 m3 day‐1. Effluents
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were collected downstream of the MWWTP secondary biological treatment,
filtered and stored at ‐20 °C until use.
Photocatalytic experiments were carried out in semi‐batch mode in a
laboratory‐scale system consisting of a 0.5 L glass‐made spherical reactor,
provided with a gas inlet, a gas outlet and a liquid sampling port. The reactor
was placed in the chamber of a commercial solar simulator (Suntest CPS, Atlas)
provided with a 1500 W air‐cooled Xe arc lamp with emission restricted to
wavelengths over 320 nm using quartz, glass and polyester cut‐off filters. The
irradiation intensity was kept at 550 W m−2 and the temperature of the system
was maintained between 20 and 40 °C throughout the experiments. If required,
a laboratory ozone generator (Anseros Ozomat Com AD‐02) was used to
produce a gaseous ozone‐oxygen stream that was fed to the reactor. In that case,
the ozone concentration was monitored by an Anseros Ozomat GM‐6000Pro gas
analyzer. A scheme of the experimental set‐up is depicted in Figure 7.1.
Figure 7.1. Scheme of the experimental set‐up.
In a typical photocatalytic ozonation experiment, the reactor was first loaded
with 0.5 L of the MWW effluent containing 2 mg L‐1 of each EC. Then, 0.25 g of
the catalyst were added and the suspension was stirred in the dark for 30 min
while bubbling air into the system. After this dark stage, the lamp was switched
on and, simultaneously, a mixture of ozone‐oxygen (10 mg L‐1 ozone
concentration) was fed to the reactor at a flow rate of 20 L h−1. The irradiation
time for each experiment was 2 h. Samples were withdrawn from the reactor at
CAPÍTULO 7 (CHAPTER 7)
242
intervals and filtered through a 0.2 μm PET membrane to remove the
photocatalyst particles.
Experiments of adsorption (i.e., absence of radiation and ozone), photolysis
(i.e., radiation in absence of catalysts and ozone), photocatalytic oxidation (i.e.,
radiation and catalyst in absence of ozone), ozonation (i.e., absence of radiation
and catalyst), and catalytic ozonation (i.e., absence of radiation) were also
carried out for comparative analysis. In addition, some photocatalytic oxidation
experiments of IBP were carried out under visible radiation using a UV cut‐off (λ
< 390 nm) polyester filter.
ECs concentrations were analyzed by HPLC‐DAD (Hitachi, Elite LaChrom)
using a Phenomenex C‐18 column (5 μm, 150 mm long, 3 mm diameter) as
stationary phase and 0.5 mL min−1 of acetonitrile‐acidified water (0.1 % H3PO4)
as mobile phase (from 5 % to 60 % in acetonitrile during 25 min and 10 min re‐
equilibration time). Identification and quantification was carried out at 220 nm.
Total organic carbon (TOC) and inorganic carbon (IC) were measured using a
Shimadzu TOC‐VSCH analyzer. Aqueous ozone was measured following the
indigo method using a UV‐Vis spectrophotometer (Evolution 201,
Thermospectronic) set at 600 nm [21]. Ozone in the gas phase was continuously
monitored by means of an Anseros Ozomat GM‐6000Pro analyzer. Finally,
chemical oxygen demand (COD) was determined by means of the dichromate
method using Dr. Lange cuvette test, biological oxygen demand (BOD5) was
analyzed by the respirometric method using an Oxitop® WTW system,
aromaticity as the absorbance of the sample at 254 nm and phosphates by means
of colorimetric methods using a Merck Spectroquant kit (UV‐Vis
spectrophotometry, Evolution 201 from Thermospectronic).
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7.3. RESULTS AND DISCUSSION
7.3.1. Characterization of the photocatalysts
Table 7.1 summarizes tungsten content, some textural parameters, band gap
energy and absorption edge wavelength of the photocatalysts. It can be seen that
the W mass composition (expressed as WO3) of the TiO2‐WO3 catalysts was close
but somewhat higher to the nominal value expected (3.8 wt. %).
Table 7.1. Properties of the catalysts.
CATALYST WO3
(wt. %)
SBET
(m2 g‐1)
VP
(cm3 g‐1)
WXPS
(at %)
Eg
(eV)
λ
(nm)
P25 ‐‐‐ 52 0.25 ‐‐‐ 3.19 389
P25‐WO3 4.1 49 0.28 1.6 3.05 407
NT ‐‐‐ 320 1.39 ‐‐‐ 3.18 390
NT‐WO3 4.5 195 1.17 0.9 2.98 416
NT‐T ‐‐‐ 208 1.19 ‐‐‐ ‐‐‐ ‐‐‐
Figure 7.2 shows N2 adsorption‐desorption isotherms where it can be noticed
that P25 based catalysts are non‐porous materials with type II isotherms
whereas NT catalysts presented type IV isotherms with H3 hysteresis loops
characteristics of mesoporous solids, according to IUPAC classifications [22]. On
the other hand, the incorporation of W did not change the isotherm aspect of
these photocatalysts respect to their supports. Main differences are observed for
NT catalysts where the procedure of W incorporation has changed the slope and
shifted the hysteresis loop in NT‐WO3 respect to NT, indicating a lower
porosity. This is mainly related to the heat‐treatment applied, as can be deduced
from the similarity of the isotherms of NT‐T and NT‐WO3. Calcination at 420 °C
triggers structural changes as discussed below [20]. From these analyses, specific
surface area (SBET) and pore volume (VP) were calculated (see Table 7.1). As can
be observed P25 based catalysts show SBET values according to that indicated by
CAPÍTULO 7 (CHAPTER 7)
244
the supplier for P25 (around 50 m2 g‐1), and similar pore volume around 0.25 m3
g‐1. Thus, the incorporation of W and calcination of the composite material does
not significantly affect its textural properties. On the other hand, NT presented
higher SBET and pore volume, according to its mesoporous structure as it was
previously reported for similar materials [20]. In this case, the incorporation of
W has favored a decrease in the surface area and pore volume of the NT‐WO3
catalyst mainly due to the calcination step, although some pore blockage after W
deposition cannot be disregarded according to the lower values of textural
parameters of this catalyst respect to the NT‐T sample.
0.0 0.2 0.4 0.6 0.8 1.00
200
400
600
800
1000
1200
N2
AD
SO
RB
ED
VO
LU
ME
(cm
3 g-1
)
P/P0
P25 P25-WO
3
NT NT-WO
3
NT-T
Figure 7.2. N2 adsorption‐desorption isotherms of the catalysts.
Structural characterization of the photocatalysts was accomplished by means
of XRD, TEM and Raman analyses and the results are depicted in Figure 7.3 to
Figure 7.5, respectively. P25 based catalysts gave place to similar XRD patterns
with main diffraction peaks attributable to anatase and rutile TiO2 phases. The
absence of peaks attributable to W species could be due to the low content in the
catalysts and/or to the high dispersion of W. This was also confirmed, as Figure
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
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7.4 shows, by TEM micrographs where W particles were undistinguishable. In
addition, heat‐treatment applied to P25‐WO3 photocatalyst does not produce
any significant change in the TiO2 structure. Regarding NT photocatalyst, a
shifted diffraction peak position at 24.9° respect to 25.4° for anatase is observed,
indicating a titanate structure [20]. TEM micrograph confirmed nanotubular
morphology for NT sample (Fig. 7.4). On the other hand, the incorporation of W
(NT‐WO3 photocatalyst) led to a transformation of the titanate structure into
anatase as evidenced by the shift of the corresponding diffraction peaks (e.g.
most intense from 24.9° to 25.3°) together with an increase in their intensity. This
transformation is mainly related to the calcination step at 420 °C as the XRD
pattern and TEM micrograph of the NT‐T sample (subjected to the same heat‐
treatment) confirmed, and also according to the results reported elsewhere [20].
TEM micrographs corroborated the transformation of titanate nanotubular
structures to elongated anatase particles (Fig. 7.4). Fig. 7.5 shows Raman spectra
of TiO2‐WO3 photocatalysts. In Fig. 7.5(A) vibration peaks at 395, 517 and 638
cm‐1 were observed in both catalysts (P25‐WO3 and NT‐WO3) which are
unambiguously attributed to anatase phase of TiO2 [23,24]. On the other hand,
only for P25 based material, rutile phase is observed as a broad peak at 448 cm‐1
according to the P25 structural composition [23,24]. These results also confirm
the transformation of titanate structure into anatase after the calcination step
[20]. The lowest intensity observed in the Raman spectra for the anatase
vibrations can be related to the lowest crystallinity of NT‐WO3 photocatalyst
according to XRD results [23]. On the other hand, Fig. 7.5(B) shows main W
contributions to the Raman spectra. The band located around 792 cm‐1
corresponds to weak second‐order feature of anatase assigned to the first over‐
tone band at 395 cm‐1 [25]. At around 954 cm‐1 appears another contribution
assigned to the stretching mode of terminal W=O bond which is characteristic of
two‐dimensional tungsten oxide surface species WOX [25,26]. In addition, the
weak band located at around 1042 cm‐1 could be related to the presence of WOX
species in tetrahedral coordination [27]. The observed contributions at 954 and
CAPÍTULO 7 (CHAPTER 7)
246
1042 cm‐1 together with the absence of any signal close to 800 cm‐1 suggest that
added W does not contribute to the formation of crystalline WO3 [27]. These
results are in a good agreement with XRD and TEM observations. Again it is
noteworthy the lowest intensity of the NT‐WO3 spectrum as commented above.
20 30 40 50 60 70 80
NT-T
R
AA
RR
AA
A
INT
EN
SIT
Y (
a.u
.)
2 (o)
P25
P25-WO3
NT
NT-WO3
A
R
Figure 7.3. XRD patterns of the catalysts. Crystalline phases detected: Anatase (A); Rutile
(R).
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Figure 7.4. TEM images of the catalysts.
CAPÍTULO 7 (CHAPTER 7)
248
Figure 7.5. Raman spectra of P25‐WO3 and NT‐WO3 photocatalysts.
Surface chemical composition of the catalysts was analyzed by XPS. Full
spectra depicted in Figure 7.6(A‐D) confirm the presence of O, Ti and C in all the
samples (O1s, Ti2p and C1s peaks). In addition, the peaks of W4d and W4f
spectral regions appear in the full spectra of the TiO2‐WO3 photocatalysts
confirming the presence of W in their surfaces. No N signal was detected thus
confirming the absence of N‐doping effect due to the NH3 used for the
incorporation of W in the composite catalysts. The surface content of W was
calculated from peak areas and Wagner atomic sensitive factors [28]. Results are
summarized in Table 7.1 where it can be noticed a lower W surface content for
the NT‐WO3 catalyst. This could be related to the higher porosity of the NT
sample which can favor the incorporation of W in inner regions of the porous
structure. On the other hand, Fig. 7.6(E) shows the high‐resolution Ti2p spectral
region of P25‐WO3 catalyst as an example. The binding energy of the Ti2p1/2 and
Ti2p3/2 core levels at 464.7 eV and 459.0 eV, respectively, together with their
separation of 5.7 eV confirm the valence state of Ti as Ti4+ in TiO2 [29,30]. Fig.
7.6(F) displays the high‐resolution and peak‐fitting results of W4f and Ti3p XPS
spectra of P25‐WO3 catalyst as an example. Analysis of the W4f region is
750 875 1000 1125 1250
1042
954
P25-WO3
NT-WO3
INT
EN
SIT
Y (
a.u.
)RAMAN SHIFT (cm-1)
792
300 400 500 600 700
448
638
517
INT
EN
SIT
Y (
a.u.
)
RAMAN SHIFT (cm-1)
P25-WO3
NT-WO3
395
(A)
(B)
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complicated due to the interference from the Ti3p level of the TiO2. However,
the position of the Ti3p peak can be fixed and its area calculated from the Ti2p
peaks areas [28]. This makes possible to distinguish between the two signals
(Ti3p and W4f) by deconvolution procedure, and to determine the valence of W
from the position of the W4f level. The binding energy of the peaks located at
37.8 eV and 35.7 eV corresponds to W4f5/2 and W4f7/2 components, respectively,
and their area ratio of 3:4 confirms the presence of W as W6+ in WO3 [18,30].
Similar results were obtained for NT‐WO3 catalyst (not shown).
The diffuse reflectance UV‐Vis spectra of the photocatalysts (Figure 7.7)
showed a higher optical absorbance in the visible region (above ca. 400 nm) for
TiO2‐WO3 catalysts due to the WO3 loading. The optical energy band gap (Eg)
was calculated by means of Tauc’s expression and these results are summarized
in Table 7.1 together with the wavelength of absorption edge. These values are
approximate due to the need of extrapolation of the resulting curve (not shown).
However a value of 3.19 eV was calculated for P25 catalyst similar to 3.2 eV
previously reported for bare TiO2 [11]. It is noticeable a decrease of band gap
energy by coupling the TiO2 with WO3 from ca. 3.2 eV to 3.05 eV and 2.98 eV
calculated for P25‐WO3 and NT‐WO3 catalysts, respectively, which is in
agreement to previously reported values for similar materials [15]. These results
indicate that both P25‐WO3 and NT‐WO3 can be promising photocatalysts to be
used under visible light.
CAPÍTULO 7 (CHAPTER 7)
250
Figure 7.6. XPS full spectra of the catalysts (A, B, C and D). High‐resolution XPS spectra
of Ti2p (E) and Ti3p/W4f (F) spectral regions of P25‐WO3 catalyst.
600 500 400 300 200 100 0
NT-WO3
Ti3pW4f
W4dC1s
Ti2p
O1s
BINDING ENERGY (eV)
600 500 400 300 200 100 0
NT
Ti3pC1s
Ti2p
O1s
CP
S (
a.u.
)
BINDING ENERGY (eV)
600 500 400 300 200 100 0
Ti3pW4fC1s
Ti2p
O1s P25-WO3
BINDING ENERGY (eV)
W4d
600 500 400 300 200 100 0
CP
S (
a.u.
)
BINDING ENERGY (eV)
P25O1s
Ti2p
C1sTi3p
468 464 460 456
Ti2p3/2
459.0
CP
S (
a.u.
)
BINDING ENERGY (eV)
Ti2p1/2
464.7
5.7 eV
42 40 38 36 34 32
W4f7/2
35.7
Ti3p37.3
W4f5/2
37.8
BINDING ENERGY (eV)
(A) (B)
(C) (D)
(E) (F)
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300 350 400 450 500 550 6000.0
0.5
1.0
1.5
AB
SO
RB
AN
CE
(u.
a.)
WAVELENGTH (nm)
P25
P25-WO3
NT
NT-WO3
Figure 7.7. DR UV–Vis spectra of the catalysts.
7.3.2. Visible light response of the photocatalysts
The effectiveness of the catalysts in the use of visible light was tested using
IBP as target compound by cutting off the wavelengths lower than 390 nm of the
Xe lamp in the solar simulator. Results of photocatalytic depletion of IBP and its
mineralization with all the catalysts synthetized are presented in Figure 7.8.
Regarding IBP depletion (Fig. 7.8(A)), once confirmed the absence of IBP
degradation through direct photolysis as expected according to its UV‐Vis
absorption spectrum, it can be noticed the positive effect of WO3 loading on the
photocatalytic activity of TiO2‐WO3 photocatalysts under visible light irradiation
compared to bare TiO2. In addition, it is noticeable the lowest IBP degradation
rate with the sample NT‐T, thus indicating that the heat‐treatment and different
structural properties of calcined NT are not responsible for the high catalytic
activity of the NT‐WO3 photocatalyst. Also, once the N‐doping effect has been
ruled out, the enhancement observed in the photocatalytic activity of the WO3
coupled materials can be related to the lowest band gap energy observed (Table
7.1) due to WO3 presence, which makes them easily excited by visible light [15].
CAPÍTULO 7 (CHAPTER 7)
252
In addition, the presence of WO3 can promote the charge transfer between
photogenerated electrons from the conduction band of TiO2 to the WO3
conduction band, accompanied by holes transfer from the valence band of WO3
to the TiO2 valence band. The charge separation mechanism has been previously
reported [15,17,18,27], and provokes an increase in the lifetime of the
photogenerated electron/hole pair avoiding its recombination to some extent. As
a consequence of both higher visible light absorption and lower recombination
rate, the photonic efficiency (i.e. reacted molecules/incident photons) of the
photocatalytic process is increased. The improvement of composite materials
respect their corresponding TiO2 was higher for NT‐WO3 photocatalyst than for
P25‐WO3. Taking into account that NT‐T sample led to a lower degradation rate
than NT, the formation of elongated anatase particles from the titanate
structures seems not to be the reason of this behavior. Thus, it seems plausible
that NT material presented a better distribution and higher dispersion of WO3
particles since also offered a developed porous structure and higher surface area
than P25 together with a large ion‐exchange capacity [19,20]. This hypothetic
higher dispersion together with the slightly higher amount of WO3 in NT‐WO3
(Table 7.1) could be the reasons of the highest visible light absorption capacity
and hence the improvement in NT‐WO3 photocatalytic activity compared to
P25‐WO3, although additional characterization analysis confirming this
hypothesis would be needed.
Regarding TOC evolution (Fig. 7.8(B)) it is noticeable the mineralization rate
observed with the NT‐WO3 photocatalyst reaching 40 % TOC removal in 2 h
whereas P25‐WO3 did not show a significant mineralization degree. These and
the previous results point out the highest photocatalytic activity of NT‐WO3
photocatalyst under visible light radiation.
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Figure 7.8. Time evolution of IBP dimensionless concentration (A) and dimensionless
TOC (B) during photocatalytic oxidation under visible light radiation (λ = 390 ‐ 800 nm).
Conditions: pH = 6.5 T = 20 ‐ 40 °C, CIBP,0 = 5 mg L‐1, CCAT = 0.5 g L‐1, Qg = 20 L h‐1 (O2).
7.3.3. Photocatalytic degradation of ECs in MWW
The effectiveness of the photocatalysts in a more realistic application was
studied using a MWW effluent as aqueous matrix spiked with IBP, MTP and
CAF. The average values of the main MWW secondary effluent parameters are
summarized in Table 7.2. It can be highlighted the amount of
carbonate/bicarbonate as inorganic carbon (IC) and their buffer role maintaining
the pH of the reaction medium around pH = 8.3.
Regarding the process applied, Table 7.3 summarizes the main results of ECs
removal at 120 min upon the treatments tested, together with Figure 7.9 and
Figure 7.10 that show the time‐evolution of ECs and TOC respectively, only for
NT‐WO3 catalyst as an example. It can be noticed that direct photolysis exerts no
effect on the case of IBP removal and only 5 and 7 % removal was observed for
CAF and MTP, respectively. The time‐evolution of ECs concentration during
photolysis is also shown in Fig. 7.9. Although these compounds do not absorb
radiation in the wavelength range used in this work, MWW content could
provoke indirect photolysis reactions [31,32]. The insignificant mineralization
during photolysis can be observed in Fig. 7.10 and the low value of final TOC
-30 0 30 60 90 1200.0
0.2
0.4
0.6
0.8
1.0
CIB
P/C
IBP
,0
TIME (min)
Dark
-30 0 30 60 90 1200.0
0.2
0.4
0.6
0.8
1.0
Photolysis P25 NT P25-WO3
NT-WO3
NT-T
TO
C/T
OC
0
TIME (min)
Dark
0
0
CT
OC
/CT
OC
,0
CAPÍTULO 7 (CHAPTER 7)
254
removal is also presented in Table 7.4. Taking into account that the contribution
of the initial concentration of ECs to the initial TOC content is around 10 % (3.7
mg L‐1), the mineralization observed would be mainly associated to the MWW
TOC content.
Table 7.2. Characterization of MWW effluent before and after some treatments with
NT‐WO3 catalyst (t = 120 min).
PARAMETER Before
treatment*
Photocatalytic
oxidation Ozonation
Catalytic
ozonation
Photocatalytic
ozonation
TOC
(mg C L‐1) 35.3 22.8 26.9 20.2 12.8
IC
(mg C L‐1) 42.0 40.1 38.3 41.5 36.4
pH 8.31 8.34 8.29 8.28 8.33
Absorbance
254 nm 0.253 0.164 0.116 0.058 0.017
COD
(mg O2 L‐1) 51 46 41 33 17
BOD5
(mg O2 L‐1) 32 n.m n.m n.m n.m
Phosphate
(mg L‐1) 4 n.m n.m n.m n.m
*Average values between all batches
n.m. not measured
Regarding the adsorption capacity of the catalysts, none of them gave place
to ECs removal higher than 10 % (Table 7.3). It is noticeable that no significant
changes are observed in terms of ECs adsorption either due to W incorporation
or textural properties modification in NT series, although the modification of the
adsorption capacity of these materials has been reported elsewhere [18].
However, a different behavior is observed for the organic matter content of the
MWW since the adsorption capacity of the catalysts increased around 15 % in
both series, P25 and NT, after W incorporation (see values in Table 7.4). It is also
noticeable a somewhat higher adsorption capacity in NT series respect to P25
due to its more developed porous structure. The fact that NT series, in which a
significant decrease in textural parameters was observed after calcination,
presented the same increase in adsorption capacity than P25 series points out
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that this phenomenon is mainly related to the W species in the catalysts surface.
In fact, it has been reported that the presence of a monolayer of WOX species on
TiO2 can significantly increase the surface acidity leading to the fact that
composite materials TiO2‐WO3 can adsorb more organic reactants [18,33].
Table 7.3. ECs removal (%) after 120 min reaction upon the different treatments applied.
TREATMENT EC No catalyst
Photolysis
IBP 0
CAF 5
MTP 7
TREATMENT EC P25 P25‐WO3 NT NT‐WO3
Adsorption
IBP 7 8 10 7
CAF 8 7 4 6
MTP 6 8 8 6
Photocatalysis
IBP 73 76 28 70
CAF 77 79 31 82
MTP 92 90 61 91
TREATMENT EC No catalysts/All catalysts
Ozonation
Catalytic ozonation
Photocatalytic ozonation
IBP
> 99.9 CAF
MTP
CAPÍTULO 7 (CHAPTER 7)
256
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
CIB
P/C
IBP
,0
TIME (min)
Dark
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
CC
AF/C
CA
F,0
TIME (min)
Dark Adsorption Photolysis Photocatalytic oxidation Ozonation Catalytic ozonation Photocatalytic ozonation
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
CM
TP/C
MT
P,0
TIME (min)
Dark
Figure 7.9. Time evolution of ECs dimensionless concentration during all the treatments
applied to MWW effluent under simulated solar light radiation (λ = 320 ‐ 800 nm) and
NT‐WO3 catalyst. Conditions: pH = 8.3, T = 20 ‐ 40 °C, CEC,0 = 2 mg L‐1, CCAT = 0.5 g L‐1,
CO3,g inlet = 10 mg L‐1, Qg = 20 L h‐1 (O2 or O3/O2).
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-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
CT
OC
/CT
OC
,0
TIME (min)
Dark
Adsorption Photolysis Photocatalytic oxidation Ozonation Catalytic ozonation Photocatalytic ozonation
Figure 7.10. Time evolution of TOC dimensionless concentration during all the
treatments applied to MWW effluent under simulated solar light radiation (λ = 320 ‐ 800
nm) and NT‐WO3 catalyst. Conditions: pH = 8.3, T = 20 ‐ 40 °C, CEC,0 = 2 mg L‐1, CCAT = 0.5
g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 20 L h‐1 (O2 or O3/O2).
Table 7.4. TOC removal (%) after 120 min reaction for the different treatments applied.
TREATMENT No catalysts
Photolysis 4
Ozonation 31
TREATMENT P25 P25‐WO3 NT NT‐WO3
Adsorption 3 17 9 23
Photocatalysis 37 42 31 37
Catalytic
ozonation 23 48 39 47
Photocatalytic
ozonation 47 55 42 64
For the photocatalytic oxidation process with P25 series, as shown in Table
CAPÍTULO 7 (CHAPTER 7)
258
7.3 the incorporation of W did not give place to higher ECs removal, attaining
around 75, 80 and 90 % for IBP, CAF and MTP, respectively, with both P25 and
P25‐WO3 catalysts. In spite of the improvement observed by introducing W
when irradiated with visible light, bare P25 presents a higher catalytic activity
under UVA radiation and thus, if the entire simulated solar light spectrum (320 ‐
800 nm) is used, the beneficial effect of W is not noticeable. Similar results were
observed in terms of mineralization (evolution not shown, final value given in
Table 7.4). On the contrary, NT photocatalyst showed the poorest behavior
during the reaction but a significant increase in the photocatalytic activity of NT‐
WO3 respect to NT precursor was noticed. Thus, in this case ECs conversions
around 70, 80 and 90 % for IBP, CAF and MTP, respectively, were achieved,
similar to those reached in P25 series. In addition, regardless of the catalyst used
the order of reactivity during photocatalytic oxidation was MTP > CAF > IBP,
different to that of the rate constants of the direct reaction between these ECs
and hydroxyl radicals, HO , (kHO∙‐IBP = 7.4x109 M‐1 s‐1 [34], kHO∙‐CAF = 5.9x109 M‐1 s‐1
[35], kHO∙‐MTP = 2.1x109 M‐1 s‐1 [36]). This may be an indicator that HO is not the
only responsible of ECs removal and other different ways such as direct h+
oxidation can be taking place.
All the ozone treatments led to complete ECs depletion in less than 45 min of
reaction time (conversion higher than 99.9 %) regardless of the presence/absence
of catalysts and/or radiation in contrast to photocatalytic oxidation. Both direct
and indirect ozone reactions may take place according to the pH of the reaction
medium (pH = 8.3) that was kept constant throughout the reaction time due to
the buffering capacity of the MWW. On the other hand, slight differences can be
noticed among ozone processes with a different behavior mainly depending on
the rate constant of the direct ozone‐EC reaction (kO3‐IBP = 9.1 M‐1 s‐1 [37], kO3‐CAF =
6.5x102 M‐1 s‐1 [35], kO3‐MTP = 6.2x103 M‐1 s‐1 (pH = 8) [38]). Although photocatalytic
ozonation gave place to a faster ECs depletion rate, it is noticeable in Fig. 7.9 that
the differences between ozone‐based processes are only significant for IBP
whose direct ozone reaction presents the lowest rate constant, being practically
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
ozonation of emerging contaminants in a municipal wastewater treatment plant effluent
259
negligible in the case of CAF and MTP. Main differences between ozone and O3‐
catalytic processes were obtained in terms of TOC removal as shown in Table 7.4
and Fig. 7.10. Ozone alone led to 31 % mineralization mainly during the first
minutes of reaction and then stopped likely due to the formation of compounds
refractory to ozone attack. Production of HO , main responsible species for
mineralization, seems to be insufficient at pH = 8.3 in the ozonation process. The
degree of mineralization decreased in the presence of P25. This material does
not exert any positive effect in catalytic ozonation leading to even lower
mineralization than ozone alone probably due to an inefficient consumption of
O3 onto the catalyst surface. In contrast, NT slightly increased TOC removal up
to 39 % although it can be associated to the higher adsorption capacity of this
material since the mineralization rate is similar to that of the ozonation process
(not shown). On the other hand, W containing photocatalysts, P25‐WO3 and NT‐
WO3, led to higher mineralization (around 47 %), with also higher TOC removal
rate than ozone alone (see Fig. 7.10 for NT‐WO3), thus indicating a catalytic
effect to some extent. This improvement can be related to (1) the catalytic effect
of WO3 and/or (2) the higher adsorption capacity of TiO2‐WO3 catalysts. The
catalytic activity of WO3 in aqueous ozone decomposition, i.e., catalytic
ozonation, has not been extensively reported, but some evidences have been
found. Nishimoto et al. observed a higher TOC removal during phenol
ozonation with WO3 suggesting its role as an ozonation catalyst in a similar
manner to MnO2 or TiO2 [39]. In addition, WO3 has been widely used as the
active component of O3 gas sensors, in which O3 undergoes dissociative
adsorption onto the WO3 surface [40,41]. On the other hand, another plausible
explanation for the catalytic activity of TiO2‐WO3 catalysts during ozonation
compared to bare TiO2 is the enhanced adsorption capacity due to the presence
of WO3 [18,33]. Thus, TiO2 has been widely used as catalyst for ozonation
processes and the proposed mechanisms involve surface reactions between
adsorbed O3 and organic compounds [42]. The fact that bare TiO2 samples
studied here do not show any catalytic activity in the process could be related to
CAPÍTULO 7 (CHAPTER 7)
260
an inefficient decomposition of O3 due to the lack of organic matter near the
catalyst surface. However, the enhanced adsorption capacity of TiO2‐WO3
catalysts can lead to an increase in the local concentration of organic matter in
the vicinity of their surface [18], thus allowing surface reactions between
adsorbed O3‐organic compounds. Both hypotheses should be considered but
some additional work to clarify the prevailing mechanism is needed. Finally, the
highest mineralization degree was obtained with the photocatalytic ozonation
process regardless of the catalyst used. This is in agreement with the expected
synergistic effect between ozone and irradiated semiconductors. This synergistic
effect is due to the reaction of O3 with the conduction band electrons of TiO2 and
WO3, or with the superoxide ion radical generated ( ‐2O ), in both cases leading to
the formation of higher concentrations of hydroxyl radicals ( HO ), main
responsible for mineralization [5,8,39].
On the other hand, MWW parameters after the most representative
treatments with NT‐WO3 catalyst are summarized in Table 7.2. As commented
above, photocatalytic ozonation gave place to higher TOC depletion than
individual treatments. Similar trends are observed both for COD and
aromaticity, reaching a 66 % COD reduction in the combined treatment
compared to 10 % in photocatalytic oxidation and 19 % in single ozonation.
Also, the effectiveness of catalytic ozonation can be observed in terms of TOC,
COD and aromaticity. In addition, it can be noticed that the pH remained
unalterable after the treatments applied as a consequence of the buffer effect of
the IC content (from carbonate/bicarbonate).
The comparison of the catalysts in the photocatalytic ozonation process is
shown with more detail in Figure 7.11 and Figure 7.12 which depict ECs
concentration and TOC with time, respectively.
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
ozonation of emerging contaminants in a municipal wastewater treatment plant effluent
261
-30 0 10 20 30 40 50 600.0
0.2
0.4
0.6
0.8
1.0
Ozonation P25 P25-WO
3
NT NT-WO
3
CIB
P/C
IBP
,0
TIME (min)
Dark
-30 0 10 20 30 40 50 600.0
0.2
0.4
0.6
0.8
1.0
CM
TP/C
MT
P,0
TIME (min)
Dark
-30 0 10 20 30 40 50 600.0
0.2
0.4
0.6
0.8
1.0
CC
AF/C
CA
F,0
TIME (min)
Dark
Figure 7.11. Time evolution of ECs dimensionless concentration during ozonation and
photocatalytic ozonation of MWW effluent under simulated solar light radiation (λ = 320
‐ 800 nm) for all the catalysts studied. Conditions: pH = 8.3, T = 20 ‐ 40 °C, CEC,0 = 2 mg L‐1,
CCAT = 0.5 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 20 L h‐1 (O3/O2).
CAPÍTULO 7 (CHAPTER 7)
262
-20 0 20 40 60 80 100 1200.0
0.2
0.4
0.6
0.8
1.0
Ozonation P25 P25-WO
3
NT NT-WO
3
CT
OC
/CT
OC
,0
TIME (min)
Dark
Figure 7.12. Time evolution of TOC dimensionless concentration during ozonation and
photocatalytic ozonation of MWW effluent under simulated solar light radiation (λ = 320
‐ 800 nm) for all the catalysts studied. Conditions: pH = 8.3, T = 20 ‐ 40 °C, CEC,0 = 2 mg L‐1,
CCAT = 0.5 g L‐1, CO3,g inlet = 10 mg L‐1, Qg = 20 L h‐1 (O3/O2).
Again, it can be noticed that ozone alone was able to completely remove the
ECs in a short time. Differences between ozonation and photocatalytic ozonation
were only observed for IBP and CAF depletion according to the lower rate
constant of their direct ozone reactions. In these cases, indirect and
photocatalytic contributions are more important than for MTP. It can be also
noticed the highest catalytic activity of WO3 composite materials compared to
TiO2 precursors although the main differences were observed in TOC evolution
(Fig. 7.12). In addition to their higher adsorption capacity, the highest
mineralization rate was observed for the TiO2‐WO3 catalysts. These materials
take the advantage of using a greater fraction of solar light radiation in the
visible region due to the presence of WO3 as demonstrated in section 7.3.2, but
also a significant contribution of dark catalytic ozone reactions. The best
performance was attained with NT‐WO3 photocatalyst which led to 64 % TOC
removal. The improvement observed in the NT series after introducing the WO3
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
ozonation of emerging contaminants in a municipal wastewater treatment plant effluent
263
is remarkably higher than in the P25 series, in good agreement with the results
obtained under visible light, probably due to a better dispersion of WO3. To
elucidate the reaction mechanism of photocatalytic ozonation using NT‐WO3
catalyst it is necessary to consider the different processes involved and will be
the subject of future work.
7.4. CONCLUSIONS
TiO2‐WO3 composite catalysts have been synthetized with around 4 wt. % of
WO3 and visible light response, and successfully applied for the removal of
emerging contaminants in urban wastewater and mineralization of the effluent
through photocatalytic ozonation. Ozone alone is able to completely remove the
emerging contaminants in the wastewater matrix but the mineralization degree
reached is relatively low. Photocatalytic ozonation gives place to the highest
mineralization rate regardless of the catalyst used. Titanate nanotubes structure
is not efficient in the photocatalytic oxidation and photocatalytic ozonation
processes, but it is a good precursor of composite catalysts due to its textural
and surface properties. Titanate nanotubes gave place to a composite catalyst
NT‐WO3 with elongated anatase particles, a well‐developed porous structure
and high dispersion of WOX species showing visible light absorption capacity.
The best performance of the NT‐WO3 catalyst compared to bare TiO2 is related
to several mixed contributions such as the use of a greater fraction of solar light
radiation, higher organic compounds adsorption capacity and catalytic activity
in ozone reactions. The contribution of the different phenomena to elucidate the
reaction mechanism with this catalyst will be the subject of future work.
AKNOWLEDGEMENTS
This work has been supported by the Spanish Ministerio de Economía,
Industria y Competitividad (MINECO) and European Feder Funds through the
project CTQ2012‐35789‐C02‐01. Authors acknowledge the SACSS‐SAIUEX and
UAI‐ICP for the characterization analyses. A. Rey thanks the University of
CAPÍTULO 7 (CHAPTER 7)
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Extremadura for a postdoctoral research contract. M.D. Hernández‐Alonso
thanks MINECO for the award of her postpoctoral contract from the “Ramón y
Cajal” program. E. Mena thanks the Consejería de Empleo, Empresa e
Innovación (Gobierno de Extremadura) and European Social Fund for providing
her a predoctoral FPI grant (Ref. PD12059).
REFERENCES
[1] Santos, J.L.; Aparicio, I.; Callejón, M.; Alonso, E. “Occurrence of
pharmaceutically active compounds during 1‐year period in wastewaters from
four wastewater treatment plants in Seville (Spain)”. J. Hazard. Mater. 164 (2009)
1509‐1516.
[2] Barceló, D.; Petrovic, M. (Eds.). “Emerging contaminants from industrial and
municipal wastes. Occurrence, analysis and effects”. The Handbook of
Environmental Chemistry 5‐S1. Springer, Berlin (Germany), 2008.
[3] Halling‐Sorensen, B.; Nors‐Nielsen, S.; Lanzky, P.F.; Ingerslev, F.; Holten
Lützhoft, H.C.; Jorgensen, S.E. “Occurrence, fate and effects of pharmaceutical
substances in the environment‐‐a review”. Chemosphere 36 (1998) 357‐393.
[4] Esplugas, S.; Bila, D.M.; Krause, L.G.T.; Dezotti, M. “Ozonation and
advanced oxidation technologies to remove endocrine disrupting chemicals
(EDCs) and pharmaceuticals and personal care products (PPCPs) in water
effluents”. J. Hazard. Mater. 149 (2007) 631‐642.
[5] Agustina, T.E.; Ang, H.M.; Vareek, V.K. “A review of synergistic effect of
photocatalysis and ozonation on wastewater treatment”. J. Photochem.
Photobiol. C Photochem. Rev. 6 (2005) 264‐273.
[6] Rodríguez, E.M.; Fernández, G.; Alvarez, P.M.; Beltrán, F.J. “TiO2 and Fe (III)
photocatalytic ozonation processes of a mixture of emergent contaminants of
water”. Water Res. 46 (2012) 152‐166.
[7] Rivas, F.J.; Beltrán, F.J.; Encinas, A. “Removal of emergent contaminants:
Integration of ozone and photocatalysis”. J. Environ. Manage. 100 (2012) 10‐15.
[8] Mena, E.; Rey, A.; Acedo, B.; Beltrán, F.J.; Malato, S. “On ozone‐
photocatalysis synergism in black‐light induced reactions: Oxidizing species
production in photocatalytic ozonation versus heterogeneous photocatalysis”.
Chem. Eng. J. 204‐206 (2012) 131‐140.
[9] Hernández‐Alonso, M.D.; Coronado, J.M.; Maira, A.J.; Soria, J.; Loddo, V.;
Augugliaro, V. “Ozone enhanced activity of aqueous titanium dioxide
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
ozonation of emerging contaminants in a municipal wastewater treatment plant effluent
265
suspensions for photocatalytic oxidation of free cyanide ions”. Appl. Catal. B
Environ. 39 (2002) 257‐267.
[10] Malato, S.; Fernández‐Ibáñez, P.; Maldonado, M.I.; Blanco, J.; Gernjak, W.
“Decontamination and disinfection of water by solar photocatalysis: Recent
overview and trends”. Catal. Today 147 (2009) 1‐59.
[11] Hernández‐Alonso, M.D.; Fresno, F.; Suárez, S.; Coronado, J.M.
“Development of alternative photocatalysts to TiO2: Challenges and
opportunities”. Energ. Environ. Sci. 2 (2009) 1231‐1257.
[12] Shifu, C.; Lei, C.; Shen, G.; Gengyu, C. “The preparation of coupled
WO3/TiO2 photocatalyst by ball milling”. Powder Technol. 160 (2005) 198‐202.
[13] Kwon, Y.T.; Song, K.I.; Lee, W.I.; Choi, G.J.; Do, Y.R. “Photocatalytic
behavior of WO3‐loaded TiO2 in an oxidation reaction”. J. Catal. 191 (2000) 192‐
199.
[14] Tryba, B.; Piszcz, M.; Morawski, A.W. “Photocatalytic activity of TiO2‐WO3
composites”. Int. J. Photoenergy (2009) (Article ID 297319) 1‐7.
[15] Ismail, M.; Bousselmi, L.; Zahraa, O. “Photocatalytic behavior of WO3‐
loaded TiO2 systems in the oxidation of salicylic acid”. J. Photochem. Photobiol.
A Chem. 222 (2011) 314‐322.
[16] Xiao, M.; Wang, L.; Huang, X.; Wu, Y.; Dang, Z. “Synthesis and
characterization of WO3/titanate nanotubes nanocomposite with enhanced
photocatalytic properties”. J. Alloys Compd. 470 (2009) 486‐491.
[17] Iliev, V.; Tomova, D.; Rakovsky, S.; Eliyas, A.; Li Puma, G. “Enhancement of
photocatalytic oxidation of oxalic acid by gold modified
WO3/TiO2 photocatalysts under UV and visible light irradiation”. J. Mol. Catal.
A Chem. 327 (2010) 51‐57.
[18] Tomova, D.; Iliev, V.; Rakovsky, S.; Anachkov, M.; Eliyas, A.; Li Puma, G.
“Photocatalytic oxidation of 2,4,6‐trinitrotoluene in the presence of ozone under
irradiation with UV and visible light”. J. Photochem. Photobiol. A Chem. 231
(2012) 1‐8.
[19] Hernández‐Alonso, M.D.; García‐Rodríguez, S.; Sánchez, B.; Coronado, J.M.
“Revisiting the hydrothermal synthesis of titanate nanotubes: new insights on
the key factors affecting the morphology”. Nanoscale 3 (2011) 2233‐2240.
[20] Hernández‐Alonso, M.D.; García‐Rodríguez, S.; Suárez, S.; Portela, R.;
Sánchez, B.; Coronado, J.M. “Highly selective one‐dimensional TiO2‐based
nanostructures for air treatment applications”. Appl. Catal. B Environ. 110
(2011) 251‐259.
CAPÍTULO 7 (CHAPTER 7)
266
[21] Bader, H.; Hoigné, J. “Determination of ozone in water by the indigo
method”. Water Res. 15 (1981) 449‐456.
[22] Sing, K.S.W.; Everett, D.H.; Haul, R.A.W.; Moscou, L.; Pierotti, R.A.;
Rouquerol, J.; Siemieniewska, T. “Reporting physisorption data for gas/solid
systems with special reference to the determination of surface area and
porosity”. Pure Appl. Chem. 57 (1985) 603‐619.
[23] Yan, J.; Wu, G.; Guan, N.; Li, L.; Li, Z.; Cao, X. “Understanding the effect of
surface/bulk defects on the photocatalytic activity of TiO2: anatase versus rutile”.
Phys. Chem. Chem. Phys. 15 (2013) 10978‐10988.
[24] Arabatzis, I.M.; Antonaraki, S.; Stergiopoulos, T.; Hiskia, A.;
Papaconstantinou, E.; Bernard, M.C.; Falaras, P. “Preparation, characterization
and photocatalytic activity of nanocrystalline thin film TiO2 catalysts towards
3,5‐dichlorophenol degradation”. J. Photochem. Photobiol. A Chem. 149 (2002)
237‐245.
[25] Vuurman, M.A.; Wachs, I.E.; Hirt, A.M. “Structural determination of
supported vanadium pentoxide‐tungsten trioxide‐titania catalysts by in situ
Raman spectroscopy and x‐ray photoelectron spectroscopy”. J. Phys. Chem. 95
(1991) 9928‐9937.
[26] Santato, C.; Odziemkowski, M.; Ulmann, M.; Augustynski, J.
“Crystallographically oriented mesoporous WO3 films: Synthesis,
characterization, and applications”. J. Am. Chem. Soc. 123 (2001) 10639‐10649.
[27] Akurati, K.K.; Vital, A.; Dellemann, J.P.; Michalow, K.; Graule, T.; Ferri, D.;
Baiker, A. “Flame‐made WO3/TiO2 nanoparticles: Relation between surface
acidity, structure and photocatalytic activity”. Appl. Catal. B Environ. 79 (2008)
53‐62.
[28] Wagner, C.D.; Davis, L.E.; Zeller, M.V.; Taylor, J.A.; Raymond, R.H.; Gale,
L.H. “Empirical atomic sensitivity factors for quantitative analysis by electron
spectroscopy for chemical analysis”. Surf. Interface. Anal. 3 (1981) 211‐225.
[29] Xiao, M.W.; Wang, L.; Huang, X.J.; Wu, Y.D.; Dang, Z. “Synthesis and
characterization of WO3/titanate nanotubes nanocomposite with enhanced
photocatalytic properties”. J. Alloys Compd. 470 (2009) 486‐491.
[30] Li, Y.; Chen, L.; Guo, Y.; Sun, X.; Wei, Y. “Preparation and characterization
of WO3/TiO2 hollow microsphere composites with catalytic activity in dark”.
Chem. Eng. J. 181‐182 (2012) 734‐739.
[31] Ryan, C.C.; Tan, D.T.; Arnold, W.A. “Direct and indirect photolysis of
sulfamethoxazole and trimethoprim in wastewater treatment plant effluent”.
PAPER 5: WO3‐TiO2 based catalysts for the simulated solar radiation assisted photocatalytic
ozonation of emerging contaminants in a municipal wastewater treatment plant effluent
267
Water Res. 45 (2011) 1280‐1286.
[32] Jacobs, L.E.; Weavers, L.K.; Houtz, E.F.; Chin, Y.P. “Photosensitized
degradation of caffeine: Role of fulvic acids and nitrate”. Chemosphere 86 (2012)
124‐129.
[33] Kwon, Y.T.; Song, K.Y.; Lee, W.I.; Choi, G.J.; Do, Y.R. “Photocatalytic
behavior of WO3‐loaded TiO2 in an oxidation reaction”. J. Catal. 191 (2000) 192‐
199.
[34] Huber, M.M.; Canonica, S.; Park, G.Y.; Von Gunten, U. “Oxidation of
pharmaceuticals during ozonation and advanced oxidation processes”. Environ.
Sci. Technol. 37 (2003) 1016‐1024.
[35] Broséus, R.; Vincent, S.; Aboulfadl, K.; Daneshvar, A.; Sauvé, S.; Barbeau, B.;
Prévost, M. “Ozone oxidation of pharmaceuticals, endocrine disruptors and
pesticides during drinking water treatment”. Water Res. 43 (2009) 4707‐4717.
[36] Sires, I.; Oturan, N.; Oturan, M.A. “Electrochemical degradation of β‐
blockers. Studies on single and multicomponent synthetic aqueous solutions”.
Water Res. 44 (2010) 3109‐3120.
[37] Huber, M.M.; Gobel, A.; Joss, A.; Hermann, N.; Loffler, D.; McArdell, C.;
Ried, A.; Siegrist, H.; Ternes, T.A.; Von Gunten, U. “Oxidation of
pharmaceuticals during ozonation of municipal wastewater effluents: A pilot
study”. Environ. Sci. Technol. 39 (2005) 4290‐4299.
[38] Benítez, F.J.; Acero, J.L.; Real, F.J.; Roldán, G. “Ozonation of pharmaceutical
compounds: Rate constants and elimination in various water matrices”.
Chemosphere 77 (2009) 53‐59.
[39] Nishimoto, S.; Mano, T.; Kameshima, Y.; Miyake, M. “Photocatalytic water
treatment over WO3 under visible light irradiation combined with ozonation”.
Chem. Phys. Letters 500 (2010) 86‐89.
[40] Guerin, J.; Aguir, K.; Bendahan, M. “Modeling of the conduction in a WO3
thin film as ozone sensor”. Sensor. Actuat. B Chem. 119 (2006) 327‐334.
[41] Oison, V.; Saadi, L.; Lambert‐Mauriat, C.; Hayn, R. “Mechanism of CO and
O3 sensing on WO3 surfaces: First principle study”. Sensor. Actuat. B Chem. 160
(2011) 505‐510.
[42] Kasprzyk‐Hordern, B.; Ziólek, M.; Nawrocki, J. “Catalytic ozonation and
methods of enhancing molecular ozone reactions in water treatment”. Appl.
Catal. B Environ. 46 (2003) 639‐669.
CAPÍTULO 8 (CHAPTER 8) PAPER 6: Reaction mechanism and kinetics of DEET visible light assisted photocatalytic ozonation with WO3 catalyst
E. Mena, A. Rey, E.M. Rodríguez, F.J. Beltrán
Applied Catalysis B: Environmental 202 (2017) 460-472
ABSTRACT. This work is focused on the mechanistic investigation of N,N-diethyl-meta-
toluamide (DEET) degradation by photocatalytic ozonation, using WO3 in suspension
and visible radiation (wavelength ≥ 390 nm). This combined process proved to be an efficient treatment to completely remove the contaminant, HO• radicals being the main species involved. Different oxidation products were identified by liquid chromatography time-of-flight mass spectrometry and ion chromatography analyses, and the evolution of their relative abundances with reaction time was established. The efficiency of photocatalytic ozonation treatment was pointed out not only in the DEET depletion rate but also in the evolution of the main intermediate species and mineralization. All the large intermediates initially formed were completely removed within 60 min reaction time and only short-chain organic acids with very low toxicity remained in solution at concentrations in agreement with the mineralization degree achieved (up to 60 % in 2 h). A reaction mechanism of photocatalytic ozonation of DEET involving different chemical reaction steps, with the final formation of short-chain organic acids and mineralization to
CO2, has been proposed. A lumped kinetic model based on TOC and hydroxyl radical
reaction was developed to simulate DEET, intermediates and short-chain organic acids evolution in terms of TOC that provides a simplified approach for this process.
Keywords: Photocatalytic ozonation, DEET, WO3, mechanism, kinetic.
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
271
8.1. INTRODUCTION
Pharmaceutical and personal care products (PPCPs) are emerging
contaminants of increasing environmental concern due to their continuous
discharge into the aquatic environment, persistence and toxic effect in human
and wildlife health. Among them, N,N‐diethyl‐meta‐toluamide (DEET) has been
widely used as the active compound in insect repellents for protection against
mosquito [1]. Due to its extensive use, in the last few years DEET has been
detected in different aquatic systems, wastewater treatment plant effluents and
even in drinking water treated, thus indicating its recalcitrance to conventional
treatments [2,3]. Although DEET is considered to be slightly toxic to aquatic
invertebrates and freshwater fish [2], it has been reported to have potential
carcinogenic properties in human nasal mucosal cells [4]. In addition, ingestion
of low doses of DEET in children has been reported to result in coma and
seizures [5]. Therefore, it is critical to develop a fundamental understanding of
the fate and degradation of DEET during water treatment processes.
DEET degradation through different ozone‐based processes like ozonation
[4], ozone combined with hydrogen peroxide and/or UV radiation [6], and
photocatalytic treatments has been studied [1,3,6‐8], the last being the most
efficient whereas single ozonation resulted to be ineffective for being so slow.
However, the combination of ozone and heterogeneous photocatalysis
(photocatalytic ozonation), has demonstrated to lead to higher mineralization
rates compared to the individual treatments, due to the generation of higher
concentration of oxidizing species, mainly hydroxyl radicals, and also the
inhibition of e‐/h+ pair recombination to some extent [9,10]. Photocatalytic
ozonation of DEET has been previously investigated using TiO2 as photocatalyst
and UV (λ = 254 nm) as radiation source [6]. For the practical deployment of
photocatalytic technologies, the use of natural solar radiation would be an
advantage. In this line, to improve the efficiency of the process, in a previous
work photocatalytic ozonation of DEET was studied using different forms of
WO3, a visible‐light‐responsive semiconductor. According to this work, a
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272
monoclinic WO3 catalyst with W reduced states gave rise to complete DEET
removal and high mineralization degree in less than 2 h [11].
However, it is known that the degradation of DEET during ozonation or
photocatalytic oxidation proceeds through complex redox reactions involving
different steps which can lead to the formation of different intermediates (from
now on transformation products, TPs) if complete mineralization is not achieved
[1,4]. Some of these TPs can be toxic and, in some cases, more persistent than the
parent compound [1]. Thus, the identification of these TPs and their reactivity is
a key step in optimizing the processes conditions in order to increase their
efficiency. To the best of our knowledge, although there are studies dealing with
the identification of TPs generated from DEET during ozonation [4,12] and
heterogeneous photocatalysis with TiO2 [1,8], there are no works focused on the
identification of DEET photocatalytic ozonation TPs. Thus, the purpose of this
work is to study the visible light assisted photocatalytic ozonation of DEET
using WO3 as catalyst with special interest in: 1) identification of DEET TPs and
other intermediates; 2) determination of main species involved in DEET
degradation with the aid of different scavengers; and 3) development of a
kinetic model for the process based on the previous results.
8.2. EXPERIMENTAL SECTION
8.2.1. Experiments
All the experiments were carried out in semi‐batch mode in a 0.5 L effective
volume glass‐made spherical reactor, provided with a gas inlet, a gas outlet and
a liquid sampling port. The reactor was placed in the chamber of a solar
simulator (Suntest CPS, Atlas) provided with a 1500 W Xe lamp with emission
restricted to wavelengths above 390 nm using a polyester cut‐off filter (Edmund
Optics). The spectral irradiance of incident radiation shown in Figure 8.1(A),
was measured with a spectral‐radiometer Black Comet C (StellarNet), provided
with an optic fiber for the wavelength range between 190 ‐ 850 nm. The
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
273
irradiation intensity was set at 550 W m−2 and the temperature increased from 20
‐ 40 °C throughout the experiments. A typical temperature vs time profile is
shown in Figure 8.1(B). An ozone generator (Anseros Ozomat Com AD‐02) was
used to produce a gaseous ozone‐oxygen stream that was fed to the reactor.
200 300 400 500 600 700 8000.0
0.2
0.4
0.6
0.8
1.0
NO
RM
ALI
ZE
D L
IGH
T IN
TE
NS
ITY
AB
SO
RB
AN
CE
(u
.a.)
WAVELENGTH (nm)
0.0
0.2
0.4
0.6
0.8
1.0
0 20 40 60 80 100 12015
20
25
30
35
40
45
T (
º C)
TIME (min)
Figure 8.1. (A) Spectral irradiance of Xe light after passing through the polyester filter
(line) and absorption spectra of DEET (dotted line). (B) Evolution of the temperature of
the reaction media with time at the experimental conditions applied in this work.
CAPÍTULO 8 (CHAPTER 8)
274
In a typical ozonation experiment, the reactor was loaded with 0.5 L of an
aqueous solution containing 15 mg L‐1 of DEET (pH0 = 6) and covered with
aluminum foil when needed (experiments in the dark). Then, the Xe lamp was
switched on and at the same time an ozone‐oxygen mixture (10 mg L‐1 ozone
concentration) was fed to the reactor at a flow rate of 15 L h−1. Photolysis and
photocatalytic oxidation of DEET were also performed in absence of ozone. For
catalytic and photocatalytic experiments the same procedure was followed but
0.125 g of the catalyst were added and the suspension was stirred for 30 min
before switching on the lamp and/or opening the gas flow (O2 or O2/O3), in order
to attain the adsorption equilibrium. The catalyst used was a monoclinic WO3‐
microspheres sample synthetized by sol‐gel procedure and calcined at 600 °C as
reported in a previous work [11]. Adsorption of DEET onto the catalyst was also
tested in absence of ozone in the dark.
On the other hand, in some experiments, appropriate amounts of scavengers
(tert‐butyl alcohol (t‐BuOH) and oxalate) were added to the DEET aqueous
solution in order to determine the nature of the main species involved in DEET
degradation. An additional experiment of photocatalytic ozonation of p‐
chlorobenzoic acid (pCBA, 5 mg L‐1) in the presence of oxalic acid (0.01 M) was
carried out at similar operating conditions for kinetics considerations. In all
cases the time for each oxidation experiment was 2 h. At different intervals
samples were withdrawn from the reactor and filtered through a 0.2 μm PET
membrane.
8.2.2. Analytical methods
DEET concentration was analyzed by HPLC‐DAD (Hitachi, Elite LaChrom)
using a Phenomenex C‐18 column (5 μm, 150 mm long and 3 mm diameter) as
stationary phase and 0.6 mL min‐1 of acetonitrile‐acidified water (0.1 % formic
acid) as mobile phase (30 ‐ 70 v/v, isocratic). Identification and quantification
was carried out at 220 nm. The concentration of pCBA was analyzed in the same
system with the same column and mobile phase at 0.7 mL min‐1, and its
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
275
identification and quantification was carried out at 238 nm. Total organic carbon
(TOC) was measured using a Shimadzu TOC‐VSCH analyzer. Aqueous ozone
concentration was measured by the indigo method using a UV‐Vis
spectrophotometer (Evolution 201, Thermospectronic) set at 600 nm [13]. In this
assay, 3 mL of sample were mixed with 3 mL of indigo solution and then filtered
to remove catalyst particles. Spectrophotometric measurements were
immediately carried out to avoid possible interferences [14]. Concentration of
ozone in the gas phase was continuously monitored by means of an Anseros
Ozomat GM‐6000 Pro analyzer. Hydrogen peroxide concentration was
determined photometrically by the cobalt/bicarbonate method at 260 nm using
the same UV‐Vis spectrophotometer [15].
DEET organic TPs were identified by HPLC‐qTOF using an Agilent 6530
accurate mass quadrupole time‐of‐flight mass spectrometer bearing with
electrospray ionization (ESI) source coupled with an Agilent 1260 series LC
system. A ZORBAX SB‐C18 column (3.5 μm, 150 mm long, and 4.6 mm
diameter) was used as stationary phase. The column was kept at constant
temperature of 30 °C during each analysis. As mobile phase, 0.2 mL min−1 of
acetonitrile‐acidified water (25 mM formic acid) was used from 10 to 100 % of
acetonitrile in 40 min with 15 min of equilibration. The injection volume was 10
μL. The qTOF instrument was operated in the 4 GHz high‐resolution mode. Ions
are generated using an electrospray ion source Dual ESI. Electrospray conditions
were the following: capillary, 3500 V; nebulizer, 30 psi; drying gas, 10 L min‐1;
gas temperature, 350 °C; skimmer voltage, 65 V; octapoleRFPeak, 750 V;
fragmentor, 175 V. The mass axis was calibrated using the mixture provided by
the manufacturer over the m/z 70 ‐ 3200 range. A sprayer with a reference
solution was used as continuous calibration in positive ion using the following
reference masses: m/z 121.0509 and 922.0098. Data were processed using Agilent
Mass Hunter Workstation Software (version B.04.00).
According to previous works, the generation of formaldehyde during DEET
degradation is expected [1,8]. Thus, concentration of formaldehyde in solution
CAPÍTULO 8 (CHAPTER 8)
276
was measured by Hantzsch reaction, measuring the absorbance at 412 nm of the
diacetyldihydrolutidine formed [16]. Finally, short‐chain organic acids and
inorganic ions were analyzed by an ion chromatograph with chemical
suppression (Metrohm 881 Compact Pro) provided with a conductivity detector
using a MetroSep A sup 5 column (250 mm long, 4 mm diameter) at 45 °C and
0.7 mL min‐1 of Na2CO3 from 0.6 ‐ 14.6 mM in 50 min (10 min post‐time for
equilibration) as mobile phase.
8.3. RESULTS AND DISCUSSION
8.3.1. Comparison of processes
In a first series of experiments, the effectiveness of adsorption, photolysis
under λ 390 nm (Ph), ozonation (O3), photolytic ozonation (Ph‐O3),
photocatalytic oxidation (PhC‐O2), catalytic ozonation (C‐O3) and photocatalytic
ozonation (PhC‐O3) on DEET removal was studied. On the one hand, from the
obtained results (not shown) it is deduced that contribution of both adsorption
onto the catalyst and direct photolysis to DEET removal is negligible, the latter
as expected since there is not overlapping between DEET absorption and lamp
emission spectra, as deduced from Fig. 8.1(A). On the other hand, time‐
evolution of normalized DEET and TOC concentration for the different
processes studied is shown in Figures 8.2(A) and 8.2(B), respectively. According
to these figures, degradation rate of DEET during photocatalytic oxidation (PhC‐
O2) was very low. Thus, less than 10 % of DEET conversion was attained after
120 min and no mineralization was observed. Taking into account that WO3 is a
visible light responsive semiconductor, these poor results could be attributable
to the fact that oxygen is not able to react with photoexcited electrons in the
conduction band of WO3 [17], leading to a high recombination rate of e‐/h+ pairs
and, as a consequence, presenting a low efficiency in the photocatalytic
oxidation process.
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visible light assisted photocatalytic ozonation with WO3 catalyst
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Figure 8.2. Evolution of DEET (A) and TOC (B) normalized concentration with time
during the different treatments applied: PhC‐O2; O3; Ph‐O3; C‐O3; PhC‐O3.
Experimental conditions: CDEET,0 = 15 mg L‐1; pH0 = 6; V = 0.5 L; Qg = 15 L h‐1;
*CO3,g = 10 mg L‐1; *I = 550 W m−2; *CWO3 = 0.25 g L‐1 (*if applied).
Regarding single ozonation (O3), at the conditions tested the conversion of
DEET after 2 h was 100 % as shown in Fig. 8.2(A). Given the low reactivity of
ozone towards DEET (kO3 = 0.123 M‐1 s‐1 at 20 °C; [4]), the results suggest the
participation of ozone indirect reactions, namely hydroxyl radical generation
from O3 decomposition, on the elimination of the contaminant. Besides, the
0 30 60 90 1200.0
0.2
0.4
0.6
0.8
1.0
CD
EE
T/C
DE
ET
,0
TIME (min)
(A)
0 30 60 90 1200.0
0.2
0.4
0.6
0.8
1.0
CT
OC/C
TO
C,0
TIME (min)
(B)
CAPÍTULO 8 (CHAPTER 8)
278
relatively low TOC removal achieved (~ 20 % after 2 h, see Fig. 8.2(B)) could be
related to the formation of intermediates refractory to ozone direct reaction [18].
All these aspects are discussed in depth in subsequent sections.
From Fig. 8.2 it can also be seen that addition of WO3 catalyst (C‐O3 system)
had no effect on the effectiveness of the ozonation in terms of DEET and TOC
removal, which indicates the inability of the catalyst to promote an efficient O3
decomposition in the dark.
On the other hand, when light (λ 390 nm) and ozone were combined (Ph‐
O3), DEET degradation rate was higher than the observed for the O3 system in
the dark. In this sense, from Fig. 8.2(A) it is deduced that the time needed to
reach a given DEET conversion by Ph‐O3 was almost half the time needed in the
dark. Besides, in line with the above the mineralization was also higher (Fig.
8.2(B)), reaching a TOC removal of 18 % and 25 % in absence and presence of
light, respectively, after 2 h of treatment. Since the direct photolysis of the
pollutant does not occur, the increased efficiency of the Ph‐O3 system compared
to O3 should be related to the interaction between ozone and light leading to the
formation of reactive species as has been previously reported [19,20].
Undoubtedly, addition of WO3 to the Ph‐O3 system, that is, photocatalytic
ozonation (PhC‐O3) led to the best results among the processes tested. As
observed in Fig. 8.2, after 15 min DEET was completely removed and the
mineralization degree was higher than 60 % after 2 h. These results suggest that,
unlike oxygen, recombination of e‐/h+ generated from WO3 photoexcitation is
avoided thanks to ozone electron trapping [17]. On its role as an electron
acceptor ozone would decompose into reactive oxygen species (mainly •HO )
thus increasing the degradation and mineralization rate of the pollutant. These
results are consistent with the evolution of dissolved ozone shown in Figure 8.3.
As observed, in the presence of light the concentration of ozone in solution
diminishes and more markedly when WO3 is also present. Therefore, it seems
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
279
that both light (λ 390 nm) and WO3 + light promote ozone decomposition
which, in turn, leads to the formation of highly reactive species that would be
responsible for the faster elimination rate of DEET and TOC observed. Another
aspect to highlight from Fig. 8.3 is the time profile of dissolved ozone (O3d)
concentration during single ozonation. As observed, it reaches a maximum at
the beginning of the experiment and then decreases gradually until reaching a
stationary value. This behavior differs from that usually observed during
ozonation of water solutions where O3d concentration is low at the beginning
and gradually increases reaching a stationary value. The explanation to this fact
lies in the temperature profile inside the reactor shown in Fig. 8.1(B). During the
first 60 min, as the temperature increases from ~ 20 °C to 40 °C ozone solubility
decreases. From that moment, since the temperature remains almost constant a
stationary value of ozone in solution is also maintained. In addition to ozone
solubility, temperature can also affect the kinetics of ozone reactions whereas its
effect on the reactivity of transient species is much lower. With this in mind, the
rate constant of the direct reaction between DEET and ozone at 15 °C, 25 °C and
40 °C has been determined by a direct method [18]. Taking into account the low
kO3 values expected according to the literature (0.123 M‐1 s‐1 at 20 °C, [4]), and the
fact that DEET does not dissociate in water, the study was performed at pH 2
(pH adjustment with perchloric acid) in a thermostated bubble column (reaction
volume 250 mL) operating under semi‐batch mode by continuous injection of 15
L h‐1 of a gaseous mixture containing 10 mg L‐1 of O3. In these experiments t‐
BuOH was used as scavenger of •HO radicals that could be generated from O3
decomposition. According to the second‐order rate constants values reported in
the literature (see Table 8.1) for the reaction of both compounds with O3 and
•HO [4,21‐23], the initial concentration of DEET and t‐BuOH was set at 4x10‐5 M
and 4x10‐3 M, respectively. At these conditions, according to Table 8.1, reaction
of DEET with •HO will be avoided whereas t‐BuOH will not be able to compete
with DEET for ozone. Also, the experimental conditions aimed to allow DEET
and ozone to react in the liquid bulk (i.e., slow kinetic regime of ozone
CAPÍTULO 8 (CHAPTER 8)
280
absorption was established) so that the DEET mass balance in the
semicontinuous perfectly mixed reactor used would be given by Eq. (8.1):
3
3d
O ‐DEETDEETO DEET
DEET
kdC‐ C C
dt z (8.1)
where CO3d and CDEET are the concentration of ozone and DEET in the liquid,
respectively; and zDEET the stoichiometry of the reaction between O3 and DEET,
which was considered to be 1 mol of O3 consumed per mol of DEET according to
its low reactivity and the structure of DEET molecule. Integration of Eq. (8.1)
leads to Eq. (8.2) which, in case CO3d remains constant, simplifies to Eq. (8.3):
0
3 3d
0
tDEET
O ‐DEET ODEET
t
CLn k C dt
C (8.2)
0
3 3d
DEETO ‐DEET O
DEET
CLn k C t
C (8.3)
According to Eq. (8.3), a plot of the left term against (CO3dt) should lead to a
straight line that intercepts the origin and whose slope is the value of kO3‐DEET.
As an example, in Figure 8.4 the results obtained at 25 °C are shown. As
observed in Fig. 8.4(A) during the run CO3d was practically constant from t = 5
min (CO3d = 3.03 (± 0.18) x 10‐5 M). On the other hand, from experimental data
fitting to Eq. (8.3) (shown in Fig. 8.4(B)) a kO3‐DEET value of 4.24 ± 0.17 M‐1 s‐1 (R2 =
0.985) was obtained at 25 °C. Following a similar procedure, kO3‐DEET values of
2.46 ± 0.09 M‐1 s‐1 and 7.05 ± 0.33 M‐1 s‐1 were obtained at 15 °C and 40 °C,
respectively. To validate these rate constant values, the slow kinetic regime of
ozone absorption (Hatta number, Ha < 0.3, [18]) was verified taking into account
Eq. (8.4) [4]:
O ‐DEET DEET O3 3
L
k C DHa
k
(8.4)
where DO3 is the diffusivity of ozone in water (1.76x10‐9 m2 s‐1; [24]); and kL the
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
281
individual liquid‐side mass transfer coefficient calculated to be 3.74x10‐4 m s‐1 by
applying the Calderbank’s Equation [25]. At the conditions applied in this work
Ha was always much lower than 0.3, allowing us to validate the kO3‐DEET values
summarized in Table 8.1. As deduced from Table 8.1, although in all cases the
low reactivity of ozone towards DEET is highlighted, kO3‐DEET values obtained in
this work are considerable greater than that reported by Benítez et al. [4]. This
could be attributable to the low CDEET0/Ct‐BuOH0 ratio applied by these authors at
which the alcohol acts as both •HO and ozone scavenger. Finally, the kO3‐DEET
values obtained at different temperatures were fitted to the Arrhenius equation.
The pre‐exponential factor and the activation energy are also indicated in Table
8.1.
0 15 30 45 60 75 90 105 1200
1x10-5
2x10-5
3x10-5
4x10-5
5x10-5
CO
3d (
M)
TIME (min)
Figure 8.3. Evolution of dissolved ozone concentration with time during the application
of different processes. Symbols: O3; Ph‐O3; PhC‐O3. Experimental conditions:
CDEET,0 = 15 mg L‐1; pH0 = 6; V = 0.5 L; Qg = 15 L h‐1; CO3,g = 10 mg L‐1; *I = 550 W m−2;
*CWO3 = 0.25 g L‐1 (*if applied).
CAPÍTULO 8 (CHAPTER 8)
282
0 10 20 30 40 50 600.0
1.0x10-5
2.0x10-5
3.0x10-5
4.0x10-5
5.0x10-5
C (
M)
TIME (min)
(A)
0.00 0.02 0.04 0.06 0.080.0
0.1
0.2
0.3
0.4
ln(C
DE
ET
,0/C
DE
ET)
CO3d·t (M·s)
Slope = 4.24 +/- 0.17
R2 = 0.985
(B)
Figure 8.4. Determination of kO3‐DEET at 25 °C. (A) Evolution of DEET () and O3d ()
concentration with time. (B) Fitting of data to Eq. (3). Experimental conditions:
CDEET,0 = 10 mg L‐1; Ct‐BuOH = 0.004 M; pH0 = 2; V = 0.25 L; Qg = 15 L h‐1; CO3,g = 10 mg L‐1.
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
283
Table 8.1. Second‐order rate constants for the reaction of ozone and hydroxyl radicals
with DEET, t‐BuOH, oxalate and p‐CBA.
Compound kO3 (M‐1 s‐1) Ref. kHO∙ (M‐1 s‐1) Ref.
DEET
20 °C: 0.121 ± 0.005
Conditions: pH = 2 ‐ 9;
CDEET,0 = 10‐6 M; Ct‐BuOH,0 = 0.1 M
[4]
4.95x109
[21]
15 °C: 2.46 ± 0.09 (R2 = 0.991)
25 °C: 4.24 ± 0.17 (R2 = 0.985)
40 °C: 7.05 ± 0.33 (R2 = 0.987)
kO3 = Ae‐Ea/RT; A = 1.2x106;
Ea = 31.3 kJ mol‐1 (R2 = 0.988)
Conditions: pH = 2, CDEET,0 = 4x10‐5 M;
Ct‐BuOH,0 = 4x10‐3 M
This
work
t‐BuOH 0.0011 [22] 6.2x108 [23]
Oxalate < 0.04 [30]
5x107
(pH = 2.8)
7.7x106 (pH = 6)
[31]
p‐CBA 0.015 [39] 5x109 [39]
8.3.2. Determination of the main species responsible for DEET degradation
and mineralization
In order to determine the nature of the species involved on the elimination of
DEET through the most effective systems, namely ozonation, photolytic
ozonation and photocatalytic ozonation, a new series of experiments was carried
out by adding substances capable of acting as scavengers of some of the species
whose generation and participation is expected. In this sense, t‐BuOH has been
selected as scavenger of •HO in the liquid bulk, and oxalate as scavenger of •
HO
both in the liquid bulk and also at the catalyst surface where, in addition,
adsorbed oxalate can be oxidized by the photogenerated positive holes (h+) [26‐
CAPÍTULO 8 (CHAPTER 8)
284
29]. As commented in the previous section, to select the appropriate substances
to be used as scavengers their reactivity towards the different species and
reagents involved must be firstly considered. Thus, in Table 8.1 the rate constant
of the reactions oxalate‐ozone and oxalate‐ •HO are also indicated [30,31]. Taking
into account the reactivity of DEET and the scavengers towards the different
species involved, this study was carried out at the following conditions: pH0 = 6;
20 ‐ 40 °C (see Fig. 8.1(B)); CDEET,0 = 8x10‐5 M; COxal,0 = 10‐3 M and/or Ct‐BuOH,0 = 0.03
M. Figure 8.5 shows the influence of the presence/absence of these scavengers in
DEET degradation rate by the different processes applied: O3 (Fig. 8.5(A)), Ph‐O3
(Fig. 8.5(B)) y PhC‐O3 (Fig. 8.5(C)). The theoretical DEET evolution with time has
also been calculated from Eq. (8.1) considering that DEET was only degraded by
direct ozone attack. For this purpose CDEET,0 value, the evolution of both CO3d
(see Fig. 8.3) and temperature (Fig. 8.1(B)) with time as well as the influence of T
on kO3‐DEET according to Arrhenius equation (see Table 8.1) have been taken into
account. These results have also been added to Fig. 8.5 for comparative reasons.
Fig. 8.5(A) shows the effect of the presence of 0.03 M t‐BuOH in 8x10‐5 M
DEET elimination rate by single ozonation. At these conditions, t‐BuOH will
react with all •HO whereas its reaction with O3 will be negligible. As observed in
Fig. 8.5(A), the presence of t‐BuOH negatively affect the elimination rate of the
contaminant, thus indicating not only the decomposition of ozone into •HO
radicals at the conditions tested but also the high contribution of •HO to DEET
degradation by single ozonation. Moreover, the good agreement between CDEET
evolution when t‐BuOH was present and that calculated by Eq. (8.1) indicates
that O3 and •HO are the only species involved on DEET degradation by single
ozonation.
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visible light assisted photocatalytic ozonation with WO3 catalyst
285
0 15 30 45 60 75 90 105 1200
2x10-5
4x10-5
6x10-5
8x10-5
1x10-4
CD
EE
T (
M)
TIME (min)
(A)
0 15 30 45 60 75 90 105 1200
2x10-5
4x10-5
6x10-5
8x10-5
1x10-4
(B)
CD
EE
T (
M)
TIME (min)
0 15 30 45 60 75 90 105 1200
2x10-5
4x10-5
6x10-5
8x10-5
1x10-4
(C)
CD
EE
T (
M)
TIME (min) Figure 8.5. Influence of the presence of different scavengers on the elimination of DEET
by O3 (A), Ph‐O3 (B) and PhC‐O3 (C). Symbols: no scavenger; t‐BuOH 0.03 M;
t‐BuOH 0.03M and Oxalate 10‐3 M. Line: Calculated from Eq. (1) assuming only direct
ozone‐DEET reaction. Experimental conditions: CDEET,0 = 15 mg L‐1; pH0 = 6; V = 0.5 L;
Qg = 15 L h‐1; CO3,g = 10 mg L‐1; *I = 550 W m−2; *CWO3 = 0.25 g L‐1 (*if applied).
CAPÍTULO 8 (CHAPTER 8)
286
Similarly, the influence of 0.03 M t‐BuOH in the degradation of 8x10‐5 M
DEET by photolytic ozonation is depicted in Fig. 8.5(B). Again, the presence of
t‐BuOH negatively affects the elimination rate of the contaminant. However, the
elimination rate of DEET calculated from Eq. (8.1) (straight line) is clearly lower
than that experimentally determined in the presence of t‐BuOH. Since it is
expected that at the conditions used most of the •HO was trapped by t‐BuOH,
there must be other species and/or mechanism different than •HO and O3 also
contributing to DEET degradation. This effect is more marked during the first 45
min and then, the rate of DEET depletion becomes similar (see Fig. 8.5(B), Eq.
(8.1) and t‐BuOH experiment). The fact that the positive effect of the presence of
light on TOC removal is much lower than on DEET (Fig. 8.2), that is, the effect is
more noticeable at short reaction times, could indicate the formation of an
intermediate capable of acting as a photosensitizer giving rise to the formation
of reactive species. Once the intermediate is degraded the beneficial effect of
light would disappear.
Finally, the effect of the presence of oxalate and/or t‐BuOH on the rate of
DEET removal by photocatalytic ozonation is shown in Fig. 8.5(C). It can be
observed that the degradation rate of DEET in the presence of t‐BuOH or
t‐BuOH + oxalate was very similar. Considering that t‐BuOH only reacts with
•HO in the bulk, these results are pointing out that •
HO and h+ photogenerated
at the catalyst surface do not contribute to DEET degradation. Moreover, it can
be observed that 100 % DEET removal was attained in less than 15 min by PhC‐
O3 when no scavengers are used, but the evolution of DEET during the same
period when t‐BuOH was present was virtually coincident with that calculated
when only its direct reaction with ozone is considered (Eq. (8.1)). Thus, •HO in
the bulk seems to be the only species involved in DEET elimination by PhC‐O3.
Therefore, as deduced from results in Fig. 8.5, at the conditions used in this
work, during photocatalytic ozonation DEET is mainly removed through its
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
287
reaction with •HO radicals in the bulk whose formation when visible radiation,
WO3 catalyst and O3 are combined (PhC‐O3) is highly enhanced. The highest
•HO concentration would also be the responsible for the highest efficiency of this
combination in terms of mineralization (Fig. 8.2(B)). In other words, the above
results confirm that not only DEET but also final products oxidation (mainly
carboxylic acids) proceeds through •HO attack during PhC‐O3 process.
Moreover, first TPs degradation is also likely due to •HO reaction in the bulk as
discussed in the following section.
8.3.3. Identification of TPs
The identification of TPs during ozonation and photocatalytic ozonation was
carried out by means of LC‐qTOF and ion chromatography. LC‐qTOF approach
provided accurate mass measurements of ions (m/z values) of the different
compounds formed, together with several peaks corresponding to equal m/z
values. In total, 22 TPs were identified, which are indicated in Table 8.2 together
with the parent compound DEET. They were designated by a number
corresponding to the m/z value and for isomers the number is followed by a
letter in alphabetical order with increasing retention time. The low experimental
relative mass errors obtained indicate the high grade of confidence in the
assignment of the elemental composition. In this sense, relative mass errors
below 5 ppm are generally accepted for the verification of the elemental
composition [32]. Table 8.2 also includes the tentative structures proposed for
the different TPs according to the specific literature [1,4,8,12,21,33,34]. These TPs
identified, similar to those found in previous works for different systems, are
consistent with the attack of •HO generated throughout the processes. Although
neither the tentative structures nor the concentration of the TPs could be
confirmed due to the lack of standards, reaction pathways for DEET
degradation have been proposed and are discussed in the following section.
In general, main first intermediates detected were large molecules, such as
CAPÍTULO 8 (CHAPTER 8)
288
C12H15NO2 (m/z = 206), C12H17NO2 (m/z = 208), C10H15NO2 (m/z = 182),
C10H15NO3 (m/z = 198) or C10H13NO (m/z = 164), in which practically the parent
compound structure (C12H17NO, m/z = 192) is still present with slight
modifications. Other smaller intermediates like C5H11NO2 (m/z = 118) or C4H11N
(m/z = 74) that have been identified are formed after the cleavage of
characteristic bonds.
Table 8.2. TPs identified by LC‐qTOF during DEET O3 and PhC‐O3 degradation.
Compound tR
(min)
Molecular
Formula
[M]
Experimental
m/z
[M‐H+]
Error
(ppm) Tentative structure
192 (DEET) 24.56 C12H17NO 192.1385 ‐2.65
240‐A 16.05 240.1234 ‐10.41
240‐B 16.92 C12H17NO4 240.1239 ‐8.32
240‐C 17.4 240.1245 ‐5.83
226‐A 15.38
C11H15NO4
226.1083 ‐4.05
226‐B 15.83 226.1085 ‐4.93
222‐A 18.88
C12H15NO3
222.1128 ‐3.29
222‐B 20.13 222.1129 ‐3.85
214 15.96 C10H15NO4 214.1072 ‐4.74
208‐A 17.54
C12H17NO2
208.1337 ‐2.38
208‐B 19.24 208.1339 ‐3.34
208‐C 20.64 208.1326 2.91
208‐D 21.69 208.1338 ‐2.86
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
289
Table 8.2 (Continuation). TPs identified by LC‐qTOF during DEET O3 and PhC‐O3
degradation.
Compound tR
(min)
Molecular
Formula
[M]
Experimental
m/z
[M‐H+]
Error
(ppm) Tentative structure
206 21.03 C12H15NO2 206.1177 ‐4.58
198‐A 13.69 198.1129 ‐2.17
198‐B 15.78 C10H15NO3 198.1129 ‐2.17
198‐C 16.83 198.1130 ‐2.68
182 22.38 C10H15NO2 182.1177 ‐5.19
178 22.70 C11H15NO 178.1229 ‐1.46
164 21.48 C10H13NO 164.1069 ‐2.50
146 9.23 C6H11NO3 146.0812 ‐2.26
118 6.55 C5H11NO2 118.0863 ‐2.42
74 6.62 C4H11N 74.0967 ‐6.74
The evolution with time of the intensity of the signal of DEET and those TPs
detected by LC‐qTOF during O3 and PhC‐O3 processes is shown in Figure 8.6.
According to Fig. 8.6 TPs formed were exactly the same regardless of the process
applied although their formation rate was noticeably much faster during the
photocatalytic ozonation treatment. Thus, during PhC‐O3 total TPs
disappearance was observed at 60 min whereas for single ozonation some of
N
O
C N
O
HO
O
CAPÍTULO 8 (CHAPTER 8)
290
these TPs still remained after 120 min.
0 15 30 45 60 75 90 105 1200
1x106
2x106
3x106
4x106
5x106
6x106
7x106P
EA
K A
RE
A
TIME (min)
0
1x108
2x108
3x108
4x108
5x108
6x108
(A)
0 15 30 45 60 75 90 105 1200
1x106
2x106
3x106
4x106
5x106
6x106
7x106
PE
AK
AR
EA
TIME (min)
(B)
0
1x108
2x108
3x108
4x108
5x108
6x108
Figure 8.6. Time profiles of the intensity of the main peaks detected during (A) O3 and (B)
PhC‐O3 of DEET. Experimental conditions: CDEET,0 = 15 mg L‐1; pH0 = 6; V = 0.5 L;
Qg = 15 L h‐1; CO3,g = 10 mg L‐1; *I = 550 W m−2; *CWO3 = 0.25 g L‐1 (*if applied).
As deduced in the previous section, the fact that •HO radicals in the bulk are
the main species involved in DEET degradation by O3 and PhC‐O3 could
explain, on the one hand, that TPs formed are the same in both processes and,
206 164 182 198-A 198-B 198-C 214 74 146 178
192 (DEET)
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visible light assisted photocatalytic ozonation with WO3 catalyst
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on the other hand, that TPs formation rate in the photocatalytic treatment
presents the highest values as a result of the largest concentration of •HO in the
reaction medium. Moreover, once the maximum concentration of each TP is
reached, the removal rate during the PhC‐O3 process is also faster than the
observed during O3 alone process. In summary, during single ozonation direct
O3 reactions are favored as a consequence of the higher concentration of
dissolved molecular ozone and, according to Fig. 8.6, a higher removal rate of
TPs is clearly obtained by PhC‐O3 system. Then, it can be pointed out that, as for
DEET, in both treatments (O3 and PhC‐O3) the •HO radical is the main
responsible species for the degradation of intermediate TPs.
Further oxidation of these TPs usually gives place to the formation of short‐
chain saturated carboxylic acids. Thus, the presence of oxalic, acetic, pyruvic and
formic acids (see molecular formula and structures in Table 8.3) was detected at
long reaction times for both O3 and PhC‐O3 processes. In Figure 8.7, the
evolution with time of CTOC present as carboxylic acids (ΣCTOC‐acids) is shown. As
expected, a higher generation rate and concentration in solution of these acids
for PhC‐O3 is observed being 18 % and 64 % of residual TOC in the form of these
organic acids after 2 h of ozonation and photocatalytic ozonation, respectively.
By comparing Fig. 8.6 and Fig. 8.7, it is deduced that formation of carboxylic
acids needs the previous oxidation of TPs, which is faster for photocatalytic
ozonation as commented before. Hence, the low TOC conversion into carboxylic
acids after 2 h of ozonation indicates that at these conditions intermediate TPs,
that could even be more harmful than the parent compound, are still present.
During photocatalytic ozonation, only short‐chain organic acids were detected
at the end of the treatment, with the complete elimination of the large TPs. Thus,
according to the literature [1,4], a significant decrease in the toxicity is expected
if this process is applied until reaching a high mineralization degree. However,
the information currently available about ecotoxicity is not sufficient and further
work would be necessary [2].
CAPÍTULO 8 (CHAPTER 8)
292
Table 8.3. Short‐chain saturated organic acids identified during DEET O3 and PhC‐O3
degradation.
Compound Molecular
Formula
Structure
Oxalic acid C2H2O4
Acetic acid C2H4O2
Pyruvic acid C3H4O3
Formic acid CH2O2
0 15 30 45 60 75 90 105 1200
1
2
3
4
CT
OC
-aci
ds(m
g L-1
)
TIME (min)
Figure 8.7. Evolution of TOC corresponding to short‐chain saturated organic acids
concentration with time during different processes. Symbols: O3; PhC‐O3.
Experimental conditions: CDEET,0 = 15 mg L‐1; pH0 = 6; V = 0.5 L; Qg = 15 L h‐1;
CO3,g = 10 mg L‐1; *I = 550 W m−2; *CWO3 = 0.25 g L‐1 (*if applied).
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0 15 30 45 60 75 90 105 1200
2x10-5
4x10-5
6x10-5
8x10-5
C (
M)
TIME (min)
Figure 8.8. Evolution of DEET eliminated and H2O2 and CH2O generated with time
during DEET degradation by O3 (circles) and PhC‐O3 (triangles) systems. Symbols: ,
DEET eliminated; , H2O2 formed; , CH2O generated. Experimental conditions:
CDEET,0 = 15 mg L‐1; pH0 = 6; V = 0.5 L ; Qg = 15 L h‐1; CO3,g = 10 mg L‐1; *I = 550 W m−2;
*CWO3 = 0.25 g L‐1 (*if applied).
According to the literature, ozonation and photocatalytic oxidation of DEET
led to the generation of formaldehyde (CH2O) [1,8]; and also, hydrogen peroxide
(H2O2) is commonly formed through direct ozone and hydroxyl radical reactions
[35,36]. Thus, the evolution of the concentration of both compounds with time is
depicted in Figure 8.8 together with the evolution of DEET eliminated during O3
and PhC‐O3. It can be clearly seen for both treatments that the disappearance of
1 mol of DEET leads to the formation and accumulation of ~1 mol of H2O2 in the
reaction medium. Particularly, during PhC‐O3 an increase of H2O2 concentration
is produced up to 45 min coinciding with the time when TPs were almost
completely removed (see Fig. 8.6(B)). From this moment, H2O2 is gradually
consumed likely in photocatalytic reactions as discussed in a previous work [11].
However, in case H2O2 reactions led to the formation of reactive oxidizing
species, they do not seem to participate in TPs removal. In fact, the analysis of
H2O2 evolution during DEET degradation by the other processes applied in this
CAPÍTULO 8 (CHAPTER 8)
294
work, together with the results during photocatalytic oxidation of DEET adding
H2O2 (PhC‐O2 + H2O2, not shown) led us to conclude that: 1) H2O2
decomposition during the processes applied mainly occurs due to its interaction
with the irradiated catalyst surface; and 2) H2O2 decomposition does not lead to
the generation of reactive species capable of oxidizing DEET and/or its TPs.
Regarding formaldehyde, a direct relation between DEET degradation and
formaldehyde formation is also observed although in this case the ratio is ~ 0.25
mol of CH2O per mol of DEET eliminated by O3 and PhC‐O3.
8.3.4. Proposed reaction mechanism
Bearing in mind the identified TPs and the fact that hydroxyl radical seems to
be the main species involved on DEET degradation by PhC‐O3 when WO3 and
light of wavelength higher than 390 nm are used, a tentative mechanism is
presented in Schemes 8.1 ‐ 8.3 and discussed below. Both the aliphatic chain and
the aromatic ring of DEET molecule are found to react with hydroxyl radicals
and pathways will be separately discussed. For the reactions at the aliphatic
chain of DEET, the identified TPs and the proposed mechanisms are presented
in Scheme 8.1. In this mechanism, the •HO radical attack on the aliphatic chain
could occur by abstraction of hydrogen to form an organic radical in which
several TPs are based on [12]. For example, the TP named 206 in Table 8.2 with
m/z 206.1177 was generated by a reaction between this organic radical and
oxygen or ozone. On the other hand, the rearrangement of the radical formed
and an additional •HO attack at the aliphatic chain lead to its de‐ethylation to
form another radical, which could further react with water to form the TP
named 164 (m/z 164.1069). Other authors have also reported the existence of
these TPs through the above two different pathways during ozonation [4], and
photocatalytic oxidation of DEET [1,8]. In addition, the rearrangement of the
organic radical could lead to the cleavage of characteristic bonds [12] and
subsequent formation of carbanions which could explain the formation of the
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visible light assisted photocatalytic ozonation with WO3 catalyst
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smaller TPs detected in the present work. On the one hand, the rearrangement
of the radical produced a new radical in addition to the carbanion of
diethylamine. From the •HO radical attack on this carbanion and reaction with
water, diethylamine can be formed (TP named 74 in Table 8.2 with m/z 74.0967).
On the other hand, the rearrangement of the radical can also form the benzyl
radical (involved in further reactions) and a carbanion which could react with
•HO radicals to produce hydroxydiethylamide (TP 118 in Table 8.2 with m/z
118.0863). Finally, in Scheme 8.1 is also considered the initial •HO attack and
hydrogen abstraction on the methyl group that is directly attached to the
aromatic ring of DEET. The TP 178 with m/z 178.1229 is derived by detachment
of this methyl group from DEET [1,8], through loss of formaldehyde after a new
•HO radical attack.
Regarding the reactions on the aromatic ring of DEET (see Scheme 8.2), the
initial attack of the hydroxyl radical could lead to the generation of
hydroxylated DEET derivatives. Thus, in the present study four isomers have
been identified for monohydroxy DEET (208‐A,B,C,D in Table 8.2 with m/z
208.1326 ‐ 208.1339) corresponding to the different positions that can be
occupied by the •HO moiety in the aromatic ring, as it has been reported for the
degradation of DEET by anodic Fenton, photolytic processes or photocatalytic
treatments [8,21,33,34]. Monohydroxy DEET further reacts with •HO to form
dihydroxylated and trihydroxylated DEET. In this sense, three different isomers
of trihydroxy DEET have been detected (240‐A,B,C in Table 8.2 with m/z
240.1234 ‐ 240.1245). In addition, hydroxylated derivatives of DEET degradation
products were detected. For example, the isomers 222‐A,B in Table 8.2 with m/z
222.1128 ‐ 222.1129 could be formed either by hydroxylation of the 206 TP, or
from isomers 208‐A,B,C,D through oxidation of the aliphatic chain. In the same
way, the isomers 226‐A,B (m/z 226.1083 ‐ 226.1085) can be generated by
hydroxylation of the 178 TP and also from isomers 240‐A,B,C by cleavage of the
C‐C bond and subsequent detachment of the methyl group from the aromatic
CAPÍTULO 8 (CHAPTER 8)
296
ring. These were also identified in previous works [1,4,8].
Besides, in Scheme 8.2, some of the identified TPs resulted from the opening
of the aromatic ring due to its breakdown through the •HO radical consecutive
attack on the same carbon atom [8]. Thus, the TP named 182 with m/z 182.1177
was assigned to mono‐oxygenated ring opening TP. After the double attack of
•HO on DEET and subsequent ring opening, the decarboxylation involving the
loss of CO2 and H2O would produce TP 182 [37] whose mono‐ and
dihydroxylation could lead to the formation of TPs 198‐A,B,C and 214 (with m/z
198.1129 ‐ 198.1130 and 214.1072), respectively. On the other hand, ozonolysis of
TP 182 could lead to the formation of the TP 146 (m/z 146.0812) through loss of
hydrogen peroxide and further •HO radical attack [18].
Finally, some of the TPs generated could react with ozone and/or hydroxyl
radicals to produce short‐chain saturated carboxylic acids (oxalic, acetic, pyruvic
and formic acids have been detected) which, given their low reactivity towards
ozone, eventually would evolve to CO2 and H2O by reaction with •HO (see
Scheme 8.3).
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visible light assisted photocatalytic ozonation with WO3 catalyst
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Scheme 8.1. Proposed DEET degradation pathway through •
HO attack on the aliphatic
chain.
CAPÍTULO 8 (CHAPTER 8)
298
Scheme 8.2. Proposed DEET degradation pathway through the •
HO attack on the
aromatic ring.
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visible light assisted photocatalytic ozonation with WO3 catalyst
299
Scheme 8.3. Formation of saturated carboxylic acids and mineralization of DEET.
8.3.5. Kinetic study
According to the above results, during photocatalytic ozonation DEET
degradation and mineralization is mainly due to •HO radical reactions without a
significant direct participation of neither ozone nor positive holes generated on
the irradiated photocatalyst surface. Thus, the mechanism of generation of
hydroxyl radicals can be described by the following reactions and mass transfer
steps:
Ozone mass transfer from the gas flow to the liquid phase:
‐3 ‐1
Lk a 5.1x10 s3g 3O O
(8.5)
where kLa is the volumetric mass‐transfer coefficient that was experimentally
determined through O3 absorption experiments for the reaction system used in
this work according to Beltrán in ref. [18].
Ozone dark decomposition:
‐1 ‐1
d1k 70 M s‐ ‐
3 2 2O HO HO O
(8.6)
6 ‐1 ‐1
d2k 2.2x10 M s ‐‐
2 3 2 3HO O HO O HO
(8.7)
‐pKa 11.4
2 2 2H O H HO (8.8)
O3/ HO● HO●
CAPÍTULO 8 (CHAPTER 8)
300
••pKa 4.8 ‐
2 2HO O H (8.9)
9 ‐1 ‐1
d3k 1.6x10 M s‐ ‐
3 32O O O HO
(8.10)
Photocatalytic reactions:
WO3h , ‐
3WO e h (8.11)
10 1 ‐1
d4k 3.6x10 M s‐ ‐
3 3e O O HO
(8.12)
rk‐3e h WO (8.13)
In this mechanism, ozonide radical ( 3O ) readily evolves to form hydroxyl
radical in reactions (8.7), (8.10) and (8.12); oxygen is not able to react with
photogenerated electrons according to [17], so this reaction has not been taken
into account. In addition, although H2O2 is decomposed onto irradiated WO3
surface, the results obtained in terms of DEET degradation and mineralization in
experiments conducted by combining PhC‐O2 + H2O2 (not shown) indicate that
no reactive species are formed during H2O2 photocatalytic decomposition.
Hydroxyl radicals may react with dissolved ozone and hydrogen peroxide
generated according to reactions (8.14) and (8.15):
9 ‐1 ‐1
d5k 3.0x10 M s
3 2 2O HO HO O
(8.14)
7 ‐1 ‐1
d6k 2.7x10 M s
2 2 2 2H O HO HO H O
(8.15)
All the known rate constants have been previously summarized in [18]. The
rate constant of ozone‐electrons reaction has been taken for hydrated electrons
from Buxton et al. [31].
In the presence of any organic contaminant at sufficient concentration,
hydroxyl radicals will mainly react with it. In terms of TOC, the reaction
mechanism of DEET and TOC removal can be described according to the
following lumped reactions:
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visible light assisted photocatalytic ozonation with WO3 catalyst
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HO ‐TOC1
k
1 2TOC HO TOC (8.16)
HO ‐TOC 2k
2 3 2TOC HO TOC 1 ‐ CO (8.17)
HO ‐TOC 3
k
3 2TOC HO CO
(8.18)
where TOC1 represents DEET expressed in carbon units; TOC3 is the total
organic carbon corresponding to the sum of short‐chain saturated organic acids
and formaldehyde in solution detected during photocatalytic ozonation; and
TOC2 is the total organic carbon corresponding to the sum of the rest of DEET
intermediates detected by LC‐qTOF. In reaction (8.17), the stoichiometric
parameter α represents the TOC2 fraction that is converted to TOC3 whereas
(1‐α) is mineralized to CO2.
TOC2 can be calculated from the TOC balance:
t 1 2 3TOC TOC TOC TOC (8.19)
where TOCt is the analyzed total organic carbon in solution at a given time, t.
Mass balances of each species can be established as follows:
Ozone in the gas phase:
3g 3g
3gi 3g 3d
O O
g O g O L O
C R T dCC ‐ C ‐ V k a ‐C 1 ‐ V
He dt
(8.20)
where g is the gas flow rate; CO3gi and CO3g are the molar ozone concentrations
in the gas phase at the reactor inlet and outlet, respectively; CO3d is the molar
concentration of dissolved ozone; β is the liquid holdup; V the reaction volume;
R the ideal gas constant; He is the Henry law constant for the ozone‐water
system at the temperature T.
CAPÍTULO 8 (CHAPTER 8)
302
Ozone in the liquid phase:
3g3d
‐3d 4 3d 5 3d
L OOL O d O d Oe HO
k a C R TdC‐ k a C ‐ k C C ‐ k C C
dt He
(8.21)
where Ce‐ is the concentration of photogenerated electrons; CHO∙ the
concentration of hydroxyl radicals; and kd4 and kd5 the rate constants of
reactions (8.12) and (8.14), respectively. In this balance, the contributions of
reactions (8.6), (8.7) and (8.10) have been neglected according to the pH of the
reaction medium (from 6 to 4).
Photogenerated electrons:
‐
3 4 3d
neWO a d O re h e
dCI k C C k C C
dt (8.22)
where WO3 is the apparent quantum yield of WO3 at the wavelength range used
in this work; Ia is the absorbed radiation flux by the catalyst WO3; the exponent
n is the order of the reaction with respect to the Ia and depends on the efficiency
of e‐/h+ formation and recombination at the catalyst surface taking a value
between 0.5 and 1 when the reaction is kinetically controlled [38]; Ch+ is the
concentration of photogenerated holes and kr is the rate constant of e‐/h+
recombination reaction (8.13). If ozone concentration is high enough, the
recombination process is minimized so it can be considered that all the
photogenerated electrons are trapped by dissolved ozone. Then Eq. (8.22) can be
simplified as:
‐
3 4 3d
neWO a d O e
dCI k C C
dt (8.23)
Hydroxyl radicals:
4 3d 5 3d 6 2 2
1 21 2
33
HOd O d O d H Oe HO HO
TOC TOCHO TOC HO HO TOC HO
TOCHO TOC HO
dCk C C k C C k C C
dtk C C k C C
k C C
(8.24)
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where CH2O2 is the concentration of hydrogen peroxide in the liquid phase;
CTOC1, CTOC2 and CTOC3 are the concentrations corresponding to TOC1, TOC2 and
TOC3, respectively; and kd6, kHO∙‐TOC1, kHO∙‐TOC2, kHO∙‐TOC3 are the rate constants of
reactions (8.15), (8.16), (8.17) and (8.18), respectively.
Eq. (8.24) can be rewritten as follows:
4 3d
HOd O i ie HO i
dCk C C C k C
dt
(8.25)
where the subscript i represents any species that reacts with hydroxyl radicals.
For Eq. (8.23) and (8.25), net reaction rates were assumed to be zero
according to the hypothesis of the stationary state for transient species. Thus,
hydroxyl radical concentration can be expressed as follows:
nWO a3 i
HOsi ii
I rC
kk C
(8.26)
In Eq. (8.26) the numerator ri represents the reaction rate of initiation for the
formation of hydroxyl radicals, whereas the denominator kS represents the
scavenging factor, taking into account that the species involved in this term are
only the inhibitors of the chain mechanism of ozone decomposition into •HO .
Total organic carbon mass balances:
1
1 11 1
TOC iTOC TOCHO TOC HO HO TOC
s
dC rk C C k C
dt k (8.27)
2
1 21 2
1 21 2
TOC
TOC TOCHO TOC HO HO TOC HO
iTOC TOCHO TOC HO TOC
s
dCk C C k C C
dt
rk C k C
k
(8.28)
CAPÍTULO 8 (CHAPTER 8)
304
3
2 32 3
2 32 3
TOC
TOC TOCHO TOC HO HO TOC HO
iTOC TOCHO TOC HO TOC
s
dCk C C k C C
dt
rk C k C
k
(8.29)
where kHO∙‐TOC1 is the rate constant of DEET‐ •HO reaction; kHO∙‐TOC2 represents
the apparent rate constant of the reaction between TOC2 (DEET intermediates
detected by LC‐qTOF) and •HO radicals; and kHO∙‐TOC3 represent an apparent
rate constant of the reaction between compounds of TOC3 (formaldehyde and
oxalic, acetic, pyruvic and formic acids) and •HO .
Firstly, the hypothesis of working in excess of ozone was proved by means of
two different experiments with different ozone concentration at the reactor inlet
(10 and 20 mg L‐1). Results of DEET degradation and mineralization are
represented in Figure 8.9(A) where it can be noticed the similar evolution of
DEET and TOC regardless of the ozone concentration used.
Thus, to solve the proposed model for TOC evolution, the initiation rate ri
was determined in an experiment of photocatalytic ozonation of pCBA in the
presence of oxalic acid. In this system, pCBA is a •HO probe which has very low
reactivity with ozone but readily reacts with •HO (see rate constants in Table 8.1)
[39]. Oxalic acid was used in this experiment at high concentration as •HO
scavenger. Thus, in the presence of both compounds, the evolution of pCBA can
be expressed as:
pCBA ipCBA pCBAHO pCBA HO HO pCBA
s
ipCBAHO pCBA
OxalHO Oxal
dC rk C C k C
dt k
rk C
k ∙C
(8.30)
where CpCBA and COxal are the concentrations of pCBA and oxalic acid,
respectively; kHO∙‐pCBA and kHO∙‐Oxal are the rate constants in Table 8.1 (kHO∙‐Oxal =
4.6x106 M‐1 s‐1, at pH = 4). In this equation, oxalic acid is considered as an
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visible light assisted photocatalytic ozonation with WO3 catalyst
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inhibitor of the chain mechanism of ozone decomposition into •HO radicals [40].
Integration of Eq. (8.30) leads to Eq. (8.31) which simplifies to Eq. (8.32) since
oxalic acid remains virtually constant while pCBA is depleted and ri will be
constant at the conditions used:
0
0
tpCBA HO ‐pCBA i
pCBA OxalHO ‐Oxal t
kC rLn dt
C k C
(8.31)
0pCBA HO ‐pCBAi
pCBA OxalHO ‐Oxal
kCLn r t
C k C
(8.32)
The representation of Eq. (8.32) is depicted in Fig. 8.9(B), and the calculated
value for ri is indicated in Table 8.4.
Table 8.4. Kinetic parameters and correlation coefficient of the model proposed for
PhC‐O3 DEET mineralization
Parameter Value
ri (M s‐1) 2.07x10‐8
kHO∙‐TOC1 (M‐1 s‐1) 5x109
kHO∙‐TOC2 (M‐1 s‐1) 1.3x109
kHO∙‐TOC3 (M‐1 s‐1) 4x108
α 0.62
R2 0.993
CAPÍTULO 8 (CHAPTER 8)
306
0 1x107 2x107 3x107 4x1070.0
0.2
0.4
0.6
0.8
1.0
ln(C
pCB
A ,
0/C
pCB
A)
kHO· - pCBA·t/(kHO· - Oxal·COxal) (M-1s)
Slope = 2.07x10-8 +/- 3.69x10-10
R2 = 0.996
(B)
0 15 30 45 60 75 90 105 1200.0
0.2
0.4
0.6
0.8
1.0
C/C
0
TIME (min)
(A)
Figure 8.9. Determination of kinetic parameters. (A) Time evolution of DEET (triangles)
and TOC (circles) normalized concentrations during PhC‐O3 at different O3 inlet gas
concentration. Symbols: , 10 mg L‐1; , 20 mg L‐1. (B) ri determination from pCBA‐
oxalic acid experiment through data fitting to Eq. (32). (C) ks determination from TOC
evolution. Symbols: TOC1 (line shows fitting of Eq. (33)); TOCt (line shows fitting to
apparent 1st order kinetics); ks. Experimental conditions: CDEET,0 = 15 mg L‐1; pH0 = 6;
V = 0.5 L ; Qg = 15 L h‐1; CO3,g = 10 mg L‐1; *I = 550 W m−2; *CWO3 = 0.25 g L‐1;
*CpCBA,0 = 5 mg L‐1; *COxal,0 = 0.01 M (*if applied).
0 1000 2000 3000 4000 5000 6000 70000.0
0.5
1.0
1.5
2.0
ks=3.67x104·exp(1.84x10-4·t)
Slope (TOCt) = 1.84x10-4+/- 6.35x10-6
R2 = 0.992
ln(C
TO
C,0
/CT
OC
)
Time (s)
Slope (TOC1) = 2.82x10-3 +/- 6.24x10-5
R2 = 0.998
(C)
2.0x104
4.0x104
6.0x104
8.0x104
1.0x105
1.2x105
1.4x105
k s (
s-1)
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During photocatalytic ozonation of DEET, the scavenging factor, kS could not
be considered constant. This scavenging factor will involve all the species
reacting with hydroxyl radicals that inhibit the ozone decomposition
mechanism. During the first moments, when DEET concentration is high and no
mineralization is observed, kS can be determined from integrated Eq. (8.27):
10
11
TOCi
HO ‐TOCTOC s
C rLn k ∙t
C k (8.33)
As observed in Fig. 8.9(C), kS was found to be 3.67x104 s‐1 (for the initial
moments of reaction) from plotting and fitting experimental data to Eq. (8.33).
However, since the composition of the reaction medium changes with time, this
kS value is not valid for the entire reaction period. In general, as reaction
progresses, the concentration of organic intermediates with inhibiting character
such as oxalic or acetic acid increase. Thus, taking into account that the main
organics involved in the scavenging factor are constituents of TOCt, an
approximation was done taking into account that kS undergoes an exponential
variation with time inversely to TOCt, at t = 0 being kS,0 = 3.67x104 s‐1, as
calculated for the beginning of the experiment. The evolution of ln(CTOC,0/CTOC)
and kS with time is depicted in Fig. 8.9(C), the selected fitting equation for kS
being:
4 4sk 3.67x10 exp 1.84x10 t) (8.34)
With all these calculated data, differential Eqs. (8.27), (8.28) and (8.29) can be
simultaneously solved. For that purpose, kHO∙‐TOC3 apparent rate constant of the
reaction between compounds of TOC3 (formaldehyde and oxalic, acetic, pyruvic
and formic acids) and •HO radicals, has been calculated as an average value
taking into account their individual rate constants [31]. Values of kHO∙‐TOC1 and
kHO∙‐TOC3 are summarized in Table 8.4. The other values, kHO∙‐TOC2 and α, were
used as fitting parameters to minimize the difference between experimental data
CAPÍTULO 8 (CHAPTER 8)
308
and those calculated with the proposed model. The resolution of the differential
equations was performed by a fourth‐order Runge‐Kutta method with
Micromath Scientist 3.0 software. Figure 8.10 shows the experimental and
calculated TOC evolution. As observed, the proposed kinetic model fits fairly
well the experimental results (also confirmed by the R2 value in Table 8.4). In
addition, as expected, for the reaction between TOC2 and •HO radicals the value
of the rate constant obtained (kHO∙‐TOC2) is lower than that of DEET‐ •HO reaction
but higher than that of TOC3‐ •HO , which is in accordance with the general trend
of subsequent lower reactivity of the TPs formed in these type of reactions [31].
On the other hand, 62 % of the TOC2 is transformed in TOC3 whereas 38 % can
be mineralized according to the apparent stoichiometric coefficient α.
0 15 30 45 60 75 90 105 1200
2
4
6
8
10
12
CT
OC
(mg
L-1
)
TIME (min)
Figure 8.10. Evolution of TOC with time during DEET degradation by PhC‐O3. Symbols:
TOCt; TOC1 (DEET); □ TOC2 (OPs); TOC3 (Carboxylic acids and formaldehyde);
lines are fitting results from the TOC depletion model. Experimental conditions:
CDEET,0 = 15 mg L‐1; pH0 = 6; V = 0.5 L ; Qg = 15 L h‐1; CO3,g = 10 mg L‐1; I = 550 W m−2;
CWO3 = 0.25 g L‐1.
The proposed kinetic model provides a simplified approach that can be
useful for design purposes, although the nature of the organic pollutant and its
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
309
reactivity with the different oxidizing species generated during photocatalytic
ozonation, together with the operating conditions, will play an important role in
order to extend the model from this particular case to more general
photocatalytic ozonation studies.
8.4. CONCLUSIONS
A synergistic effect between ozone and visible light irradiated WO3 has been
proved. The highest efficiency of photocatalytic ozonation process compared to
photocatalysis and single ozonation is reflected in both, DEET depletion and
mineralization rates, hydroxyl radicals in the bulk being the main species
responsible according to scavenging experiments. At the conditions used in this
work, photocatalytic ozonation process led to complete removal of 15 mg L‐1
DEET in 15 min with mineralization up to 60 % in 2 h. The detailed study of the
evolution with time of 22 TPs detected during ozonation and photocatalytic
ozonation allowed the proposal of a general reaction mechanism through
different pathways mainly based on the hydroxyl radical attack. The reaction
mechanism involves steps of mono‐ and poly‐hydroxylation and/or oxidation,
de‐alkylation and finally opening of the aromatic ring which evolves through
further oxidation, to the formation of short‐chain saturated organic acids and
mineralization to CO2. The evolution of DEET, intermediates and short‐chain
saturated organic acids was fitted to a lumped kinetic model based on TOC‐
hydroxyl radical reactions and provides a simplified approach for this process
that can be useful for design purposes. However, the nature of the organic
pollutant and its reactivity towards the different reactive species generated
during photocatalytic ozonation (ozone, hydroxyl radicals, positive holes, etc.);
as well as the operating conditions (pH, temperature, radiation intensity and
wavelength, etc.) would play an important role in order to extend the model
from this particular case and will be considered for a future work.
CAPÍTULO 8 (CHAPTER 8)
310
AKNOWLEDGEMENTS
Authors thank the Spanish MINECO and European Feder Funds
(CTQ2015/64944‐R) and Junta de Extremadura (Ayuda a Grupos Exp. GR15‐033)
for economic support; and SAEM‐SAIUEX for the LC‐qTOF analyses. E. Mena
thanks the Consejería de Empleo, Empresa e Innovación (Junta de Extremadura)
and European Social Fund for her FPI grant (Ref. PD12059).
REFERENCES
[1] Antonopoulou, M.; Giannakas, A.; Deligiannakis, Y.; Konstantinou, I.
“Kinetic and mechanistic investigation of photocatalytic degradation of the N,N‐
diethyl‐m‐toluamide”. Chem. Eng. J. 231 (2013) 314‐325.
[2] Costanzo, S.D.; Watkinson, A.J.; Murby, E.J.; Kolpin, D.W.; Sandstrom, M.W.
“Is there a risk associated with the insect repellent DEET (N,N‐diethyl‐m‐
toluamide) commonly found in aquatic environments?”. Sci. Total Environ. 384
(2007) 214‐220.
[3] Adams, W.A.; Impellitteri, C.A. “The photocatalysis of N,N‐diethyl‐m‐
toluamide (DEET) using dispersions of Degussa P‐25 TiO2 particles”. J.
Photochem. Photobiol. A Chem. 202 (2009) 28‐32.
[4] Benítez, F.J.; Acero, J.L.; García‐Reyes, J.F.; Real, F.J.; Roldán, G.; Rodríguez,
E.; Molina‐Díaz, A. “Determination of the reaction rate constants and
decomposition mechanisms of ozone with two model emerging contaminants;
DEET and Nortriptyline”. Ind. Eng. Chem. Res. 52 (2013) 17054‐17073.
[5] Petrucci, N.; Sardini. S. “Severe toxic reactions and death following the
ingestion of diethyltoluamide‐containing insect repellents”. Pediatr. Emerg.
Care 16 (2000) 341‐342.
[6] Benítez, F.J.; Acero, J.L.; Real, F.J.; Roldán, G.; Rodríguez, E. “The
effectiveness of single oxidants and AOPs in the degradation of emerging
contaminants in waters: A comparison study”. Ozone Sci. Eng. 35 (2013), 263‐
272.
[7] Li, W.; Nanaboina, V.; Zhou, Q.; Korshin, G.V. “Effects of Fenton treatment
on the properties of effluent organic matter and their relationships with the
degradation of pharmaceuticals and personal care products”. Water Res. 46
(2012) 403‐412.
[8] Medana, C.; Calza, P.; Dal Bello, F.; Raso, E.; Minero, C.; Baiocchi, C.
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
311
“Multiple unknown degradants generated from the insect repellent DEET by
photoinduced processes on TiO2”. J. Mass Spectrom. 46 (2011) 24‐40.
[9] Mena, E.; Rey, A.; Acedo, B.; Beltrán, F.J.; Malato, S. “On ozone‐
photocatalysis synergism in black‐light induced reactions: Oxidizing species
production in photocatalytic ozonation versus heterogeneous photocatalysis”.
Chem. Eng. J. 204‐206 (2012) 131‐140.
[10] Agustina, T.E.; Ang, H.M.; Vareek, V.K. “A review of synergistic effect of
photocatalysis and ozonation on wastewater treatment”. J. Photochem.
Photobiol. C Photochem. Rev. 6 (2005) 264‐273.
[11] Mena, E.; Rey, A.; Beltrán, F.J.; Contreras, S. “Visible light photocatalytic
ozonation of DEET in the presence of different forms of WO3”. Catal. Today 252
(2015) 100‐106.
[12] Tay, K.S.; Abd. Rahman, N.; Bin Abas, M.R. “Degradation of DEET by
ozonation in aqueous solution”. Chemosphere 76 (2009) 1296‐1302.
[13] Bader, H.; Hoigné, J. “Determination of ozone in water by the indigo
method”. Water Res. 15 (1981) 449‐456.
[14] Ilisz, I.; Bokros, A.; Dombi, A. “TiO2‐based heterogeneous photocatalytic
water treatment combined with ozonation”. Ozone Sci. Eng. 26 (2004) 585‐594.
[15] Masschelein, W.; Denis, M.; Ledent, R. “Spectrophotometric determination
of residual hydrogen peroxide”. Water & Sewage Works (1977) 69‐72.
[16] Flyunt, R.; Leitzke, A.; Mark, G.; Mvula, E.; Reisz, E.; Schick, R.; von
Sonntag, C. “Determination of •OH, O2•‐, and hydroperoxide yields in ozone
reactions in aqueous solution”. J. Phys. Chem. B 107 (2003) 7242‐7253.
[17] Nishimoto, S.; Mano, T.; Kameshima, Y.; Miyake, M. “Photocatalytic water
treatment over WO3 under visible light irradiation combined with ozonation”.
Chem. Phys. Lett. 500 (2010) 86‐89.
[18] Beltrán, F.J. “Ozone reaction kinetics for water and wastewater systems”.
Boca Raton, CRC Press, 2004, Florida (USA).
[19] Sánchez, L.; Domènech, X.; Casado, J.; Peral, J. “Solar activated ozonation of
phenol and malic acid”. Chemosphere 50 (2003) 1085‐1093.
[20] Quiñones, D.H.; Rey, A.; Álvarez, P.M.; Beltrán, F.J.; Plucinski, P.K.
“Enhanced activity and reusability of TiO2 loaded magnetic activated carbon for
solar photocatalytic ozonation”. Appl. Catal. B Environ. 144 (2014) 96‐106.
[21] Song, W.; Cooper, W. J.; Peake, B. M.; Mezyk, S. P.; Nickelsen, M. G.;
O’Shea, K. E. “Free‐radical‐induced oxidative and reductive degradation of N,
CAPÍTULO 8 (CHAPTER 8)
312
N′‐diethyl‐m‐toluamide (DEET): Kinetic studies and degradation pathway”.
Water Res. 43 (3) (2009) 635−642.
[22] Yao, C.C.D.; Haag, W.R. “Rate constants for direct reactions of ozone with
several drinking water contaminants”. Water Res. 25 (1991) 761‐773.
[23] Alam, M.S.; Rao, B.S.M.; Janata, E. “OH reactions with aliphatic alcohols:
evaluation of kinetics by direct optical absorption measurement. A pulse
radiolysis study”. Radiat. Phys. Chem. 67 (2003) 723‐728.
[24] Johnson, P.N.; Davis, R.A. “Diffusivity of ozone in water”. J. Chem. Eng.
Data 41 (1996) 1485‐1487.
[25] Froment, G.F.; Bischoff, K.B. “Chemical reactors analysis and design”. John
Wiley & Sons, 1979, New York (USA).
[26] Staehelin, S.; Hoigné, J. “Decomposition of ozone in water: rate of initiation
by hydroxide ions and hydrogen peroxide”. Environ. Sci. Technol. 16 (1982) 666‐
681.
[27] Beltrán, F.J.; Aguinaco, A.; García‐Araya, J.F. “Mechanism and kinetics of
sulfamethoxazole photocatalytic ozonation in water”. Water Res. 43 (2009) 1359‐
1369.
[28] Zhang, L.S.; Wong, K.H.; Zhang, D.Q.; Hu, C.; J.C. Yu, C.Y. Chan, Wong,
P.K. “Zn:In(OH)ySz solid solution nanoplates: Synthesis, characterization, and
photocatalytic mechanism”. Environ. Sci. Technol. 43 (2009) 7883‐7888.
[29] Rodríguez, E.M.; Márquez, G.; Tena, M.; Álvarez, P.M.; Beltrán, F.J.
“Determination of main species involved in the first steps of TiO2 photocatalytic
degradation of organics with the use of scavengers: The case of ofloxacin”. Appl.
Catal. B Environ. 178 (2015) 44‐53.
[30] Hoigné, J.; Bader, H. “Rate constants of reactions of ozone with organic and
inorganic compounds in water—II: Dissociating organic compounds”. Water
Res. 17 (1983) 185‐194.
[31] Buxton, G.V.; Greenstock, C.L.; Helman, W.P. “Critical review of rate
constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl
radicals (.OH/.O−) in aqueous solution”. J. Phys. Chem. Ref. Data 17 (1988) 513‐
886.
[32] Coelho, A. D.; Sans, C.; Agüera, A.; Gomez, M.J.; Esplugas, S.; Dezotti, M.
“Effects of ozone pre‐treatment on diclofenac: Intermediates, biodegradability
and toxicity assessment”. Sci. Total Environ. 407 (2009) 3572−3578.
[33] Zhang, H.; Lemley, A. T. “Reaction mechanism and kinetic modeling of
DEET degradation by flow‐through anodic Fenton treatment (FAFT)”. Environ.
PAPER 6: Reaction mechanism and kinetics of DEET
visible light assisted photocatalytic ozonation with WO3 catalyst
313
Sci. Technol. 40 (2006) 4488−4494.
[34] Calza, P.; Medana, C.; Raso, E.; Giancotti, V.; Minero, C. “N,N‐diethyl‐m‐
toluamide transformation in river water”. Sci. Total Environ. 409 (2011)
3894−3901.
[35] Leitzke, A.; von Sonntag, C. “Ozonolysis of unsaturated acids in aqueous
solution: Acrylic, methacrylic, maleic, fumaric and muconic acids”. Ozone Sci.
Eng. 31 (2009) 301‐308.
[36] Aguinaco, A.; Beltrán, F.J.; Sagasti, J.J.P.; Gimeno, O. “In situ generation of
hydrogen peroxide from pharmaceuticals single ozonation: A comparative
study of its application on Fenton like systems”. Chem. Eng. J. 235 (2014) 46‐ 51.
[37] Beltrán, F.J.; Gimeno, O.; Rivas, F.J.; Carbajo, M. “Photocatalytic ozonation
of gallic acid in water”. J. Chem. Technol. Biotechnol. 81 (2006) 1787‐1796.
[38] Malato, S.; Fernández‐Ibáñez, P.; Maldonado, M.I.; Blanco, J.; Gernjak, W.
“Decontamination and disinfection of water by solar photocatalysis: Recent
overview and trends”. Catal. Today 147 (2009) 1‐59.
[39] Elovitz, M.S.; von Gunten, U. “Hydroxyl radical/ozone ratios during
ozonation processes. I. The Rct concept”. Ozone Sci. Eng. 21 (1999) 239‐260.
[40] Leitner, N.K.V.; Doré, M. “Mecanisme dʹaction des radicaux OH sur les
acides glycolique, glyoxylique, acetique et oxalique en solution aqueuse:
Incidence sur la consammation de peroxyde dʹhydrogene dans les
systemes H2O2‐UV et O3‐H2O2”. Water Res. 31 (1997) 1383‐1397.
CAPÍTULO 9 Conclusiones
En este capítulo se presentan las conclusiones más relevantes que se han ido extrayendo a partir de los resultados obtenidos en el desarrollo del trabajo de investigación que ha dado lugar a la presente Tesis Doctoral.
Conclusiones
317
La realización de esta Tesis Doctoral ha tenido como objetivo principal
contribuir al desarrollo de nuevas estrategias para el tratamiento de
contaminantes emergentes en aguas, centrándose en el proceso de ozonización
fotocatalítica como alternativa eficiente para su eliminación aprovechando la luz
solar como fuente de radiación y sintetizando fotocatalizadores que permitan un
mayor aprovechamiento de la misma. Para lograr este objetivo general se
plantearon una serie de objetivos específicos, tal como se detalla en el Capítulo 1
de este trabajo. A continuación se presentan las conclusiones generales
obtenidas en función del objetivo perseguido.
I) En primer lugar, del estudio realizado con el fin de demostrar y
justificar la mayor eficacia del tratamiento de ozonización fotocatalítica frente a
los procesos individuales (fotocatálisis heterogénea y ozonización simple), se
concluye que:
‐ Los resultados obtenidos sobre la producción de especies oxidantes en los
distintos procesos ha permitido establecer la importancia de las reacciones
directas e indirectas del ozono y la sinergia entre sistemas. A pH ácido,
cuando las reacciones indirectas de ozono son despreciables, el ozono ejerce
un efecto positivo en la velocidad de formación de especies oxidantes foto‐
generadas, aumentando el rendimiento cuántico de 0,34 mol einstein‐1 en
fotocatálisis a 0,80 mol einstein‐1 en ozonización fotocatalítica. A pH = 7 el
incremento fue mucho mayor (de 0,29 a 3,27 mol einstein‐1), si bien en este
caso no puede descartarse la contribución del ozono por vía indirecta. El
aumento del rendimiento cuántico en presencia de ozono, responsable de la
sinergia observada en el proceso combinado, se debe fundamentalmente al
desempeño del ozono como aceptor de electrones del semiconductor
irradiado, minimizando así el proceso de recombinación.
II) En segundo lugar, del estudio realizado sobre la síntesis de nuevos
fotocatalizadores con vistas a incrementar el aprovechamiento de la radiación
solar en la degradación mediante ozonización fotocatalítica de contaminantes
emergentes en agua pura y agua residual, puede concluirse que:
CAPÍTULO 9
318
‐ El óxido de wolframio WO3 ha demostrado ser un material activo y estable
en la degradación de contaminantes en agua en presencia de ozono
empleando tanto luz visible como radiación solar. A diferencia del oxígeno,
el ozono es capaz de reaccionar con los electrones fotogenerados en la
superficie irradiada del semiconductor. Las formas monoclínica y
ortorrómbica del WO3 son las más activas, viéndose la actividad favorecida
por la presencia de estados reducidos o vacantes de oxígeno en la estructura
del WO3, lo que favorece el transporte de electrones en la superficie del
catalizador. Los resultados indican que son los parámetros estructurales y
superficiales y no el desarrollo de superficie específica los que juegan un
papel más importante en la actividad del material. Bajo luz visible, con el
mejor material de esta serie se obtuvo, en función de las condiciones de
ensayo, un consumo específico de ozono de 15 ‐ 49 mg O3/mg COT eliminado
para conseguir una mineralización del 50 % mediante ozonización
fotocatalítica, empleando DEET como compuesto de prueba.
‐ Los resultados sobre síntesis y aplicación de óxidos de cerio CeO2 con
distinta morfología (nanovarillas y nanocubos), indican que las nanovarillas
de CeO2, por su mayor área superficial irradiada y la mayor cantidad de
defectos superficiales y vacantes de oxígeno (lo que condujo a una menor
energía de salto de banda), presentan una mayor actividad con respecto a los
nanocubos en la ozonización fotocatalítica bajo luz visible. Por el contrario,
bajo radiación solar la presencia en los nanocubos de caras {100} expuestas les
confiere una mayor actividad fotocatalítica intrínseca. Con el mejor material
de entre los preparados en esta serie se obtuvo, en las condiciones de ensayo
y empleando tanto luz visible como radiación solar, un consumo específico
de ozono de 70 mg O3/mg COT eliminado para conseguir una mineralización
del 50 % mediante ozonización fotocatalítica empleando DEET como
compuesto de prueba.
‐ De la síntesis y aplicación de catalizadores compuestos de WO3‐TiO2 con un
4 % en peso de WO3 sobre dos soportes de TiO2 con distintas propiedades
Conclusiones
319
estructurales y texturales (TiO2 comercial P25 y nanotubos de TiO2, NT), se
deduce que bajo luz solar ambos materiales resultaron más eficientes que sus
precursores en la degradación de contaminantes emergentes en agua residual
mediante ozonización fotocatalítica, siendo el material NT‐WO3 el que
presentó un mejor comportamiento. Este catalizador estaba compuesto de
partículas alargadas de anatasa, una estructura porosa muy desarrollada y
una alta dispersión de especies WOx. Estas propiedades provocan un mejor
aprovechamiento de la radiación solar visible y, además, una mayor
capacidad de adsorción de compuestos orgánicos en comparación con el TiO2
P25. Con este material, en las condiciones de ensayo y bajo radiación solar se
obtuvo un consumo específico de ozono de 19 mg O3/mg COT eliminado
para conseguir una mineralización del 50 % de un efluente de EDAR dopado
con contaminantes emergentes mediante ozonización fotocatalítica.
‐ Con independencia de las condiciones de ensayo, de la naturaleza de la
matriz acuosa y del tipo de radiación (luz visible o radiación solar), todos los
catalizadores sintetizados favorecieron la sinergia entre el ozono y el
semiconductor. Atendiendo al comportamiento de los distintos materiales
bajo luz visible, para el siguiente estudio se seleccionó por su mejor consumo
específico de ozono, un catalizador de WO3 con estructura monoclínica.
III) Para finalizar, al objeto de profundizar en el comportamiento de los
catalizadores de WO3 en la degradación de contaminantes en agua mediante
ozonización fotocatalítica bajo luz visible y, con ello, poder proponer para este
sistema un modelo cinético, se identificaron las principales especies oxidantes
involucradas y los intermedios de reacción generados en la degradación del
DEET, compuesto modelo elegido debido a su baja reactividad frente al ozono
molecular. Los resultados obtenidos indican que:
‐ Bajo luz visible los radicales hidroxilo en el seno del líquido son las
principales especies involucradas en la degradación y mineralización de
DEET mediante ozonización fotocatalítica. De forma general, en el caso de
contaminantes que presenten una alta reactividad frente al ozono molecular
CAPÍTULO 9
320
este jugará un papel en la eliminación de los mismos, mientras que los
radicales hidroxilo serán fundamentales en la mineralización de los
intermedios más refractarios al ozono.
‐ Los intermedios de degradación de DEET mediante ozonización simple y
fotocatalítica fueron prácticamente los mismos, lo que ha permitido
establecer un mecanismo de reacción basado en el ataque del radical
hidroxilo y que transcurre a través de la mono‐ y poli‐hidroxilación y
oxidación, de‐alquilación y, finalmente, apertura del anillo aromático,
conduciendo a la formación de ácidos orgánicos de cadena corta y la
mineralización a CO2.
‐ En base al mecanismo propuesto y al papel mayoritario del radical hidroxilo
en el mismo, se desarrolló un modelo cinético en términos de COT que
permite simular la concentración de DEET, la de sus principales intermedios
de degradación y la de los ácidos orgánicos de cadena corta, proporcionando
un enfoque simplificado del proceso de ozonización fotocatalítica. No
obstante, la naturaleza del contaminante orgánico, su reactividad frente a las
distintas especies reactivas en el proceso (ozono, radical hidroxilo, huecos
positivos, etc.), así como las condiciones de operación (pH, temperatura, tipo
de radiación empleada, intensidad, etc.), jugarán un papel muy importante
para poder extender el modelo más allá de este caso en particular.
NOMENCLATURA
A Constante en medidas de espectroscopía ultravioleta‐visible de
reflectancia difusa
Absm Absorbancia de una muestra
Abs0 Absorbancia del blanco
act Actinómetro
BC Banda de conducción de un semiconductor
BV Banda de valencia de un semiconductor
c Camino óptico en la Ley de Beer
C Contaminante orgánico genérico
Ci Concentración de un compuesto i en disolución acuosa
Ci,0 Concentración inicial de un compuesto i en disolución acuosa
Ci,g Concentración de un compuesto i en fase gas
CI Carbono inorgánico
COT Carbono orgánico total
CT Carbono total
d Distancia interplanar en medidas de difracción de rayos X
DBO Demanda biológica de oxígeno
Nomenclatura
322
DBO5 Demanda biológica de oxígeno tras 5 días de incubación
DEET N,N‐dietil‐meta‐toluamida
DMA Directiva Marco Europea del Agua
DQO Demanda química de oxígeno
e Electrón
BCe Electrón generado en la banda de conducción de un
semiconductor
se Electrón móvil que migra a la superficie del semiconductor
Te Electrón atrapado formando un estado de menor movilidad
EDAR Estación depuradora de aguas residuales
Eg Energía de salto de banda de un semiconductor
h Constante de Planck
h Hueco positivo
BVh Hueco positivo generado en la banda de valencia de un
semiconductor
sh Hueco positivo móvil que migra a la superficie del semiconductor
Th Hueco positivo atrapado formando un estado de menor
movilidad
He Constante de Henry
HO Ion hidróxido
•HO Radical hidroxilo
sHO Radical hidroxilo superficial
2HO Ion hidroperóxido
2HO Radical hidroperóxido
h Radiación
IBP Ibuprofeno
Ii Intensidad de pico en medidas de espectroscopía fotoelectrónica
de rayos X
Nomenclatura
323
I0 Intensidad de radiación incidente
K Constante relativa al factor de forma de un cristal en medidas de
difracción de rayos X en la ecuación de Scherrer
kHO‐C Constante de reacción entre los radicales hidroxilo y un
contaminante orgánico C
kL Coeficiente individual de transferencia de materia
kLa Coeficiente volumétrico de transferencia de materia
kO3‐C Constante de reacción directa entre el ozono y un contaminante
orgánico C
L Paso efectivo de radiación a través del reactor en medidas
actinométricas
o tamaño de cristal en medidas de difracción de rayos X
LD Límite de detección
n Exponente que toma el valor de ½ para transiciones electrónicas
directas en medidas de espectroscopía ultravioleta‐visible de
reflectancia difusa
u orden de difracción en medidas de difracción de rayos X
NCA Normas de calidad ambiental
ni Número de átomos por cm3 del elemento i en medidas de
espectroscopía fotoelectrónica de rayos X
NT Nanotubos
2sO Ion oxígeno terminal de la red de un semiconductor
2O Radical superóxido
•
3O Radical ozónido
O3dis Ozono disuelto
OMS Organización Mundial de la Salud
PAO Proceso avanzado de oxidación
pHPZC pH del potencial de carga cero
PNCA Plan Nacional de Calidad de las Aguas
PNRA Plan Nacional de Reutilización de Aguas
PT Contenido en fósforo total
P25 Catalizador de TiO2 comercial P25 (Aeroxide®)
Nomenclatura
324
Qg Caudal de gas
R Valor de reflectancia medida respecto a la unidad en medidas de
espectroscopía ultravioleta‐visible de reflectancia difusa
Si Factor de sensibilidad atómica en medidas de espectroscopía
fotoelectrónica de rayos X
T Temperatura
tR Tiempo de retención
u.a. Unidades arbitrarias
US‐EPA Agencia de los Estados Unidos para la protección del
Medioambiente
UV Radiación ultravioleta
UVA Radiación ultravioleta A
UVB Radiación ultravioleta B
UVV Radiación ultravioleta de vacío
V Volumen de reacción
VINY Volumen de inyección
Vm Volumen de muestra
VT Volumen total
Coeficiente de absorción en medidas de espectroscopía
ultravioleta‐visible de reflectancia difusa
Anchura a la mitad de la intensidad máxima de un pico
seleccionado en medidas de difracción de rayos X
H Variación de entalpía
G Variación de entropía
Coeficiente de extinción molar en la Ley de Beer
Rendimiento cuántico de una reacción fotoquímica
hv Rendimiento cuántico de foto‐generación de especies oxidantes
Longitud de onda de la radiación
Frecuencia de la radiación en medidas de espectroscopía
ultravioleta‐visible de reflectancia difusa
2 Ángulo de difracción en medidas de difracción de rayos X
NOMENCLATURE
A Pre‐exponential constant in the Arrhenius equation
AAP Acetaminophen
ANT Antipyrine
AOP Advanced oxidation process
a.u. Arbitrary units
BOD Biological oxygen demand
BOD5 Biological oxygen demand during 5 days of incubation
CAF Caffeine
CAR Carbamazepine
CAT Catalyst
CB Conduction band in a semiconductor
Ci Concentration of a compound i in the liquid phase
Ci,0 Initial concentration of a compound i in the liquid phase
Ci,g Concentration of a compound i in the gas phase
Ci,g inlet Inlet concentration of a compound i in the gas phase
COD Chemical oxygen demand
Nomenclature
326
DCF Diclofenac
DEET N,N‐diethyl‐meta‐toluamide
DO3 Ozone diffusivity in water
DR‐UV‐Vis Diffuse reflectance UV‐Vis spectroscopy
E Degree of reaction rate enhancement
e Electron
aqe Electron in the aqueous phase
Ea Activation energy
EC Emerging contaminant
Eg Optical energy band gap of a semiconductor
FS Integrated absorption fraction of radiation
h Positive hole
Ha Hatta number
HCT Hydrochlorothiazide
He Henry’s constant
•HO Hydroxyl radical
sHO Surface bounded hydroxyl radical
2HO Hydroperoxide radical
HPLC‐DAD High‐performance liquid chromatography – Diode Array detector
HPLC‐qTOF High‐performance liquid chromatography – quadrupole time of
flight
h Radiation
I Photon flux
Ia Photon flux absorbed by the catalyst
IBP Ibuprofen
IC Inorganic carbon
ICP‐OES Inductively coupled plasma – optical emission spectrometry
Ii Intensity of a Raman peak at i cm‐1.
INT Intermediate organic compound
Nomenclature
327
I0 Incident photon flux
k Kinetic constant
KET Ketorolac
kHO‐i Rate constant for the reaction between i and hydroxyl radical
kL Individual liquid phase mass transfer coefficient
kLa Volumetric mass transfer coefficient
kO3‐i Rate constant for the reaction between i and ozone
kS Scavenging factor
kTOC Pseudo‐first order apparent rate constant of TOC removal
L Crystal size
LC50 Lethal concentration of a substance expected to kill 50 % of
organisms in a given population under a defined set of conditions
M Mass
max Maximum
MTP Metoprolol
MWW Municipal wastewater
MWWTP Municipal wastewater treatment plant
n Reaction order
NC Nanocubes
NT Nanotubes
NR Nanorods
2
2
TiOO Terminal oxygen ion of the TiO2 lattice
2O Superoxide radical
•
3O Ozonide radical
O3d Dissolved ozone
O3g Ozone in the gas phase at the reactor outlet
O3gi Ozone in the gas phase at the reactor inlet
Oxal Oxalic acid
pCBA p‐chlorobenzoic acid
Nomenclature
328
PhC‐O2 Photocatalytic oxidation
PhC‐O3 Photocatalytic ozonation
pHPZC pH at the point of zero charge
PPCP Pharmaceutical and Personal Care Product
P25 Commercial TiO2 P25 (Aeroxide®)
Qg Gas flow rate
R Universal gas constant
rHO‐O3 Rate of ozone indirect reactions
rhv Rate of photocatalytic reactions
ri Rate of reaction or formation of i
rO3 Rate of ozone direct reactions
RTE Radiation transfer equation
SBET BET surface area
SEM Scanning electron microscopy
SFM Sulfamethoxazole
T Temperature
t Time
TEM Transmission electron microscopy
TGA‐DTA Thermal gravimetric and differential temperature analysis
TP Transformation product
TOC Total organic carbon
TOC,a TOC corresponding to short‐chain organic acids
u.a. Units of absorbance
UV Ultraviolet radiation
UVA Ultraviolet A radiation
UVC Ultraviolet C radiation
V Reaction volume
VB Valence band in a semiconductor
Vis Visible
Vp Pore volume
Nomenclature
329
v/v Volume fraction
wt Weight
wt/wt Weight fraction
XPS X‐ray photoelectron spectroscopy
XRD X‐ray diffraction
z Stoichiometric coefficient of a reaction
Stoichiometric parameter that represents the fraction of TOC
converted
Liquid holdup
Increment
Molar extinction coefficient
i Quantum yield of the photocatalytic reaction of i
hv Quantum yield of photo‐generated oxidizing species
Photonic efficiency
lim Limiting photonic efficiency
Radiation wavelength
g Gas flow rate
2 Diffraction angle in XRD analysis