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DESARROLLO DE METODOLOGÍA ANALÍTICA
PARA LA DETERMINACIÓN DE FILTROS SOLARES
EN MUESTRAS AMBIENTALES Y COMPUESTOS
RELACIONADOS EN ALIMENTOS ENVASADOS
NOELIA NEGREIRA FERROL
Memoria para optar al grado de Doctora en Química
Santiago de Compostela, Diciembre 2010
UNIVERSIDAD DE SANTIAGO DE COMPOSTELA
Facultad de Química Departamento de Química Analítica, Nutrición y Bromatología Instituto de Investigación y Análisis Alimentario (IIAA)
D. Antonio Moreda Piñeiro, Profesor Titular de Universidad y Director
del Departamento de Química Analítica, Nutrición y Bromatología de la
Universidad de Santiago de Compostela,
Informa:
Que Dña. Noelia Negreira Ferrol presenta el trabajo “DESARROLLO DE
METODOLOGÍA ANALÍTICA PARA LA DETERMINACIÓN DE FILTROS
SOLARES EN MUESTRAS AMBIENTALES Y COMPUESTOS
RELACIONADOS EN ALIMENTOS ENVASADOS” que ha realizado en este
departamento bajo la dirección de D. Isaac Rodríguez Pereiro y Dña. Elisa
Rubí Cano, Profesores Titulares de Universidad, para optar al grado de
Doctora en Química.
Y para que así conste, firmo el presente informe en Santiago de
Compostela, 3 de diciembre de 2010.
D. Antonio Moreda Piñeiro
D. Isaac Rodríguez Pereiro y Dña. Elisa Rubí Cano, Profesores Titulares
de Universidad del Departamento de Química Analítica, Nutrición y
Bromatología de la Universidad de Santiago de Compostela,
Autorizan:
A la licenciada Dña. Noelia Negreira Ferrol a la presentación del trabajo
recogido en la memoria titulada “DESARROLLO DE METODOLOGÍA
ANALÍTICA PARA LA DETERMINACIÓN DE FILTROS SOLARES EN
MUESTRAS AMBIENTALES Y COMPUESTOS RELACIONADOS EN
ALIMENTOS ENVASADOS” que ha realizado bajo su dirección en el
Departamento de Química Analítica, Nutrición y Bromatología de la Facultad
de Química de la Universidad de Santiago de Compostela para optar al grado
de Doctora en Química
Y, para que así conste, firmamos el presente informe en Santiago de
Compostela, 3 de diciembre de 2010.
D. Isaac Rodríguez Pereiro Dña. Elisa Rubí Cano
AGRADECIMIENTOS
Al finalizar un trabajo tan lleno de dificultades, como lo es una Tesis
Doctoral para la que se requiere de mucho esfuerzo y dedicación por parte de la
doctoranda y de los directores, resulta inevitable mostrar mis más sinceros
agradecimientos a aquellas personas que de una forma desinteresada han
contribuido a la elaboración de esta Tesis y sin las cuales, no hubiera sido
posible su finalización.
En primer lugar, quiero dar las gracias a mis directores, D. Isaac
Rodríguez Pereiro y Dña. Elisa Rubí Cano, por haberme dado la oportunidad
de realizar este trabajo, por poner vuestro apoyo y confianza en mí. En especial,
a Isaac por tus ideas, orientación, respuestas ante inquietudes surgidas, en
definitiva, por tu incalculable aportación al desarrollo de esta tesis y mi
formación como investigadora.
Quisiera también agradecer al Departamento de Química Analítica,
Nutrición y Bromatología, al Instituto de Investigación y Análisis Alimentarios
(IIAA), y en especial a D. Rafael Cela, director del grupo de investigación de
Cromatografía de Gases y Quimiometría, por haber puesto a mi disposición
todos los medios técnicos y materiales para la realización de esta Tesis.
Asimismo, debo agradecer al Ministerio de Educación por la beca FPU
disfrutada y a los fondos autonómicos, estatales y europeos que han financiado
mi trabajo (2010-02, DE2009-0020, DGICT CTQ2009-08377, CTQ2006-03334 y
PGIDIT06PXIB237039PR).
A mis compañeros del IIAA, tanto a los que ya se han ido (Pablo, Lucía y
María F.) como a los quedan (en especial, Iria y Paula) y, sobre todo, a los que
están empezando desearles mucha suerte y paciencia.
A Jorge, por ser la persona con la que compartí más momentos dentro y
fuera del laboratorio durante esta etapa. Tu compañía hizo que me olvidara de
los ratos amargos y tu alegría me dieron la fuerza necesaria para seguir
adelante. Gracias por tu comprensión, tu paciencia y tu cariño. Gracias por esos
momentos inolvidables.
A mis amigas, María, Bea, Lupe y por supuesto, Isabel, gracias por vuestra
compañía, por ayudarme a sobrellevar los malos momentos, y por ofrecerme
amistad y diversión. A Carmen Trillo, por escucharme y aconsejarme en
momentos de angustia, por brindarme tu apoyo y tus ánimos.
Y, por supuesto, el agradecimiento más profundo a mi familia, en especial
a mis padres y a mi hermana, por vuestro apoyo, cariño, consejos, por estar
siempre a mi lado y por hacerme feliz.
En resumen, agradezco a todas y cada una de las personas que han vivido
conmigo la realización de este trabajo, con sus altos y bajos, y en especial, a
aquellas que han sido mi soporte en momentos de angustia y desesperación.
Más que el brillo de la victoria,
nos conmueve la entereza ante la adversidad
(Octavio Paz)
ÍNDICE
Índice
ÍNDICE
I. JUSTIFICACIÓN Y OBJETIVOS ...........................................................1
II. INTRODUCCIÓN ..................................................................................... 5
A. Filtros solares.......................................................................................................... 7
1. ASPECTOS GENERALES ................................................................................. 7
1.1. Definición y clasificación .................................................................. 7
1.2. Filtros solares permitidos por la Unión Europea. ........................ 9
1.3. Filtros solares considerados en este estudio ................................ 15
2. DISTRIBUCIÓN MEDIOAMBIENTAL ........................................................ 21
2.1. Muestras de agua ............................................................................ 22
2.1.1. Niveles de filtros solares en agua ............................................ 22
2.1.2. Toxicidad y reacciones de transformación de los filtros
solares en el medio acuoso ............................................................... 29
2.2. Lodos y sedimentos ........................................................................ 32
2.3. Biota ................................................................................................... 36
3. METODOLOGÍA ANALÍTICA ..................................................................... 38
3.1. Preparación de muestra .................................................................. 38
3.1.1. Muestras de agua ....................................................................... 38
3.1.1.1. Extracción en fase sólida (SPE) .................................... 38
3.1.1.2. Técnicas basadas en la microextracción en fase sólida 46
3.1.1.2.1. Microextracción en fase sólida (SPME) ................... 46
3.1.1.2.2. Microextracción con barras agitadoras (SBSE) ......... 52
3.1.1.2.3. Microextracción mediante sorbentes empaquetados
(MEPS) .................................................................... 55
3.1.1.3. Técnicas basadas en la microextracción en fase
líquida……………………………………………………...57
3.1.1.3.1. Microextracción con gota suspendida (SDME) ........ 57
3.1.1.3.2. Microextracción líquido-líquido con membranas .... 58
3.1.1.3.3. Microextracción en fase líquida con fibra hueca (HF-
LPME) ..................................................................... 58
Índice
3.1.1.3.4. Microextracción líquido-líquido dispersiva
(DLLME)…………………………………………..59
3.1.2. Muestras sólidas ........................................................................ 67
3.1.2.1. Dispersión de la matriz en fase sólida (MSPD) ........... 67
3.1.2.2. Extracción con disolventes presurizados (PLE) .......... 70
3.1.2.3. Otras técnicas de extracción ........................................ 76
3.1.3. Muestras de biota ...................................................................... 78
3.2. Técnicas de determinación ............................................................. 81
3.2.1. Cromatografía de gases acoplada a espectrometría de masas
simple (GC-MS) y en tándem (GC-MS/MS) ................................. 82
3.2.2. Cromatografía líquida acoplada a espectrometría de masas
(LC-MS) ............................................................................................... 91
B. Fotoiniciadores .................................................................................................... 97
1. ASPECTOS GENERALES ............................................................................... 97
2. ESTRUCTURA Y PROPIEDADES ................................................................ 98
3. PRESENCIA EN ALIMENTOS ...................................................................... 99
4. METODOLOGÍA ANALÍTICA PARA LA DETERMINACIÓN DE
FOTOINICIADORES EN ALIMENTOS ................................................ 101
III. METODOLOGÍA DESARROLLADA .......................................................... 105
A. Filtros solares .................................................................................................... 107
1. MUESTRAS ACUOSAS ................................................................................ 107
1.1. Introducción ................................................................................... 107
1.2. Esquemas de los métodos desarrollados para muestras
acuosas.. .......................................................................................... 110
1.3. Publicación: “Dispersive liquid-liquid microextraction followed by gas
chromatography-mass spectrometry for the rapid and sensitive determination of UV
filters in environmental water samples” ....................................................... 117
1.4. Publicación: “Silicone discs as disposable enrichment probes for gas
chromatography-mass spectrometry determination of UV filters in water
samples”……. ........................................................................... ………145
Índice
1.5. Publicación: “Sensitive determination of salicylate and benzophenone type UV
filters in water samples using solid-phase microextraction, derivatization and gas
chromatography tandem mass spectrometry” ............................................... 173
1.6. Publicación: “Solid-phase extraction followed by liquid chromatography tandem
mass spectrometry for the determination of hydroxylated benzophenone UV absorbers
in environmental water samples" ............................................................... 201
1.7. Publicación: “Study of some UV filters stability in chlorinated water and
identification of halogenated by-products by gas chromatography-mass spectrometry”
............................................................................................................ 225
2. MUESTRAS SÓLIDAS .................................................................................. 251
2.1. Introducción ................................................................................... 253
2.2. Esquemas de los métodos desarrollados para muestras
sólidas….......................................................................................... 255
2.3. Publicación: “Optimization of pressurized liquid extraction and purification
conditions for gas chromatography-mass spectrometry determination of UV filters in
sludge” .................................................................................................. 257
2.4. Publicación: “Determination of selected UV filters in indoor dust using matrix
solid-phase dispersion and gas chromatography tandem mass spectrometry” ....... 281
B. Fotoiniciadores .................................................................................................. 305
1. FOTOINICIADORES EN ALIMENTOS ..................................................... 307
1.1. Introducción ................................................................................... 307
1.2. Esquema del método desarrollado para fotoiniciadores en
alimentos......................................................................................... 308
1.3. Publicación: “Solid-phase microextraction followed by gas chromatography mass
spectrometry for the determination of ink photo-initiators in packed milk” .......... 309
IV. CONCLUSIONES ................................................................................. 337
V. BIBLIOGRAFÍA .................................................................................... 345
VI. ABREVIATURAS Y ACRÓNIMOS .................................................. 357
VII. Anexo: OTRAS PUBLICACIONES ................................................... 363
Justificación y objetivos
3
I. JUSTIFICACIÓN Y OBJETIVOS
Actualmente, se utilizan un gran número de compuestos químicos como
aditivos en alimentos, productos de cuidado personal, textiles, etc. En los
últimos años, el interés por los posibles efectos de estos compuestos, conocidos
como contaminantes emergentes, se ha incrementado de forma notable. Dentro
de este grupo se engloban los filtros solares, también denominados filtros UV.
Estos compuestos constituyen los principios activos de protectores solares, cuyo
consumo ha aumentado notablemente en los últimos años [Richardson, 2006]
debido a la concienciación de los efectos nocivos de las radiaciones ultravioletas
sobre la salud. Además, los filtros solares aparecen también en la formulación
de otros cosméticos de uso diario como cremas de belleza, lociones para la piel,
barras de labios, sprays para el pelo, tintes, champú e incluso en ropa y
plásticos [Richardson, 2006] [Salvador, 2005-A; Salvador, 2005-B] [Chisvert, 2001-A]
[Schlumpf, 2004]. Las aplicaciones anteriores provocan la introducción, directa o
indirecta, de filtros UV en el medio acuático. Sus efectos medioambientales son
todavía desconocidos, sin embargo es sabido que algunos de ellos presentan
actividad estrogénica [Díaz-Cruz, 2009] e incluso tendencia a acumularse en
sedimentos y lodos.
Otra familia de compuestos considerada en este trabajo es la
correspondiente a los llamados fotoiniciadores, usados como aditivos en las
tintas de envases alimentarios con el fin de acelerar el secado de las impresiones
realizadas sobre los mismos. Desde un punto de vista estructural, la mayoría de
los fotoiniciadores son derivados de la benzofenona o del ácido p-
aminobenzoico, empleados como filtros UV en productos de cuidado personal.
Recientemente, algunos fotoiniciadores han sido detectados en diversos
alimentos, entre ellos leche, causando alarma social y pérdidas económicas
derivadas de la destrucción de estos alimentos.
Justificación y objetivos
4
El objetivo general de esta Tesis Doctoral ha sido la optimización de
nuevas metodologías para la determinación de filtros solares en muestras
ambientales y de fotoiniciadores en leche. En todos los casos se ha minimizado,
en la medida de lo posible, el consumo de disolventes y la generación de
residuos peligrosos, intentando mejorar las prestaciones analíticas (límites de
cuantificación, exactitud, coste, tiempo de respuesta, etc.) de los métodos ya
disponibles en la bibliografía.
La determinación de filtros solares se llevó a cabo en matrices acuosas
(agua de río, piscina y residual) y sólidas (polvo y lodos). Los métodos
analíticos propuestos se basan en el empleo de técnicas de preparación de
muestra adecuadas a cada matriz. Entre ellas se encuentran la extracción en fase
sólida (SPE), la microextracción en fase sólida (SPME), la absorción sobre
siliconas, y la microextracción líquido-líquido dispersiva (DLLME) para
muestras de agua. En el caso de las matrices sólidas se desarrollaron métodos
de dispersión de la matriz en fase sólida (MSPD) y extracción con fluidos
presurizados (PLE). Las técnicas de determinación empleadas fueron la
cromatografía de gases y la cromatografía de líquidos en combinación con
espectrometría de masas. Una vez optimizadas y validadas las condiciones de
trabajo, los métodos fueron aplicados al estudio de la distribución de filtros
solares en muestras de agua superficial y residual, lodos de estaciones
depuradoras y polvo procedente de atmósferas interiores. También se han
realizado estudios de laboratorio para evaluar la reactividad de ciertos filtros
UV cuando entran en contacto con aguas cloradas.
Por último, un capítulo de esta Tesis se ha dedicado a la optimización y
validación de un método rápido y robusto para la determinación de
fotoiniciadores en muestras de leche. En este caso, se combinó la utilización de
la microextracción en fase sólida (SPME) con cromatografía de gases acoplada a
espectrometría de masas para la determinación selectiva de los compuestos
anteriores evitando la utilización de disolventes orgánicos.
Introducción-Filtros solares
7
II. INTRODUCCIÓN
A. Filtros solares
1. ASPECTOS GENERALES
1.1. Definición y clasificación
Los protectores solares son preparados farmacéuticos de aplicación tópica
que tienen la propiedad de absorber o reflejar la radiación ultravioleta de origen
solar, o de fuentes artificiales, atenuando la acción perjudicial de los rayos
solares. Generalmente, los compuestos activos en protectores solares protegen
contra radiación tipo UVB (290-320 nm) y algunos contra UVA (320-400 nm), la
cual puede penetrar en profundidad en la piel y, después de una larga
exposición, dañarla fuertemente [Chisvert, 2001-A; Chisvert, 2001-B] [Salvador,
2005-A; Salvador, 2005-C]. Actualmente, hay una tendencia a aumentar el factor
de protección lo que, generalmente, significa mayor concentración de filtros UV
en los protectores solares. Frecuentemente, se usan varios filtros solares para
cubrir un amplio rango de longitudes de onda [Balmer, 2005].
La luz solar provoca daño cutáneo porque las radiaciones ultravioleta
(UV) son absorbidas por el ADN, ARN, proteínas, lípidos de membranas y
orgánulos celulares presentes en la epidermis y dermis, incluyendo el sistema
vascular. Los efectos son acumulativos y están en relación con la duración,
frecuencia e intensidad de radiación. El efecto inmediato es la inflamación y el
tardío, cáncer de piel. El 95% de radiaciones que inciden sobre nuestra piel son
infrarrojos (longitudes de onda superiores a 760 nm), y luz visible (400-760 nm).
Sólo el 5% es radiación UV, de la cual el 2% corresponde a UVB y el 98% a UVA
que, a su vez puede dividirse en UVA largos o UVA-I (340-400 nm) y UVA
cortos o UVA-II (320-340 nm). La denominada radiación UVC (con longitudes
de onda inferiores a 290 nm) no llega a nuestra piel ya que es absorbida por la
Introducción-Filtros solares
8
capa de ozono, aunque comenzó a tomar importancia debido a la progresiva
disminución de ésta en los últimos años.
Atendiendo a su naturaleza química, los filtros solares pueden clasificarse
en:
Filtros solares de naturaleza inorgánica. Debido al tamaño y a la
uniformidad de las partículas que los componen, son fotoestables y
seguros. En concentraciones altas pueden sufrir aglomeración
presentando un aspecto blanquecino. Actúan mediante atenuación de la
radiación UV, resultado de la combinación de mecanismos de reflexión,
dispersión y absorción. Los ejemplos más represantativos de este tipo de
compuestos son:
- Óxido de zinc: es un óxido metálico empleado históricamente
como protector de la piel. Es seguro para la aplicación en piel
inflamada y con afectación de la barrera cutánea. Protege de
radiaciones UVB, UVA-II y parcialmente de UVA-I.
- Dióxido de titanio: es un óxido metálico, casi inerte. Protege
frente a UVB y UVA-II.
Filtros solares de naturaleza orgánica. Son los más empleados y se
caracterizan por poseer estructuras aromáticas simples o múltiples, a
menudo substituidas con grupos hidrofóbicos para mejorar sus
propiedades. Estos compuestos son aplicados superficialmente sobre la
piel y su penetración subcutánea es limitada [Straub, 2002]. Sin embargo,
estudios recientes han demostrado su presencia en fluidos biológicos,
tales como orina [Ye, 2005] [Vidal, 2007] [Kawaguchi, 2008-B; Kawaguchi,
2009] [Ito, 2009] [Knisue, 2010], plasma [Sarveiya, 2004] y semen [León,
2010]. Pueden dividirse en dos grupos según absorban radiación UVA o
UVB y su combinación, junto con los filtros de naturaleza inorgánica,
Introducción-Filtros solares
9
permite obtener preparados que cubren un amplio rango de longitudes
de onda y aumentan la eficacia de protección solar.
1.2. Filtros solares permitidos por la Unión Europea.
En general, las concentraciones máximas permitidas de filtros UV en
productos de uso tópico (protectores solares) de acuerdo a la legislación
Europea están entre el 5 y el 10% [Chisvert, 2002-A; Chisvert, 2002-B] [Schakel,
2004]. Existen otras sustancias que podrían ser usadas como filtros solares, que
se encuentran autorizadas por la legislación americana (FDA) o la japonesa,
pero no todas son permitidas por la Unión Europea. Además, compuestos cuyo
uso está permitido actualmente posiblemente se retiren del mercado a medio
plazo. Por ejemplo, el ácido p-aminobenzoico (PABA) causa problemas
dermatológicos [Chisvert, 2001-A] y aún así, sigue permitido en un 15% por
FDA (Estados Unidos) y en un 5% por la EU [Wang, 2007]. Muchos otros son
objeto de estudio entre ellos el ácido 2-hidroxi-4-metoxibenzofenona-5-
sulfónico (BP-4), la 2-hidroxi-4-metoxibenzofenona (BP-3), el
butilmetoxidibenzoilmetano, el 2-etilhexil-4-dimetilaminobenzoato (EHPABA),
el 2-etilthexil-p-metoxicinamato (EHMC), el homosalato (HMS) y el 2-
etilhexilsalicilato (EHS) [Chisvert, 2001-A]. Por último, es necesario destacar que
la legislación comentada anteriormente es sólo aplicable a protectores solares,
pero no a otros productos de cuidado personal, tales como cremas faciales,
colonias, etc. que pueden contener otros compuestos como filtros de la
radiación ultravioleta.
A continuación, se muestra una tabla con la lista de filtros UV que pueden
contener los protectores solares comercializados en la Unión Europea, así como
la concentración máxima autorizada y el tipo de radiación UV que absorben,
Tabla 1.
Introducción-Filtros solares
10
Tabla 1: Filtros solares permitidos por la legislación europea1.
Nombre químico Nombre común
(acrónimo)
Tipo
absorción
UV
Concentración
máxima
autorizada(%)
Filtros de naturaleza inorgánica
Dióxido de Titanio Dióxido de Titanio (TiO2) A, B 25
Filtros de naturaleza orgánica
Derivados de la benzofenona
2-hidroxi-4-metoxibenzofenona Benzofenona-3 (BP-3),
Oxibenzona A, B 10
Ácido 2-hidroxi-4-
metoxibenzofenona-5-sulfónico y
su sal sódica
Benzofenona-4
(BP-4) A, B
5
(expresada
como ácido)
Ácido p-aminobenzoico y sus derivados
Ácido 4-aminobenzoico PABA B 5
2-etilhexil-4-
dimetilaminobenzoato
2-etilhexil-4-
dimetilaminobenzoato
(EHPABA)
B 8
Etil-4-aminobenzoato etoxilado PEG-25 PABA B 10
Ácido benzoico,2-(4-
(dietilamino)-2-hidroxibenzoil-,
hexiléster
Dietilamino hidroxibenzoil
benzoato de hexil A 10
Salicilatos
2-etilhexilo salicilato Etilhexil salicilato (EHS) B 5
3,3,5-trimetilciclohexilo salicilato Homosalato
(HMS) B 10
Metoxicinamatos
4-metoxicinamato de 2-etilhexilo
o metoxicinamato de octilo
Etilhexil metoxicinamato
(EHMC) B 10
4-metoxicinamato de isoamilo Isoamil metoxicinamato
(IAMC) B 10
1 Anexo VI: REGULATION (EU) 1223/2009 del Parlamento Europeo del 30 de
noviembre 2009 relativa a productos cosméticos.
Introducción-Filtros solares
11
Tabla 1 cont.: Filtros solares permitidos por la legislación europea.
Nombre químico Nombre común
(acrónimo)
Tipo absorción
UV
Concentración máxima
autorizada(%) Derivados del alcanfor
3-Bencilideno alcanfor 3-Bencilideno
alcanfor (3-BC) B 2
3-(4’-Metilbencilideno)-d-1 alcanfor 4-metilbencilideno
alcanfor (4-MBC) B 4
Ácido α-(2-oxoborn-3-iliden)-toluen-4-
sulfónico y sus sales
Ácido bencilideno
alcanfor sulfónico
(BCS)
B
6
(expresada
como ácido)
Sulfato de metilo de N,N,N-trimetil[(oxo-
2-bornilideno-3)metil]-6-anilinio
Alcanfor
benzalconio
metosulfato (CBM)
B 6
Ácido 3,3’-(1,4-Fenilendimetileno)bis(7,7-
dimetil-2-oxobiciclo-[2,2,1]hept-1-
ilmetanosulfónico) y sus sales
Ácido tereftaliden
dialcanfor sulfónico
(TDS)
A
10
(expresada
como ácido)
Polímero de N-{(2 y 4)-[(2-oxoborn-3-
iliden)metil]bencil} acrilamida
Poliacrilamidometil
bencilideno alcanfor
(PBC)
B 6
Derivados de la triazina
2,4,6-Trianilina-(p-carbo-2’-etilhexil-1’oxi)-
1,3,5-triazina
Etilhexil triazona
(ET) B 5
Benzoato de bis (2-etilhexil) 4,4’-[[6-[[[(1,1-
dimetiletil)amino]carbonil]fenil]amino]-
1,3,5-triazina-2,4-diil]diimino]bis-
Dioctil butamido
triazona
(DBT)
B 10
2,2’-[6-(4-Metoxifenil)-1,3,5- triazina-2,4-
diil]bis[5- [(2etilhexil)oxi]fenol
Bis-etilhexiloxifenol
metoxifenil triazina
(EMT)
A, B 10
Derivados del benzotriazol
2-(2H-benzotriazol-2-il)-4-metil-6-[2-metil-
3-(1,3,3,3-tetrametil-1-(((trimetilsilil)oxi)-
disiloxanil)propil], fenol
Drometrizol
trisiloxano (DRT) A, B 15
2,2’-metilen-bis-[4-(1,1,3,3-tetrametilbutil)-
6-(2H-benzotriazol-2-il) fenol
Metilen
bisbenzotriazolil
tetrametilbutilfenol
(MBP)
A, B 10
Introducción-Filtros solares
12
Tabla 1 cont.: Filtros solares permitidos por la legislación europea.
Nombre químico Nombre común
(acrónimo)
Tipo
absorción
UV
Concentración
máxima
autorizada(%)
Derivados del benzoimidazol
Ácido 2-Fenilbencimidazol-5-
sulfónico y sus sales de potasio, de
sodio y de trietanolamina
Ácido fenilbencimidazol
sulfónico (PBS) B
8
(expresada
como ácido)
Sal monosódica del ácido 2,2’-bis-
(1,4-fenilen)1H-benzimidazol-4,6-
disulfónico
Ácido
fenildibenzimidazol
tetrasulfónico (PDT)
A
10
(expresada
como ácido)
Otros
1-(4-tert-butil-fenil)-3-(4-
metoxifenil)propano-1,3-diona
Butil
metoxidibenzoilmetano
(BDM)
A 5
Éster 2-etilhexil del ácido 2-ciano-
3,3-difenilacrílico
Octocrileno
(OCR) B
10
(expresada
como ácido)
Dimetilcodietilbencilmalonato Polisilicona-15 (P-15) B 10
A continuación, se muestran las estructuras químicas para algunos filtros
solares, permitidos por la Unión europea, y para los que apenas existe
metodología para su determinación en muestras medioambientales (Fig.1). No
se incluyen los compuestos considerados en este estudio ya que sus
propiedades se presentan en el siguiente apartado de esta memoria.
Ácido p-aminobenzoico y sus derivados:
Z
PABA x+y+z=25PEG-25 PABA
Dietilamino hidroxibenzoil benzoato de hexil
Introducción-Filtros solares
13
Derivados del alcanfor:
3-Bencilideno alcanfor(3-BC)
Poliacrilamidometil bencilideno alcanfor
(PBC)
Ácido bencilidenoalcanfor sulfónico
(BCS)
Alcanfor benzalconio metosulfato
(CBM)
Ácido tereftaliden dialcanfor sulfónico
(TDS)
Derivados de la triazina:
Etilhexil triazona(ET)
Bis-etilhexiloxifenol
metoxifenil triazina(EMT)
Doctil butamidotriazona
(DBT)
R1=R2=R3
R1=R2
R3
R3
R1=R2
Introducción-Filtros solares
14
Derivados del benzotriazol:
Drometrizol trisiloxano (DRT)
Metilen bisbenzotriazoliltetrametilbutilfenol
(MBP)
2
Derivados del benzoimidazol:
Ácido fenildibencimidazoltetrasulfónico
(PDT)
Ácido fenilbencimidazolsulfónico
(PBS)
Otros:
Butil metoxidibenzoilmetano
(BDM)
Figura 1: Estructuras químicas de distintas familias de filtros solares [Salvador, 2005-
A].
Introducción-Filtros solares
15
1.3. Filtros solares considerados en este estudio
A continuación, se presentan de forma detallada las características físico-
químicas más relevantes de los filtros solares considerados en este estudio,
además de otros compuestos con estructuras químicas muy semejantes2.
a) La familia de las benzofenonas se caracteriza por presentar la misma
estructura base, con sustituyentes hidroxi y metoxi en diferentes
posiciones de los anillos aromáticos. En la Unión Europea (UE) sólo está
permitido el uso de 2-hidroxi-4-metoxibenzofenona, también
denominada benzofenona-3 (BP-3 o Bz-3), y el ácido 2-hidroxi-4-
metoxibenzofenona-5-sulfónico (BP-4) en protectores solares, pero en
otros países como en Japón también se permite el uso de 2,4-
dihidroxibenzofenona o 2,4-dihidroxifenilmetanona (BP-1), 2,2’,4,4’-
tetrahidroxibenzofenona (BP-2) y 2,2’-dihidroxi-4,4’-metoxibenzofenona
(BP-6) [Shaath, 2007]. Además, la BP-1 es el principal metabolito de la BP-
3 detectado en muestras de orina [Gonzalez, 2008] [Díaz-Cruz, 2008]. Otras
benzofenonas, tales como la 2,2’-dihidroxi-4-metoxibenzofenona (BP-8)
son incluidas como fotoestabilizadores en muchos otros productos de
cuidado personal, barnices, ropa y plásticos para envasado de alimentos
[Richardson, 2008] [Jeon, 2006]. A continuación se presentan las
estructuras químicas de las benzofenonas consideradas en este estudio
(Fig. 2).
2 Calculated using Advanced Chemistry Development (ACD/Labs) Software
V8.14 for Solaris (© 1994-2009 ACD/Labs).
Introducción-Filtros solares
16
BP‐4 BP‐8
BP‐3BP‐2BP‐1
BP‐6
Figura 2: Estructuras de las benzofenonas.
Las propiedades físico-químicas de las benzofenonas permitidas por la
UE (BP-3 y BP-4) y de las demás benzofenonas consideradas en esta memoria
(BP-1, BP-2, BP-6 y BP-8) se recogen en las siguientes tablas (Tabla 2 y Tabla 3). A
excepción de la BP-4, se trata de compuestos ligeramente ácidos (pKas entre 7 y
7,6 unidades) y de moderada polaridad. Por su parte, la BP-4 es altamente
soluble en agua y presenta un carácter fuertemente ácido.
Tabla 2: Propiedades físico-químicas de las benzofenonas permitidas por la UE.
Propiedad Compuesto
BP-3 BP-4
Nº CAS 131-57-7 4065-45-6
Peso molecular (g mol-1) 228,2 308,3
Densidad (g cm-3) 1,20 ± 0,06 1,45 ± 0,06
pKa 7,6 ± 0,4
-0,7 ± 0,5
Entalpía de vaporización (KJ mol-1) 64 ± 3 -
logKow 3,6 ± 0,4 0,9 ± 0,5
Presión de vapor (mTorr) 0,0526 -
Punto de ebullición (ºC) 370 ± 27 491
Punto de fusión (ºC) 141 ± 17 145-190
Solubilidad molar (mol L-1), pH 1 9,2E-4 0,11
Solubilidad molar (mol L-1), pH 7 1,2E-3 3,24
Solubilidad molar (mol L-1), pH 10 0,21 3,24
-, dato no disponible
Introducción-Filtros solares
17
Tabla 3: Propiedades físico-químicas de BP-1, BP-2, BP-6 y BP-8.
Propiedad Compuesto
BP-1 BP-2 BP-6 BP-8
Nº CAS 131-56-6 131-55-5 131-54-4 131-53-3
Peso molecular (g mol-1) 214,2 246,2 274,3 244,2
Densidad (g cm-3) 1,30 ± 0,06 1,53 ± 0,06 1,29 ± 0,06 1,30 ± 0,06
pKa 7,5 ± 0,4 7,0 ± 0,4 7,0 ± 0,4 7,0 ± 0,4
Entalpía de vaporización (KJ mol-1) 67 ± 3 84 ± 3 72 ± 3 65 ± 3
logKow 3,2 ± 0,4 3,1 ± 0,4 4,1 ± 0,0 3,9 ± 0,4
Presión de vapor (mTorr) 0,012 6,69E-9 2,49E-5 0,0373
Punto de ebullición (ºC) 391 ± 27 531 ± 25 439 ± 45 375 ± 0,0
Punto de fusión (ºC) 215 ± 17 289 ± 20 164 ± 22 146 ± 19
Solubilidad molar (mol L-1), pH 1 3,0E-3 4,0E-3 2,4E-4 9,5E-4
Solubilidad molar (mol L-1), pH 7 3,0E-3 0,013 7,0E-4 1,8E-3
Solubilidad molar (mol L-1), pH 10 4,67 4,06 3,15 4,09
b) El derivado del ácido p-aminobenzoico, más conocido como PABA,
considerado a lo largo de este estudio fue su éster: 2-etilhexil-4-
dimetilaminobenzoato (EHPABA), empleado como filtro solar. A
continuación se presentan sus estructuras (Fig. 3) y sus propiedades
físico-químicas (Tabla 4).
PABA EHPABA
Figura 3: Estructuras del PABA y del EHPABA.
Introducción-Filtros solares
18
Tabla 4: Propiedades físico-químicas del PABA y del EHPABA.
Propiedad Compuesto
PABA EHPABA
Nº CAS 150-13-0 21245-02-3
Peso molecular (g mol-1) 137,1 277,4
Densidad (g cm-3) 1,32 ± 0,06 0,99 ± 0,06
pKa (grupo carboxilo) 4,9 ± 0,1 -
pKa (grupo amino) 2,5 ± 0,1 2,4 ± 0,1
Entalpía de vaporización (KJ mol-1) 62 ± 3 63 ± 3
logKow 0,8 ± 0,2 6,1 ± 0,2
Presión de vapor (mTorr) 0,0345 0,00457
Punto de ebullición (ºC) 340 ± 25 383 ± 25
Punto de fusión (ºC) 159 ± 23 122 ± 14
Solubilidad molar (mol L-1), pH 1 1,61 1,8E-4
Solubilidad molar (mol L-1), pH 3 0,075 9,5E-6
Solubilidad molar (mol L-1), pH 5 0,13 7,6E-6
Solubilidad molar (mol L-1), pH 7 6,67 7,6E-6
c) Los salicilatos considerados en este estudio fueron el 2-etilhexilsalicilato
(EHS), el homosalato (HMS) y el bencilsalicilato (BzS) que, aunque su
uso no está permitido en protectores solares, aparece en la formulación
de muchos cosméticos y perfumes. A continación se muestran sus
estructuras (Fig. 4) y sus propiedades físico-químicas (Tabla 5). Al igual
que las benzofenonas poseen un ligero carácter ácido; sin embargo, su
solubilidad en medio acuoso es muy inferior.
OH
O
O
EHS HMS BzSOH
O
O
Figura 4: Estructuras del EHS, HMS y BzS.
Introducción-Filtros solares
19
Tabla 5: Propiedades físico-químicas del EHS, HMS y BzS.
Propiedad Compuesto
EHS HMS BzS
Nº CAS 118-60-5 118-56-9 118-58-1
Peso molecular (g mol-1) 250,3 262,3 228,2
Densidad (g cm-3) 1,04 ± 0,06 1,1 ± 0,1 1,22 ± 0,06
pKa 8,1 ± 0,3 8,1 ± 0,3 8,1 ± 0,3
Entalpía de vaporización (KJ mol-1) 57 ± 3 61 ± 3 58 ± 3
logKow 5,8 ± 0,2 5,8 ± 0,3 4,0 ± 0,3
Presión de vapor (mTorr) 0,0807 0,0417 0,075
Punto de ebullición (ºC) 332 ± 15 341 ± 15 320 ± 0
Punto de fusión (ºC) 128 ± 13 132 ± 13 164 ± 14
Solubilidad molar (mol L-1), pH 1-6 1,1E-4 7,0E-5 7,6E-4
Solubilidad molar (mol L-1), pH 10 7,9E-3 5,4E-3 0,056
d) Los metoxicinamatos considerados fueron el isoamil-p-metoxicinamato
(IAMC) y el 2-etilhexil-p-metoxicinamato (EHMC) ambos permitidos por
la UE en concentraciones máximas individuales del 10%. Otros dos
filtros solares considerados en esta tesis son el octocrileno (OCR), de
estructura similar a los cinamatos y el 3-(4-Metilbencilideno alcanfor),
también conocido como Eusolex 6300 o por las siglas 4-MBC. Sus
estructuras químicas (Figura 5) y propiedades físico-químicas (Tabla 6) se
recogen a continación.
OCR4-MBC
IAMC EHMC
Figura 5: Estructuras del IAMC, EHMC, 4-MBC y OCR.
Introducción-Filtros solares
20
Tabla 6: Propiedades físico-químicas de los metoxicinamatos (IAMC y EHMC), OCR y
4-MBC.
Propiedad Compuesto
IAMC EHMC 4-MBC OCR
Nº CAS 71617-10-2 5466-77-3 36861-47-9 6197-30-4
Peso molecular (g mol-1) 248,3 290,4 254,4 361,5
Densidad (g cm-3) 1,03 ± 0,06 1,00 ± 0,06 1,06 ± 0,06 1,06 ± 0,06
Entalpía de vaporización (KJ mol-1) 61 ± 3 66 ± 3 62 ± 3 74 ± 3
logKow 4,1 ± 0,2 5,7 ± 0,2 4,9 ± 0,3 7,5 ± 0,8
Presión de vapor (mTorr) 0,0189 0,0889E-2 9,99E-3 2,56E-6
Punto de ebullición (ºC) 363 ± 17 405 ± 20 372 ± 22 478 ± 33
Punto de fusión (ºC) 152 ± 16 172 ± 16 167 ± 13 239 ± 12
Solubilidad molar (mol L-1), pH 1-10 2,4E-4 1,7E-5 1,4E-5 5,5E-7
Introducción-Filtros solares
21
2. DISTRIBUCIÓN MEDIOAMBIENTAL
Los filtros solares son usados en una amplia variedad de productos de
cuidado personal y compuestos farmacéuticos, por lo que entran en el medio
ambiente acuático indirectamente a través de las aguas residuales domésticas, y
directamente a través de las piscinas, playas, ríos y los efluentes de las plantas
de tratamiento de aguas residuales, donde no son eliminados de manera
completa [Straub, 2002] [Richardson, 2006].
Debido al elevado carácter lipofílico de algunos de estos compuestos
pueden acumularse en biota y matrices sólidas, alcanzando niveles similares a
los de PCBs y DDT [Heneweer, 2005]. Estudios recientes sobre la distribución
medioambiental de los filtros UV incluyen su determinación en agua de lagos
[Rodil, 2009-C] [Moeder, 2010] [Haunschmidt, 2010], ríos [Liu, 2010], efluentes e
influentes de plantas de tratamiento de aguas residuales urbanas [Balmer, 2005]
[Rodil, 2008-A] [Richardson, 2006]; así como, en sedimentos [Jeon, 2006] [Rodil,
2008-C] y lodos [Plagellat, 2006] [Rodil, 2009-D] [Nieto, 2009] [Wick, 2010]. Incluso
se encontraron residuos de estos analitos en pescado [Balmer, 2005] [Zenker,
2008] [Fent, 2010-B]. En esta tesis, además del desarrollo de metodología de
preparación de muestra para la determinación de filtros solares en algunas de
las matrices anteriores, se aportan datos relativos a su presencia en aguas
superficiales y residuales, muestras de lodo y polvo de atmósferas interiores.
Introducción-Filtros solares
22
2.1. Muestras de agua
2.1.1. Niveles de filtros solares en agua
Las concentraciones de filtros UV encontrados en agua varían según la
época del año, siendo mayores en los meses de verano, cuando el uso de cremas
para la protección solar es máximo. No obstante, se han detectado filtros solares
en muestras tomadas en otoño e invierno [Negreira, 2009-A], lo que demuestra
que estos compuestos son empleados en muchos productos de cuidado
personal, y no sólo en protectores solares. En muestras de agua superficial, los
niveles detectados varían desde los 5 ng L-1 en río [Rodil, 2008-B] hasta los 4000
ng L-1 en lago [Rodil, 2009-C]. En agua residual tratada, los niveles van desde 1
ng L-1 [Trenholm, 2008] hasta 4000 ng L-1 [Kasprzyk-Hordern, 2008]; en cambio en
agua residual sin tratar, las concentraciones son mayores abarcando desde los
25 ng L-1 para la BP-2 [Kasprzyk-Hordern, 2008] hasta los 20000 ng L-1 para el
EHMC [Kupper, 2006] [Balmer, 2005]. La diferencia de concentraciones entre
efluente e influente proporciona una primera estimación de la eficacia de
eliminación de estos compuestos en las plantas de tratamiento de agua residual.
De todos modos, estos valores deben considerarse con suma precaución ya que
en la mayoría de los estudios se hace referencia sólo a las concentraciones en
disolución, sin tener en cuenta la fracción adsorbida en el lodo. A continuación,
se muestra la revisión bibliográfica relativa a la distribución de filtros solares en
muestras de agua.
Balmer y col. [Balmer, 2005] fueron pioneros en la determinación de filtros
solares de naturaleza orgánica en aguas residuales y superficiales de lagos
suizos. Los análisis de aguas superficiales realizados en el año 2002 ofrecen
valores de 28 ng L-1 para 4-MBC y 35 ng L-1 para BP-3, mientras que en el año
1998 los valores fueron superiores a 82 ng L-1 para 4-MBC y 125 ng L-1 para BP-
3. Las concentraciones en aguas residuales sin tratar se encontraron en el rango
de 600 a 6500 ng L-1 para 4-MBC, de 700 a 7800 ng L-1 para BP-3, de 500 a 19000
Introducción-Filtros solares
23
ng L-1 para EHMC y de 200 a 12000 ng L-1 para OCR. En aguas residuales
tratadas, los valores fueron: de 60 a 2700 ng L-1 para 4-MBC, de 10 a 700 ng L-1
para BP-3, de 10 a 100 ng L-1 para EHMC y de 10 a 270 ng L-1 para OCR. Las
concentraciones de filtros solares en efluentes de plantas de tratamiento de
aguas residuales fueron considerablemente más bajas que las correspondientes
en influentes, indicando una eliminación significativa de los mismos. Los
porcentajes medios de eliminación fueron de 18-82% para 4-MBC, 68-99% para
BP-3, 88-99% para OCR y 97-99% para EHMC. Los análisis de muestras
tomadas en diferentes días indicaron que la eliminación variaba de día a día,
resultado de condiciones cambiantes en las estaciones depuradoras (descarga
de agua, tormenta, tiempo de residencia).
Li y col. [Li, 2007] cuantificaron BP-3, 4-MBC, EHMC y OCR en diferentes
unidades de una estación depuradora de agua residual en China que recibe
efluente de otra planta de la misma área y, una vez allí, sufre el tratamiento
terciario. La planta estudiada consta de los siguientes tratamientos:
coagulación-floculación, microfiltración y ozonización. Las concentraciones
encontradas en el influente fueron de 97 a 722 ng L-1 para BP-3, de 475 a 2128 ng
L-1 para 4-MBC, de 54 a 116 ng L-1 para EHMC y de 34 a 153 ng L-1 para OCR.
Las eficacias totales de eliminación en la planta fueron de 28-31% para BP-3, 37-
40% para 4-MBC, 40-43% para EHMC y 36-38% para OCR [Li, 2007]. Estos
porcentajes son relativamente más bajos que los obtenidos por Balmer y col.
[Balmer, 2005].
Poiger y col. [Poiger, 2004] determinaron concentraciones de filtros solares
a diferentes profundidades en los lagos Zurich y Hüttnersee antes, durante y
después del verano. Las concentraciones medidas fueron mayores en verano y
más altas en el lago Hüttnersee que en el lago Zurich. Los valores encontrados
en el lago Zurich estuvieron en el rango de 2 a 22 ng L-1 para 4-MBC, de 2 a 25
ng L-1 para EHMC y de 2 a 4 ng L-1 para BP-3. Los valores encontrados en
Introducción-Filtros solares
24
Hürtnersee se situaron en el rango de 5 a 125 ng L-1 para BP-3, de 2 a 82 ng L-1
para 4-MBC, de 2 a 27 ng L-1 para OCR y de 2 a 19 ng L-1 para EHMC.
Giokas y col. [Giokas, 2004] cuantificaron residuos de filtros solares en
diferentes muestras de agua. En agua de mar encontraron 2 ng L-1 de BP-3; en
agua de piscina, 4 ng L-1 de BP-3, 7 ng L-1 de 4-MBC y 5 ng L-1 de EHMC.
Lambropoulou y col. [Lambropoulou, 2002] también cuantificaron filtros
solares en el ambiente acuático. Las concentraciones obtenidas en agua de
piscina para BP-3 fueron de 2400 a 3300 ng L-1 y de 2100 ng L-1 para el PABA.
En residuos de ducha, los resultados fueron de 8200 a 9900 ng L-1 para BP-3 y de
5300 a 6200 ng L-1 para PABA.
Rodil y col. [Rodil, 2008-A] determinaron, en diferentes tipos de muestras
acuosas, los filtros solares: BP-4, BP-3, 4-MBC, IAMC, OCR y EHPABA. Los
valores encontrados en tres muestras diferentes de influente, tomadas en
estaciones depuradoras situadas en el Noroeste de España, fueron de 237 a 1481
ng L-1 para BP-4, de 31 a 168 ng L-1 para BP-3, 122 ng L-1 para 4-MBC y 36 ng L-1
para OCR. En los correspondientes efluentes se encontraron valores de 376 a
1947 ng L-1 para BP-4, 16 ng L-1 para BP-3, de 23 a 122 ng L-1 para 4-MBC, 20 ng
L-1 para OCR y 59 ng L-1 para IAMC. En aguas de río encontraron valores de
849 ng L-1 para BP-4 y 27 ng L-1 para BP-3. En dos muestras diferentes de agua
de mar sólo se encontró BP-4 en concentraciones de 38 a 138 ng L-1. De su
estudio se deduce que la BP-4 es uno de los filtros solares más difíciles de
eliminar en las estaciones depuradoras de aguas residuales, y también uno de
los más móviles en el medio acuático.
Kupper y col. [Kupper, 2006] encontraron concentraciones en influente
(agua residual sin tratar) para EHMC y OCR de 20070 y 1680 ng L-1,
respectivamente.
Introducción-Filtros solares
25
Cuderman y col. [Cuderman, 2007] investigaron la presencia de diferentes
filtros UV en 21 muestras de aguas recreacionales de Eslovenia encontrando
mayoritariamente BP-3 a niveles de 114 ng L-1 en agua de río, en el rango de 32
a 85 ng L-1 en lagos y en el rango de 103 a 400 ng L-1 en piscinas. Los demás
filtros solares no fueron encontrados a niveles superiores al límite de detección
del método.
Kasprzyk-Hordern y col. [Kasprzyk-Hordern, 2008] determinaron 4 filtros
solares de la familia de las benzofenonas (BP-1, BP-2, BP-3 y BP-4) con
concentraciones entre 4 ng L-1 para BP-2 y 227 ng L-1 para BP-4 en muestras de
agua de río, de 1 ng L-1 (BP-2) a 4309 ng L-1 (BP-4) en efluente y de 25 ng L-1 (BP-
2) a 5790 ng L-1 (BP-4) en influente.
Moeder y col. [Moeder, 2010] analizaron muestras de agua procedentes de
una planta de tratamiento de Liepzig y de un lago próximo a una zona de
recreo, utilizando diferentes métodos para la determinación de filtros solares.
Las concentraciones obtenidas fueron de 113 ng L-1 para 4-MBC, 385 ng L-1 para
BP-3, 362 ng L-1 para EHMC y de 440 ng L-1 para OCR en el agua de la planta de
tratamiento. Para la muestra de agua del lago se obtuvieron concentraciones de
2472 ng L-1 para 4-MBC, 62 ng L-1 para BP-3, 3686 ng L-1 para OCR y 150 ng L-1
para EHMC.
Haunschmidt y col. [Haunschmidt, 2010] tomaron muestras de agua en
lagos en los que se practican actividades recreativas, obteniendo
concentraciones de 32 a 40 ng L-1 para BP-3 y de 1400 a 1710 ng L-1 para OCR,
mientras que EHS, HMS, 4-MBC, EHPABA no fueron detectados.
Pedrouzo y col. [Pedrouzo, 2009] determinaron BP-1, BP-8, BP-3, OCR y
EHPABA en muestras de agua residual procedentes de Cataluña. Los valores
en efluente variaron de 20 a 100 ng L-1 para BP-3, 19 ng L-1 para EHPABA, 11 ng
L-1 para BP-1 y en influente, de 11 a 286 ng L-1 para BP-3, 103 ng L-1 para
Introducción-Filtros solares
26
EHPABA y de 47 a 155 ng L-1 para BP-1. En un trabajo posterior [Pedrouzo,
2010], los mismos autores determinaron BP-8, BP-3, OCR y EHPABA en
muestras de agua superficial y residuales. En agua de río sólo detectaron BP-3
con valores desde 6 a 28 ng L-1. En efluente, las concentraciones encontradas
fueron de 25 y 55 ng L-1 para EHPABA y BP-8, respectivamente. En cambio, en
influente todos los compuestos fueron detectados con concentraciones en el
rango de 59 a 185 ng L-1 para BP-8, de 75 a 127 ng L-1 para BP-3, 129 ng L-1 para
OCR y de 55 ng L-1 para EHPABA.
Tarazona y col. [Tarazona, 2010] desarrollaron un método para la
determinación de benzofenonas hidroxiladas en 3 muestras de agua de mar. Las
concentraciones encontradas fueron de 280 ng L-1 para BP-1 y de 1340 a 3300 ng
L-1 para BP-3.
Wick y col. [Wick, 2010] determinaron 4 benzofenonas (BP-1, BP-2, BP-3 y
BP-4) en muestras acuosas. Las concentraciones medidas fueron de 2 ng L-1 (BP-
1) a 1980 ng L-1 (BP-4) en agua superficial, de 12 ng L-1 (BP-1) a 572 ng L-1 (BP-4)
en efluente y de 35 ng L-1 (BP-2) a 5130 ng L-1 (BP-4) en influente.
Liu y col. [Liu, 2010] determinaron 4 filtros solares (EHS, BP-3, 4-MBC y
OCR) en agua de río. Las concentraciones encontradas oscilaron entre 8 ng L-1
para EHS hasta 59 ng L-1 para la BP-3.
A continuación, se muestra una Tabla resumen (Tabla 7) donde se
recogen las concentraciones de filtros solares encontrados en la bibliografía para
diferentes muestras de agua superficial (agua de río, lago, mar y piscina) y agua
residual tratada (efluente) y sin tratar (influente). En aquellos trabajos en los
que se procesan varias muestras del mismo tipo, se indican las concentraciones
mínimas y máximas encontradas.
Introducción-Filtros solares
27
Tabla 7: Niveles de filtros solares en muestras acuosas.
Tipo de
agua Compuestos detectados Concentración (ng L-1) Referencia
Río
BP-1 47 [Jeon, 2006]
BP-3 23 [Kawaguchi, 2006]
BP-3 114 [Cuderman, 2007]
BP-3 14 [Kawaguchi, 2008-A]
BP-3 85 [Okanouchi,2008]
BP-4, BP-3 27 (BP-3)-849 (BP-4) [Rodil, 2008-A]
EHPABA 5 [Rodil, 2008-B]
BP-1, BP-2, BP-3, BP-4 4 (BP-2)-227 (BP-4) [Kasprzyk-Hordern,
2008]
BP-3 6-28 [Pedrouzo, 2010]
BP-1, BP-2, BP-3, BP-4 1 (BP-1)-51 (BP-4) [Wick, 2010]
EHS, BP-3, 4-MBC, OCR 8 (EHS)-59 (BP-3) [Liu, 2010]
Arroyo BP-1, BP-2, BP-3, BP-4 2 (BP-1)-1980 (BP-4) [Wick, 2010]
Lago
4-MBC, EHMC, BP-3 2-125 (BP-3) [Poiger, 2004]
4-MBC, BP-3 28 (4-MBC)-125 (BP-3) [Balmer, 2005]
EHS, HMS, BP-3,
4-MBC, IAMC, EHMC,
EHPABA, OCR
2 (EHPABA)-250 (OCR) [Rodil, 2008-B]
EHS, BP-3, 4-MBC, IAMC,
EHMC, EHPABA, OCR 40 (BP-3)-4381 (OCR) [Rodil, 2009-C]
BP-3 32-85 [Cuderman, 2007]
4-MBC, BP-3, EHMC, OCR 62 (BP-3)-3686 (OCR) [Moeder, 2010]
BP-3, OCR 32 (BP-3)-1710 (OCR) [Haunschmidt, 2010]
Mar
BP-3 2 [Giokas, 2004]
BP-4 38-138 [Rodil, 2008-A]
BP-3, BP-1 280 (BP-1)-3300 (BP-3) [Tarazona, 2010]
Piscina
BP-3, PABA 2100 (PABA)-3300 (BP-3) [Lambropoulou, 2002]
4-MBC, EHMC, BP-3 4 (BP-3)-7 (4-MBC) [Giokas, 2004]
BP-3 103-400 [Cuderman, 2007]
IAMC 700 [Vidal, 2010]
Introducción-Filtros solares
28
Tabla 7 cont.: Niveles de filtros solares en muestras acuosas.
Tipo de
agua Compuestos detectados Concentración (ng L-1) Referencia
Residual
tratada
(efluente)
4-MBC, BP-3, EHMC, OCR 10 (OCR)-2700 (4-MBC) [Balmer, 2005]
BP-3 1-13 [Trenholm, 2008]
BP-4, BP-3, 4-MBC,
OCR, IAMC 16 (BP-3)-1947 (BP-4) [Rodil, 2008-A]
EHS, HMS, BP-3,
4-MBC, IAMC, EHMC,
EHPABA, OCR
2 (EHPABA)-54 (BP-3) [Rodil, 2008-B]
BP-3, 4-MBC, OCR,
EHMC, EHS 32 (EHS)-899 (EHMC) [Rodil, 2009-E]
BP-3, 4-MBC, OCR 18 (BP-3)-179(OCR) [Rodil, 2009-C]
BP-1, BP-3, EHPABA 11 (BP-1)-100 (BP-3) [Pedrouzo, 2009]
BP-8, EHPABA 25 (EHPABA)-55 (BP-8) [Pedrouzo, 2010]
BP-1, BP-2, BP-3, BP-4 1 (BP-2)-4309 (BP-4) [Kasprzyk-Hordern,
2008]
4-MBC, BP-3, EHMC, OCR 113 (4-MBC)-440 (OCR) [Moeder, 2010]
BP-1, BP-2, BP-3, BP-4 12 (BP-1)-572 (BP-4) [Wick, 2010]
Residual
sin tratar
(influente)
4-MBC, BP-3, EHMC, OCR 200 (OCR)-19000
(EHMC) [Balmer, 2005]
EHMC, OCR 1680 (OCR)-20070
(EHMC) [Kupper, 2006]
4-MBC, BP-3, EHMC, OCR 34 (OCR)-2128 (4-MBC) [Li, 2007]
BP-1, BP-2, BP-3, BP-4 25 (BP-2)-5790 (BP-4) [Kasprzyk-Hordern,
2008]
BP-3 300-2300 [Trenholm, 2008]
BP-4, BP-3, 4-MBC, OCR 31 (BP-3)-1481 (BP-4) [Rodil, 2008-A]
BP-3, 4-MBC, OCR,
IAMC, EHMC, EHS 69 (IAMC)-899 (EHMC) [Rodil, 2009-E]
BP-3, IAMC, 4-MBC,
OCR, EHMC, EHS 66 (IAMC)-5322 (OCR) [Rodil, 2009-C]
BP-1, BP-3, EHPABA 11 (BP-3)-286 (BP-3) [Pedrouzo, 2009]
BP-1, BP-3, OCR, EHPABA 55 (EHPABA)-185 (BP-1) [Pedrouzo, 2010]
BP-1, BP-2, BP-3, BP-4 35 (BP-2)-5130 (BP-4) [Wick, 2010]
Introducción-Filtros solares
29
En general, BP-3, 4-MBC, EHMC y OCR son los filtros solares
considerados en la mayoría de los estudios. Además, en los últimos años se
incluyen la BP-4 y los metabolitos de la BP-3. En lo relativo a su
comportamiento en estaciones depuradoras, BP-4 y 4-MBC son las especies que
presentan los porcentajes de eliminación más bajos, considerando sus
concentraciones en disolución a la entrada y salida de las plantas depuradoras.
En esta Tesis doctoral se aportan datos de distribución de filtros solares en río,
piscina y agua residual [Negreira, 2009-A; Negreira, 2009-B; Negreira, 2010-A],
siendo el valor más alto encontrado de 26000 ng L-1 para el OCR en agua de
piscina [Negreira, 2010-A].
2.1.2. Toxicidad y reacciones de transformación de los
filtros solares en el medio acuoso
La presencia de filtros UV en el medio acuático se ha relacionado con
alteraciones encontradas en los peces tales como: inducción de vitelogenina
(proteína de la yema del huevo, importante para la reproducción), alteraciones
de las gónadas, disminución de la fertilidad y de la tasa reproductiva y la
feminización de las características sexuales de los machos [Díaz-Cruz, 2009].
Existen estudios in vivo e in vitro que evidencian este problema y sugieren un
riesgo potencial para la salud humana. Además estos compuestos son
bioacumulativos y capaces de penetrar la barrera cutánea. Así se ha detectado
la presencia de BP-3 y EHMC en leche humana [Hany, 1995].
Algunos filtros solares presentan mayor o menor actividad disruptora
como el EHMC [Kunz, 2006-A; Kunz, 2006-B] [Seidlová-Wuttke, 2006] [Straub,
2002] [Díaz-Cruz, 2009], el 4-MBC [Maerkel, 2005] [Seidlová-Wuttke, 2006]
[Schlumpf, 2008-A; Schlumpf, 2008-B] [Díaz-Cruz, 2009], la BP-4 [Fent, 2010-A], la
BP-3 [Díaz-Cruz, 2009], la BP-1 [Kunz, 2006-B] [Fent, 2010-A] y la BP-2 [Kunz,
2006-B] [Fent, 2010-A]; de los cuales, el EHMC, 4-MBC y BP-3 ya han sido
englobados en el grupo de los productos químicos considerados como
estrogénicos [Díaz-Cruz, 2009].
Introducción-Filtros solares
30
Los estudios de toxicidad in vivo se han realizado administrando filtros
UV en la comida de roedores a la generación de los padres (antes del
apareamiento, durante el embarazo y la lactancia) y a las crías hasta la edad
adulta. Los machos recién nacidos, expuestos a filtros UV, exhibían alteraciones
en el crecimiento de la próstata y las hembras en la expresión de los genes del
útero [Díaz-Cruz, 2009] [Schlumpf, 2008]. También se apreciaron interacciones
del 4-MBC con el tiroides, con el mARN y los niveles de proteínas en el útero,
próstata y cerebro de los adultos [Schlumpf, 2008]. Además, el PABA y sus
derivados, pueden incrementar la citotoxicidad bacteriana e interactuarían con
el ADN tras la radiación UV potenciando la fotocarcinogénesis [Debuys, 2000].
Una línea de investigación de gran actualidad, en relación con la presencia
de contaminantes emergentes en el medio acuático, se centra en evaluar sus
posibles reacciones de transformación, en la identificación de los productos que
se generan (by-products) y en la estimación de su estabilidad, persistencia y
toxicidad en el medio acuático [Richardson, 2003]. Un porcentaje importante de
estas reacciones son de tipo fotoquímico o redox. Las segundas, normalmente
implican procesos de oxidación de los analitos por diversos agentes
introducidos de forma intencionada, o no, en el medio acuático. Entre ellos, el
más habitual es el cloro libre (combinación de ácido hipocloroso e hipoclorito
sódico). En algunos casos los productos de cuidado personal reaccionan con el
cloro libre produciendo especies volátiles fácilmente eliminables. Este tipo de
reacciones pueden tener interés a la hora de proponer tratamientos terciarios,
que mejoren el porcentaje de eliminación de estos compuestos en las estaciones
depuradoras de aguas residuales [Huber, 2005]. Sin embargo, en ocasiones, los
subproductos formados en los procesos de oxidación pueden resultar más
persistentes y tóxicos que los analitos de partida; además, algunas de estas
reacciones pueden ser muy favorables desde el punto de vista cinético [Gallard,
2002].
Introducción-Filtros solares
31
La estabilidad de los filtros solares en presencia de luz y de agentes
oxidantes, empleados en procesos de desinfección de agua potable o residual,
es un aspecto poco estudiado en la bibliografía; no obstante, atendiendo a la
presencia de grupos fenólicos o amino en su estructura parece factible que
puedan experimentar reacciones de fotodescomposición y/o oxidación. Se han
encontrado dos trabajos en relación con su fotodegradación [Sakkas, 2003] [Rodil,
2009-A]. Sakkas y col. [Sakkas, 2003] se centraron en el éster octilado del PABA,
describiendo la formación de productos de metilación del grupo amino e
hidroxilaciones en el anillo aromático debido a reacciones de tipo fotoquímico.
Las reacciones siguieron una cinética de primer orden, y la vida media del filtro
UV se encontró entre 27 y 39 horas, considerando radiación solar natural y
distintos tipos de agua. Rodil y col. [Rodil, 2009-A] estudiaron la fotoestabilidad
de 6 filtros solares (OCR, BP-3, 4-MBC, IAMC, EHMC y EHPABA) en
disolución acuosa y demostraron que EHMC, IAMC y EHPABA se degradaban
siguiendo cinéticas de primer orden, con vidas medias comprendidas entre 20 y
59 horas. Estos autores consiguieron identificar dos derivados de la
fotodegradación del EHPABA como consecuencia de la pérdida de los grupos
metilo unidos al átomo de nitrógeno, sin embargo, no fueron capaces de
detectar ningún fotoproducto del IAMC ni del EHMC.
En cuanto a la estabilidad de los filtros solares en presencia de agentes
oxidantes, Sakkas y col. [Sakkas, 2003] han descrito la presencia de derivados
clorados del octil PABA en agua de piscina. Puesto que estos compuestos no
tienen ninguna aplicación comercial conocida, se supone que son el resultado
de la reacción entre el cloro existente en el agua y el éster del PABA, presente en
los protectores solares que usan los bañistas. En cualquier caso, en el artículo
anterior, no se aportan datos sobre las vidas medias ni sobre mecanismos de
formación de estas especies.
En esta Tesis doctoral, se estudió la estabilidad en agua clorada de tres
filtros UV: EHPABA, EHS y BP-3. Se demostró que la BP-3 y el EHPABA
Introducción-Filtros solares
32
presentaban una elevada reactividad en muestras de agua conteniendo
concentraciones de cloro libre similares a las utilizadas en agua de grifo o de
piscina [Negreira, 2008].
2.2. Lodos y sedimentos
Muchos filtros solares presentan elevados coeficientes de partición
octanol/agua (Kow), por lo que es de esperar su acumulación en sedimentos,
lodos y materia particulada. Los niveles detectados en sedimentos son bajos. En
general, no superan los 16 ng g-1 [Ricking, 2003] [Jeon, 2006] con excepción de
Rodil y col. [Rodil, 2008-C] que encontraron valores de 14 a 34 ng g-1 para
EHMC y de 61 a 93 ng g-1 para OCR. Sin embargo, los valores encontrados en
lodos son mucho más elevados, destacando la presencia de OCR (el más
lipofílico de los filtros solares determinado en muestras medioambientales,
logKow 7,5) que alcanza valores de 18000 ng g-1 [Plagellat, 2006]. La determinación
de filtros solares en lodos de depuradoras es un aspecto de interés para calcular
el riesgo de re-introducir estos compuestos en el medio terrestre debido al uso
de lodos como fertilizantes en agricultura, así como para calcular sus
porcentajes reales de eliminación en las estaciones depuradoras.
Plagellat y col. [Plagellat, 2006] determinaron filtros solares en lodos de
estaciones depuradoras de aguas residuales urbanas en Suiza, encontrando
concentraciones en invierno y verano de 1730 y 1820 ng g-1 para 4-MBC, 105 y
115 ng g-1 para EHMC, 4270 y 5410 ng g-1 para OCR, respectivamente.
Nieto y col. [Nieto, 2009] analizaron muestras de lodos de depuradora que
fueron previamente homogeneizadas, congeladas y liofilizadas utilizando el
sistema de secado por congelación. Se calculó la concentración de filtros solares
en distintas épocas del año obteniéndose valores para BP-3 entre 10 y 20 ng g-1,
entre 700 y 1842 ng g-1 para OCR y de 132 a 170 ng g-1 para EHPABA, referidos
Introducción-Filtros solares
33
a peso seco. En el estudio anterior no se contempla la determinación de 4-MBC
ni de EHMC.
Rodil y col. [Rodil, 2009-D] determinaron los niveles de varios filtros UV en
una muestra de lodos de una estación depuradora en Alemania, que recibe el
agua residual de una población de 100.000 habitantes, y en un material de
referencia (MR) de lodos obtenidos del IRMN (Centro Común de Investigación
de la Comisión Europea, Geel, Bélgica). En el primer caso, las mayores
concentraciones correspondieron a OCR, EHS y HMS; mientras que en el MR,
los niveles máximos corresponden a 4-MBC seguido de OCR.
Wick y col. [Wick 2010] determinaron 4 compuestos de la familia de las
benzofenonas (BP-1, BP-2, BP-3 y BP-4) en muestras de lodo recogidas en una
planta de tratamiento de aguas residuales provenientes de una población de
320.000 habitantes en Alemania. Las concentraciones encontradas fueron bajas
desde 5 ng g-1 para la BP-1 hasta 132 ng g-1 para la BP-3.
A continuación, se muestran dos tablas resumen donde se recogen las
concentraciones de filtros solares encontrados en muestras de sedimentos (Tabla
8) y lodos (Tabla 9). En conjunto, la información disponible es muy limitada en
comparación con el estudio de su distribución en muestras de agua.
Introducción-Filtros solares
34
Tabla 8: Niveles de filtros solares en sedimentos.
Tipo de muestra Localización Compuestos
determinados
Concentración
(ng g-1) Referencia
Sedimento
( río) Alemania
BP
EHMC
0,5-3
4 [Ricking, 2003]
Sedimento Korea
BP
BP-3
BP-1
BP-8
2-10
nd
nd
0,5-2
[Jeon, 2006]
Suelo Korea
BP
BP-3
BP-1
BP-8
0,03-17
0,03-4
nq
0,5-4
[Jeon, 2006]
Sedimento
(lago) Alemania
EHMC
OCR
14-34
61-93 [Rodil, 2008-C]
n.d., no detectado
n.q., no cuantificado
Introducción-Filtros solares
35
Tabla 9: Niveles de filtros solares en lodos.
Tipo de muestra Localización
Compuestos determinados
Concentración (ng g-1) Referencia
Lodo de aguas residuales
Suiza 4-MBC EHMC OCR
150-4980 10-390 320-18740
[Plagellat, 2006]
Lodo de aguas residuales
España
BP-3 OCR EHPABA BP-1 BP-8
10-20 700-1842 132-170 nd nd
[Nieto, 2009]
Lodo IRMN
-
BP-3 IAMC 4-MBC OCR EHPABA EHMC EHS HMS
6,6 5,0 3893 2479 1,4 127 49 22
[Rodil, 2009-D]
Lodo de aguas residuales
Alemania (Leipzig)
BP-3 IAMC 4-MBC OCR EHPABA EHMC EHS HMS
29 20 73 585 1,9 35 280 331
[Rodil, 2009-D]
Lodo de aguas residuales
Alemania
BP-1 BP-2 BP-3 BP-4
5,1 11 132 29
[Wick, 2010]
n.d., no detectado
En la bibliografía, no aparecen datos relativos a la distribución de filtros
solares en atmósferas interiores. No obstante, en esta tesis doctoral [Negreira,
2009-C] se analizaron varias muestras de polvo procedentes de viviendas y de
vehículos.
Introducción-Filtros solares
36
2.3. Biota
Balmer y col. [Balmer, 2005] determinaron algunos filtros solares de
naturaleza orgánica en diferentes tipos de pescado de agua dulce encontrando
valores de 44 a 166 ng g-1 para 4-MBC, de 66 a 123 ng g-1 para BP-3, 64 a 72 ng g-
1 para EHMC y 25 ng g-1 para OCR, referidos a materia grasa. En la misma
línea, Buser y col. [Buser, 2006] también determinaron 4-MBC y OCR en pescado
de ríos suizos. Las concentraciones encontradas se situaron en el rango de 50 a
2400 ng g-1 en pescado de río y de 20 a 170 ng g-1 en pescado de lago, ambos
datos referidos a materia grasa.
Meinerling y col. [Meinerling, 2006] determinaron BP-3, 4-MBC, OCR y
EHMC en pescado y encontraron concentraciones en el rango de 3 a 21 ng g-1
para la BP-3, 3 ng g-1 para 4-MBC, 4 ng g-1 para OCR y 6 ng g-1 para EHMC,
referidas a músculo de pescado liofilizado.
Zenker y col. [Zenker, 2008] validaron un método para la determinación de
7 filtros solares (BP-1, BP-2, BP-3, BP-4, EHMC, Et-PABA y 4-MBC) en muestras
de pescado recogidas en diferentes puntos del río Glatt y sus afluentes, cerca de
Zürich, Suiza. Sólo detectaron la presencia de EHMC con una concentración de
42 ng g-1 en la muestra perteneciente a la salida de un lago y 142 ng g-1 a la
salida de una planta de agua residual, ambas referidas a contenido graso. En la
misma línea, Fent y col. [Fent, 2010-B] también encontraron EHMC en pescado
perteneciente al mismo río con concentraciones en el intervalo de 49 a 129 ng g-1
referidas a contenido graso.
A continuación, se muestra una tabla resumen (Tabla 10) donde se
recogen las distintas concentraciones de filtros solares en pescado. En general,
la distribución y acumulación de filtros solares en biota, ha sido estudiada en
áreas geográficas muy concretas y los datos disponibles parecen todavía
insuficientes para evaluar su potencial de bioacumulación en las cadenas
tróficas.
Introducción-Filtros solares
37
Tabla 10: Niveles de filtros solares encontrados en pescado.
Tipo de pescado Localización Compuestos
detectados
Concentración
(ng g-1) Referencia
Pescado blanco
(Coregonus sp.)
Rutilo
(Rutilus rutilus)
Perca
(Perca fluviatilis)
Suiza
(lago)
4-MBC
BP-3
EHMC
OCR
44-166
66-123
64-72
25
[Balmer, 2005]
Trucha marrón
(Salmo trutta fario)
Suiza
(río)
4-MBC
OCR
50-2400
20-170 [Buser, 2006]
Trucha marrón
(S. trutta fario)
Suiza
(lago)
4-MBC
OCR
<20-170
nd [Buser, 2006]
Trucha arco iris Alemania
(río)
BP-3
4-MBC
OCR
EHMC
3-21
3
4
6
[Meinerling,
2006]
Bagre o cacho
(Leuciscus cephalus)
Lengüeta
(Barbus barbus)
Suiza
(río) EHMC
42-142 [Zenker, 2008]
49-129 [Fent, 2010-B]
n.d., no detectado
Introducción-Filtros solares
38
3. METODOLOGÍA ANALÍTICA
3.1. Preparación de muestra
La preparación de muestra es una etapa crítica en cualquier
procedimiento analítico. Suele incluir una serie de pasos como son: la extracción
de los analitos desde la matriz de la muestra, la concentración de los analitos
hasta un nivel medible, la eliminación de especies interferentes o la conversión
química de los analitos a otras formas más fácilmente detectables [Cela, 2002].
En esta Tesis Doctoral se determinaron varios filtros solares en distintas
matrices tanto sólidas (polvo y lodo), como acuosas, empleando diferentes
técnicas de preparación de muestra cuyo fundamento se describe a
continuación. También se presenta una revisión bibliográfica de las
metodologías usadas para la determinación de los compuestos objeto de
estudio.
3.1.1. Muestras de agua
Los métodos de preparación de muestras acuosas utilizados en esta Tesis
Doctoral han sido: extracción en fase sólida (SPE), microextracción en fase
sólida (SPME), microextracción líquido-líquido dispersiva (DLLME) y
microextracción en fase sólida con siliconas en formatos no comerciales.
3.1.1.1. Extracción en fase sólida (SPE)
La SPE es una técnica de uso extendido que permite transferir los analitos
de la muestra líquida a una fase sólida (retención) y recuperarlos
cuantitativamente con un disolvente adecuado (elución). En este proceso se
logra separar los analitos de los compuestos interferentes (purificación) y
concentrarlos en un volumen pequeño de disolvente. La SPE es la técnica de
referencia para la concentración de filtros solares en muestras acuosas,
Introducción-Filtros solares
39
encontrando numerosas aplicaciones en la bibliografía que se detallan a
continuación.
Poiger y col. [Poiger, 2004] cuantificaron cinco filtros solares en agua de río
y de baño. La preparación de la muestra fue llevada a cabo mediante SPE. El
volumen de muestra usado fue de 1 litro con adición de bencilcinamato, como
surrogado interno, a un nivel de 100 ng L-1. Las extracciones fueron llevadas a
cabo con columnas de vidrio reutilizables que contenían un adsorbente
macroporoso de poliestireno. Posteriormente, los analitos fueron cuantificados
por GC-MS, empleando un espectrómetro de masas equipado con un
analizador de sector magnético, bajo ionización de impacto electrónico (EI, 70
eV, 200 ºC). La columna usada fue de tipo apolar. Los límites de detección
alcanzados fueron de 2 ng L-1 y las recuperaciones obtenidas variaron entre 77%
para 4-MBC, 90% para EHMC, 64% para BP-3 y 57% para OCR.
Giokas y col. [Giokas, 2004] utilizaron la SPE para la extracción de filtros
solares en diferentes muestras de agua (piscina, residuos de ducha de hoteles).
La etapa de concentración se llevó a cabo sobre discos de C18 de 500 mg (47
mm de diámetro). El volumen de muestra fue de 500 mL. La elución se realizó
con acetato de etilo:diclorometano (1:1, v/v) recogiendo dos alícuotas de 5 mL
que fueron evaporadas y reconstituidas con 0,010 mL de hexano para la
posterior determinación por GC-MS, y con 0,050 mL de metanol para su
inyección en cromatografía líquida. La inyección de 3 µL de extracto en GC-MS
proporcionó recuperaciones en el rango de 95 a 99%, y límites de cuantificación
de 0,9 a 1,4 ng L-1. La determinación mediante cromatografía líquida, con
detector UV-VIS, alcanzó límites de cuantificación entre 8 y 24 ng L-1.
Sakkas y col. [Sakkas, 2003] usaron SPE para estudiar la presencia de
EHPABA en diferentes muestras acuosas. Las muestras (5 mL) fueron pasadas a
través de cartuchos C18. En la etapa de elución se empleó
diclorometano:acetato de etilo en proporción 1:1 recogiendo dos alícuotas de 3
Introducción-Filtros solares
40
mL que, después de combinadas, fueron evaporadas y redisueltas con 0,1 mL
de metanol. La determinación se realizó mediante cromatografía líquida. Las
recuperaciones obtenidas fueron de 98% para agua de mar, 101% para agua de
piscina y de 98% para agua destilada y el límite de detección se situó en 300 ng
L-1.
Balmer y col. [Balmer, 2005] determinaron tres filtros solares (BP-3, 4-MBC,
EHMC) en aguas residuales, superficiales y en pescado de lagos suizos. La
preparación de la muestra se realizó mediante SPE, empleando columnas de
vidrio reutilizables que contenían aproximadamente 10 mL de un polímero
adsorbente y macroporoso de poliestireno divinilbenzeno. El volumen de
muestra fue de 300 mL y el flujo a través de los cartuchos se ajustó a 10 mL min-
1. La elución se llevó a cabo con metanol:diclorometano y la limpieza posterior
de los extractos se realizó con sílica. Seguidamente, los extractos fueron
concentrados con corriente de nitrógeno y analizados mediante GC-MS. Las
recuperaciones se encontraron en el rango de 78 a 129% y los límites de
detección fueron de 2 ng L-1 para aguas superficiales y 10 ng L-1 para aguas
residuales.
Cuderman y col. [Cuderman, 2007] usaron la SPE como método de
concentración de diferentes filtros solares (HMS, 4-MBC, BP-3, OCR y EHMC)
en diversas muestras acuosas. Los cartuchos empleados fueron Strata X (60 mg).
El volumen de muestra fue de 500 mL ajustado a pH 3. Una vez finalizada la
etapa de concentración, se lavó el cartucho con 1,5 mL de agua que contenía 1%
de metanol para eliminar impurezas. Después del secado, se eluyeron los
analitos con acetato de etilo:diclorometano, en proporción 1:1 (v/v). Los
extractos (3 x 0,5 mL) fueron combinados y evaporados a sequedad bajo
corriente de nitrógeno, antes de su redisolución con 0,4 mL de tolueno. Las
recuperaciones obtenidas se encontraron en el rango de 82 a 98% para agua
desionizada y de 50 a 98% en agua superficial.
Introducción-Filtros solares
41
Li y col. [Li, 2007] determinaron BP-3, 4-MBC, EHMC y OCR en muestras
de agua residual usando SPE como método de concentración. Volúmenes de 1
litro de muestra fueron acidificados a pH 3 y se les adicionó un 1% de metanol,
v/v. Los cartuchos empleados fueron de C18, acondicionados previamente con
5 mL de acetato de etilo:diclorometano en proporción 1:1 (v/v), 5 mL de
metanol y 5 mL de agua desionizada. Los cartuchos se secaron durante 5
minutos aplicando vacío y se eluyeron con alícuotas de 5 ml de acetato de
etilo:diclorometano (1:1, v/v). Los extractos fueron evaporados y redisueltos
con 1 mL de n-hexano. La determinación posterior se llevó a cabo mediante GC-
MS. Las recuperaciones obtenidas se encontraron en el rango de 67 a 118% y el
límite de detección fue de 10 ng L-1.
Rodil y col. [Rodil, 2008-A] presentaron un nuevo método analítico basado
en SPE y LC-(ESI)MS/MS para la determinación de seis filtros solares (BP-4,
BP-3, 4-MBC, IAMC, EHPABA y OCR) en muestras acuosas. Los adsorbentes
considerados fueron Oasis HLB (60 y 200 mg), Sep-Pak Plus C18 (aprox. 360
mg) y Bond Elut Plexa (60 y 200 mg). De ellos, los que aportaron mejores
recuperaciones fueron los Oasis HLB (60 mg). Estos fueron acondicionados con
3 mL de metanol, 3 mL de agua pura y 3 mL de disolución tampón. El método
optimizado consistió en adicionar 20 mL de una disolución de tampón (2%
metanol, 50 mM tri-n-butilamina ajustada a pH 4,5 con ácido fórmico) a una
muestra de 200 mL (pH 2) de agua con objeto de obtener el par iónico de la BP-
4, facilitando así su retención en el cartucho de fase reversa. Una vez pasada la
muestra a través de los cartuchos se lavaron con 3 mL de tampón y 3 mL de
Milli-Q y, posteriormente, se secaron durante 30 min con nitrógeno. La elución
se hizo con metanol (3 x 10 mL) y los extractos combinados (30 mL aprox.)
fueron concentrados y ajustados a un volumen final de 1 mL con metanol:agua
(1:1). Se estudiaron también otros disolventes de elución como acetonitrilo y
acetona con recuperaciones más bajas. Esta metodología ofrece límites de
detección en el rango de 7 a 46 ng L-1 y recuperaciones en el rango de 63 a 108%.
El mismo método es propuesto para la determinación de filtros solares junto
Introducción-Filtros solares
42
con herbicidas, organofosforados usados como plastificantes y retardantes de
llama, y otros compuestos farmacéuticos pero, en este caso se emplearon los
cartuchos Oasis HLB de 200 mg [Rodil, 2009-B].
Trenholm y col. [Trenholm, 2008] desarrollaron un método para la
determinación de 24 contaminantes emergentes, entre ellos la BP, la BP-3 y el
alcanfor en aguas residuales empleando SPE, como técnica de concentración, y
GC-MS/MS, como técnica de determinación para el alcanfor y BP y LC-MS/MS
para la BP-3. Para la determinación por GC-MS/MS, 500 mL de muestra fueron
concentrados a través de cartuchos HLB (200 mg) previamente acondicionados
con 5 mL de diclorometano, 5 mL de metil-t-butil éter (MTBE), 5 mL de metanol
y 5 mL de agua ultrapura. Los surrogados adicionados fueron 0,25 µg de [13C6]-
o-fenilfenol, BHT-d24, dibutilftalato-d4 y 1 µg de [13C6]-vanilina. Después de
pasar la muestra a través de los cartuchos a un flujo de 15 mL/min, estos se
lavaron con 5 mL de agua ultrapura y se secaron bajo corriente de nitrógeno
durante 30 min. La elución se llevó a cabo con 5 mL de metanol:MTBE 10:90
(v/v), seguido de 5 mL de diclorometano. Los eluatos fueron combinados,
concentrados y reconstituidos con 0,5 mL de isooctano. El procedimiento de
SPE desarrollado para LC-MS/MS fue muy similar al de GC salvo algunas
diferencias: los cartuchos fueron acondicionados con 5 mL de MTBE, 5 mL de
metanol y 5 mL de agua. Los surrogados adicionados fueron 0,02 µg de [13C6]-
triclosán, [13C6]-simazina, [13C6]-atrazine y 0,1 µg de [13C6]-bisfenol A y [13C6]-o-
fenilfenol. La elución de los cartuchos se llevó a cabo con 5 mL de
metanol:MTBE 10:90 (v/v) y estos se concentraron a 500 µL. El método
proporciona límites en el rango de 1 a 50 ng L-1 en agua y recuperaciones entre
67 y 87%.
Kasprzyk-Hordern y col. [Kasprzyk-Hordern, 2008] desarrollaron un método
para la determinación de una gran variedad de compuestos, entre los cuales se
encuentran BP-1, BP-2, BP-3 y BP-4, en muestras de agua superficial y residual.
Los adsorbentes de SPE probados fueron Oasis HLB, MCX, MAX, WCX, WAX
Introducción-Filtros solares
43
(60 mg), Chormabond C18 (200 mg), Isolute ENV+ (100 mg) e Isolute HCX (200
mg). De todos ellos, los cartuchos Oasis MCX resultaron ser los más eficientes.
El volumen de muestra empleado fue de 1000 mL para agua superficial y 250
mL para agua residual. Las muestras se acidificaron a pH 2 y la elución tuvo
lugar con 2 mL de metanol y 2 mL de metanol con un 5% de amoníaco. El
extracto resultante se concentró y se reconstituyó con 0,5 mL de fase móvil,
para su posterior determinación mediante UPLC-(ESI)MS-MS. Los límites de
cuantificación del método se situaron en el rango de 0,1 ng L-1 (agua superficial)
a 35 ng L-1 (agua residual sin tratar) y las recuperaciones obtenidas variaron
desde 17 a 117%.
Pedrouzo y col. [Pedrouzo, 2009] presentan un método basado en SPE y
UPLC-(ESI)MS/MS para la determinación de varios productos de cuidado
personal, entre ellos 5 filtros solares (BP-1, BP-3, BP-8, OCR y EHPABA), en
muestras acuosas. La SPE fue llevada a cabo con 2 adsorbentes poliméricos:
Oasis HLB (500 mg) para agua residual y Bond Elut Plexa (200 mg) para agua
de río. Ambos adsorbentes fueron acondicionados con 5 mL de metanol y 2 mL
de milli-Q. Los volúmenes de muestra extraídos fueron de 100 mL para
influente, 250 mL para efluente y 500 mL para agua de río. Las muestras fueron
pasadas a través de los cartuchos a un flujo de 10-15 mL min-1. A continuación
se lavaron con agua conteniendo un 15% de metanol y se secaron durante 5
min, antes de ser eluidos con 5 mL de metanol y 5 mL de diclorometano. Los
extractos fueron concentrados a 3-4 mL y posteriormente diluidos a 5 mL con
Milli-Q para inyectar un volumen de 50 µL en el sistema cromatográfico. Las
recuperaciones obtenidas para los filtros solares en 500 mL de agua de río
oscilaron entre 46 y 97%. Para aguas residuales, las recuperaciones estuvieron
en el rango de 20 a 71% (250 mL de efluente) y de 27 a 86% (100 mL de
influente). Los límites de detección estuvieron en el rango de 1 a 4 ng L-1 para
río, de 3 a 10 ng L-1 para efluente y de 5 a 20 ng L-1 para influente.
Introducción-Filtros solares
44
Oliveira y col. [Oliveira, 2010] proponen un método automático para la
determinación de 3 filtros solares: BP-3, EHMC y HMS en muestras de agua de
piscina y de mar. La nueva metodología consiste en una SPE-on line con el
sistema cromatográfico, en este caso LC-UV/VIS. El sistema microanalítico
propuesto, usando multijeringas como unidad de propulsión, combina todas las
etapas de la SPE incluyendo la renovación del adsorbente para prevenir la
contaminación cruzada entre muestras, y el ajuste, después de la extracción, de
la composición del eluato para prevenir el ensanchamiento de la banda de
inyección en cabeza de columna. Con objeto de acelerar la separación en LC, se
usó una columna monolítica de C18 y se llevó a cabo una elución en modo
isocrático. Las recuperaciones logradas estuvieron en el rango de 51 a 140% y
los límites de detección variaron entren 0,81 y 3,2 µg L-1. El interés del método
anteriormente descrito se basa en su completo grado de automatización; sin
embargo, los límites de detección alcanzados son demasiado elevados para su
aplicación rutinaria a muestras de aguas superficiales.
A continuación, se muestra una tabla resumen con las aplicaciones más
significativas de SPE a la determinación de filtros UV en muestras de agua,
Tabla 11.
Introducción-Filtros solares
45
Tabla 11: Resumen de aplicaciones de SPE como técnica de concentración para la
determinación de filtros solares en muestras de agua.
Matriz Adsorbente Eluyente (volumen, mL) R.
(%) Referencia
Mar y piscina C18 Acetato de etilo:diclorometano,
1:1 (2 x 3 mL)
98-
101 [Sakkas, 2003]
Río y baño Poliestireno Metanol (5 mL),
diclorometano (10 mL)
57-
90 [Poiger, 2004]
Piscina,
residuos de
duchas
Discos C18
(500 mg)
Acetato de etilo:diclorometano,
1:1 (2 x 5 mL)
95-
99 [Giokas, 2004]
Residual Poliestireno-
divinilbenzeno Metanol:diclorometano
78-
129 [Balmer, 2005]
Río, lago,
piscina, mar
Strata X
(60 mg)
Acetato de etilo:diclorometano,
1:1 (3 x 0,5 mL)
50-
98
[Cuderman,
2007]
Residual C18
(200 mg)
Acetato de etilo:diclorometano,
1:1 (5 mL)
67-
118 [Li, 2007]
Residual HLB
(200 mg)
Metanol:MTBE, 10:90 (5 mL);
Diclorometano (5 mL)
67-
87
[Trenholm,
2008]
Superficial y
residual
Oasis MCX
(60 mg)
Metanol (2 mL), Metanol-
5%hidróxido amónico (2 mL)
17-
117
[Kasprzyk-
Hordern, 2008]
Río, mar
y residual
Oasis HLB
(60 mg) Metanol (3 x 10 mL)
63-
108 [Rodil, 2008-A]
Grifo, ría
y residual
Oasis HLB
(200 mg) Metanol (3 x 10 mL)
56-
132 [Rodil, 2009-B]
Residual
Río
Oasis HLB
(500 mg)
Bond Elut Plexa
(200 mg)
Metanol (5 mL),
diclorometano (5 mL)
46-
97
20-
86
[Pedrouzo,
2009]
Aunque SPE permite llevar a cabo en una única etapa la extracción y
concentración de los analitos, la cantidad de muestra a procesar es más elevada
que en el caso de las técnicas de microextracción, además el coste de los
adsorbentes empleados es considerable. Las técnicas más modernas de
preparación de muestra tienden hacia la simplificación y miniaturización del
Introducción-Filtros solares
46
proceso analítico, reduciendo tiempo, esfuerzo, cantidad de muestra y consumo
de disolventes orgánicos [Pawliszyn, 2002] [Kloskowski, 2007]. Como resultado de
la miniaturización de la LLE (extracción líquido-líquido) y de la SPE, surgieron
diferentes metodologías de microextracción. La mayoría de ellas han sido
aplicadas a la concentración de filtros solares en muestras de agua. A
continuación se revisan estas aplicaciones, prestando especial interés a las
modalidades empleadas en este estudio.
3.1.1.2. Técnicas basadas en la microextracción en fase sólida
Como resultado de la minituarización de la SPE, surgieron varias
técnicas de microextracción en fase sólida: la microextracción en fase sólida
(SPME) [Belardi, 1989] y otras más recientes como: la microextracción con barras
agitadoras (SBSE) [Baltussen, 1990] y la microextracción mediante sorbentes
empaquetados (MEPS) [Mohamed, 2010]. Además, en esta Tesis Doctoral se
propone el uso de siliconas de grado técnico, en formato no comercial, como
sustituyente de las barras agitadoras recubiertas con polidimetilsiloxano
(PDMS) empleadas en SBSE, para la extracción de filtros solares en muestras
acuosas. A continuación, se describe el fundamento de las técnicas empleadas
en esta memoria (SPME y siliconas), y se revisan sus aplicaciones, así como las
de SBSE y MEPS, a la determinación de filtros solares.
3.1.1.2.1. Microextracción en fase sólida (SPME)
La microextracción en fase sólida (SPME) es una técnica de preparación de
la muestra desarrollada en 1989 por Belardi y Pawliszyn [Belardi, 1989]
[Pawliszyn, 2002] y que se basa en la utilización de una fibra de sílice fundida,
recubierta con una fase estacionaria adsorbente/absorbente de naturaleza
polimérica. El dispositivo empleado puede verse en la figura siguiente (Fig. 6).
Los analitos presentes en la muestra, por lo general, no se extraen
Introducción-Filtros solares
47
cuantitativamente sobre la fibra, sino que simplemente se establece un
equilibrio entre ambas fases. La SPME no necesita de la utilización de
disolventes orgánicos, requiere una manipulación mínima de la muestra y es
aplicable tanto a compuestos volátiles como semivolátiles [Cámara, 2002] [Cela,
2002].
Cuerpo de la jeringa
Émbolo
Guía del émbolo
Muelle
Aguja de acero
Fibra de sílice
Séptum
Férrula
Pieza de acero fijación fibra
Cuerpo de la jeringa
Émbolo
Guía del émbolo
Muelle
Aguja de acero
Fibra de sílice
Séptum
Férrula
Pieza de acero fijación fibra
Muelle
Aguja de acero
Fibra de sílice
Séptum
Férrula
Pieza de acero fijación fibra
Muelle
Aguja de acero
Fibra de sílice
Séptum
Férrula
Pieza de acero fijación fibra
Figura 6. Esquema de un dispositivo comercial de SPME
La concentración de las muestras mediante SPME consta de dos etapas:
1 ª Extracción o muestreo. La fibra se pone en contacto con la muestra, o en
espacio de cabeza, situada en un vial cerrado con un séptum y una cápsula. Se
deja un tiempo determinado para que los analitos se absorban/adsorban en la
fibra, produciéndose el reparto de los mismos entre la muestra y el
recubrimiento de la fibra. A continuación, la fibra se retrae en la aguja
protectora.
2 ª Desorción. Inmediatamente después, la fibra se introduce en un inyector
de un instrumento analítico (GC o HPLC), donde los analitos son desorbidos
térmicamente o por disolución en la fase móvil, según la técnica de
determinación empleada. Esta etapa se lleva a cabo en 1-2 minutos.
Los modos de extracción empleados en SPME son los siguientes:
Extracción directa. En este modo, la fibra se introduce directamente
en la muestra líquida y los analitos son transportados directamente
Introducción-Filtros solares
48
desde la matriz de la muestra a la fase extractante. Está
recomendada para compuestos poco volátiles, en muestras con un
nivel bajo o moderado de interferencias.
Extracción en espacio de cabeza (HS). La fibra se expone al espacio de
cabeza de la muestra, de manera que los analitos son transportados
primero al espacio de cabeza y luego son concentrados en la fibra.
Así se protege la fase extractante de compuestos de alto peso
molecular y poco volátiles que no son de interés.
Extracción indirecta a través de una membrana protectora. Con esta
membrana se evita el deterioro de la fibra cuando se extraen
muestras complejas. Su utilización es adecuada cuando se requiere
determinar analitos con volatilidades demasiado bajas para su
extracción en el modo de espacio de cabeza. Se trata de la
modalidad menos utilizada dada su lenta cinética de extracción y la
posible competencia entre la membrana y el recubrimiento de la
fibra por los analitos.
A continuación, se presenta un esquema de SPME para los modos de
extracción directo y espacio de cabeza (Fig. 7).
Figura 7: Esquema del proceso de SPME-GC: (a) Inmersión directa; (b) Espacio de
cabeza; (c) Desorción térmica en GC.
La SPME es un proceso de equilibrio entre múltiples fases. Normalmente,
para simplificar la descripción del mismo, sólo se consideran tres fases: el
recubrimiento de la fibra, la fase gaseosa o espacio de cabeza y una matriz
homogénea como el agua pura. Mediante un sencillo balance de masas se
Introducción-Filtros solares
49
puede relacionar la cantidad total de analito en el vial de SPME con la masa
extraída por la fibra y la remanente en el espacio de cabeza y en la matriz, como
se indica en la siguiente ecuación:
ffsshhso VCVCVCVC (1)ffsshhso VCVCVCVC ffsshhso VCVCVCVC (1)
Co: Concentración inicial del analito en la matriz.
Cf∞: Concentración del analito en la fase estacionaria al alcanzar el
equilibrio.
Ch∞: Concentración del analito en el espacio de cabeza al alcanzar el
equilibrio.
Cs∞: Concentración del analito en la matriz al alcanzar el equilibrio.
Vf: Volumen del recubrimiento en la fibra de SPME.
Vh: Volumen del espacio de cabeza.
Vs: Volumen de la matriz.
Los coeficientes de reparto entre las tres fases son:
h
ffh C
CK
s
hhs C
CK
s
ffs C
CK(2) (3) (4)
Así, la cantidad de analito en el recubrimiento de la fibra (n = CfVf) se
puede expresar como:
shhsfhsfh
sofhsfh
VVKVKK
VCVKKn
(5)
Cuando se alcanza el equilibrio:
hsfhfs KKK (6)hsfhfs KKK (6)
Y, substituyendo en la ecuación anterior:
shhsffs
soffs
VVKVK
VCVKn
(7)
Cuando no existe espacio de cabeza, es decir, la muestra ocupa por
completo el vial, el término KhsVs es despreciable:
sffs
soffs
VVK
VCVKn
(8)
Introducción-Filtros solares
50
Si además estipulamos que el volumen de la muesta acuosa (Vs) es mucho
mayor que el volumen de recubrimento de la fibra (Vf), se puede cumplir que
KfsVf << Vs, por eso:
n = KfsVfCo (9)
De acuerdo con la expresión anterior, una vez alcanzado el equilibrio, la
cantidad de analito extraído es independiente del volumen de la muestra.
La constante Kfs controla la eficacia y la selectividad del proceso de
extracción. Su valor es función de las características del recubrimiento, de la
muestra y de las propiedades físico-químicas de cada analito en cuestión.
Actualmente, se comercializan varios tipos de fases estacionarias con diferentes
espesores y polaridades, que muestran mayor o menor afinidad por diferentes
analitos. En la Tabla 12 se indican algunas de las fases más empleadas. En breve,
está previsto la comercialización de fibras recubiertas con C18, más adecuadas
para su desorción con disolventes orgánicos que las hasta ahora disponibles.
Tabla 12: Recubrimentos más utilizados en SPME.
Fase estacionaria Espesor (µm) Polaridad Tª Máxima (ºC)
PDMS
100
Apolar
280
30 280
7 340
PDMS/DVB 65 Semipolar 270
PA 85 Polar 320
CAR/PDMS 75 Semipolar 320
DVB/CAR/PDMS 50/30 Semipolar 270
PEG 60 Polar 250
Carbopack 15 Semipolar 340
Los trabajos que proponen el uso de SPME para la concentración de filtros
solares son los siguientes:
Felix y col. [Felix, 1998] aplicaron la SPME, en combinación GC-MS, a la
determinación de BP-3 y sus metabolitos (BP-1, BP-8) en orina. Compararon tres
fibras diferentes: una de 30 µm de polidimetilsiloxano (PDMS), de 85 µm de
Introducción-Filtros solares
51
poliacrilato (PA) y de 65 µm de carbowax-divinilbenceno (CW-DVB). La fibra
de CW-DVB (actualmente descatalogada) fue la que proporcionó una mayor
eficacia de extracción. Los parámetros evaluados fueron: el rango lineal, que se
encontró entre 10 y 500 ng mL-1, y los límites de cuantificación que fueron de
1667 ng L-1, 10000 ng L-1 y 3704 ng L-1 para BP-3, BP-1 y BP-8 respectivamente.
Lambropoulou y col. [Lambropoulou, 2002] desarrollaron un método para la
determinación de BP-3 y EHPABA en agua de baño y de piscina. La
determinación se llevó a cabo usando cromatografía de gases con detector de
ionización de llama y también, espectrometría de masas. Las fibras más eficaces
fueron PDMS de 100 µm y PA de 85 µm. La SPME fue realizada en modo
directo y en espacio de cabeza. Las recuperaciones relativas (respecto a agua
ultrapura) obtenidas fueron de 82% a 98% con límites de cuantificación de 1000
a 4100 ng L-1 para GC-FID y 730 a 4430 ng L-1 para GC-MS.
Evidentemente, los valores anteriores son claramente insuficientes para
la determinación de filtros UV en muestras de agua superficial. En esta tesis, se
ha profundizado en la aplicabilidad de SPME a la determinación de estos
compuestos en muestras de agua, centrando el estudio en compuestos fenólicos
del grupo de los salicilatos y las benzofenonas, combinando extracción y
derivatización con objeto de mejorar las características analíticas del método
resultante [Negreira, 2009-A].
Posteriormente, Liu y col. [Liu, 2010] optimizaron un método de SPME
para la determinación de cuatro filtros UV (BP-3, EHS, 4-MBC y OCR), junto
con fragancias, en agua de río. La fibra empleada fue la de PDMS y ésta fue
sumergida directamente en 3 mL de muestra ajustada a pH 7 y conteniendo un
10% de NaCl. El muestreo se realizó a 24ºC durante 90 min., posteriormente la
fibra se desorbió en el inyector de un sistema GC-MS durante 7 min. Los límites
de detección y de cuantificación estuvieron en el rango de 0,2 a 2,0 ng L-1 y de
0,7 a 6,7 ng L-1, respectivamente.
Introducción-Filtros solares
52
3.1.1.2.2. Microextracción con barras agitadoras (SBSE)
La extracción en fase sólida con barras agitadoras (stir-bar soptive
extraction, SBSE) tiene el mismo fundamento que la SPME. La principal
diferencia entre ambas técnicas es el diseño del sistema de extracción y la
cantidad de sorbente utilizado. La barra agitadora (Twister) está recubierta con
PDMS y, dado que su superficie y volumen son superiores a los de la fibra de
SPME, la cantidad de recubrimiento es superior; en consecuencia, la eficacia de
extracción es también mayor que la de la SPME. La muestra líquida se agita con
la barra durante un cierto tiempo, tras el cual ésta se retira y se desorbe. La
desorción puede realizarse con un disolvente orgánico o, para alcanzar límites
de cuantificación más bajos, se puede llevar a cabo térmicamente en
combinación con GC-MS. En este caso, los tiempos de desorción típicos son de
10 minutos, por lo que los analitos deben reenfocarse en cabeza de columna,
antes de su separación. Esto se lleva a cabo, habitualmente, con un sistema de
enfriamiento criogénico. Por otro lado, a diferencia de SPME, con SBSE
normalmente se alcanzan recuperaciones cuantitativas para compuestos de
polaridades medias y bajas.
Fundamento teórico
La base teórica de SBSE es exactamente la misma que la de SPME. La
principal diferencia es que se utiliza un volumen de polímero mucho mayor
que en caso de las fibras convencionales de SPME, entre 20 y 150 µL de PDMS
dependiendo de las aplicaciones, frente a los 0,5 µL de una fibra de PDMS
comercial. Así, se puede aumentar el rendimiento del proceso.
La eficacia de extracción, E, puede calcularse como:
100100000
VC
VC
n
nE fff (10)
Introducción-Filtros solares
53
Siendo nf la cantidad de analito retenida sobre el polímero, en
condiciones de equilibrio, y n0 la cantidad total de compuesto presente en la
muestra.
100
ssff
ff
VCVC
VCE (11)
fsfsf
s
fs
sf
f
KKV
V
K
VV
VE
1
1100
1
1100 (12)
La eficacia de extracción está relacionada con el parámetro β, cociente
entre el volumen de muestra y de PDMS, y la constante de partición Kfs. En
SBSE, es habitual usar las Kow como estimado de los valores de Kfs.
Aplicaciones
Los primeros trabajos que aplican SBSE para la extracción y
concentración de filtros solares están enfocados a la determinación de
benzofenonas, entre las que figuran la BP, BP-1 y BP-3 en muestras de orina
[Kawaguchi, 2008-B] y de agua [Kawaguchi, 2008-A; Kawaguchi, 2006]. El método
propuesto consiste en introducir una barra cubierta con PDMS en 10 mL de
muestra y agitar durante 120 min a temperatura ambiente (25ºC). A
continuación, el Twister se pasa al sistema de desorción térmica conectado on-
line con GC-MS [Kawaguchi, 2008-A; Kawaguchi, 2006]. En uno de esos trabajos,
realizan la derivatización in-situ de las benzofenonas con anhídrido acético,
antes de la desorción térmica (TD) en GC-MS [Kawaguchi, 2008-A]. Los límites
de detección alcanzados se recogen en la Tabla 13.
Rodil y col. [Rodil, 2008-B] aplicaron SBSE en combinación con TD y GC-MS
a la determinación de filtros solares (EHS, HMS, IAMC, 4-MBC, BP-3, EHMC,
EHPABA, OCR) en muestras acuosas. La barra agitadora cubierta con
polidimetilsiloxano (PDMS) fue introducida en 20 mL de agua a pH 2 (10% de
Introducción-Filtros solares
54
metanol) y agitada a 1000 rpm, durante 180 min a temperatura ambiente. La
desorción se llevó a cabo a 250 ºC durante 15 min. Los límites de cuantificación
se situaron en el rango de 0,6 a 26 ng L-1, los coeficientes de determinación en el
rango de 0,9941 a 0,9999, la eficacia de extracción fue superior al 63% y las
desviaciones estándar relativas inferiores al 16%.
Pedrouzo y col. [Pedrouzo, 2010] usaron también SBSE y cromatografía
líquida (UPLC-(ESI)MS-MS) para la determinación de 4 filtros solares (BP-8, BP-
3, OCR y EHPABA) en muestras de agua. Tomando un volumen de muestra de
50 mL, y considerando 1 mL de acetonitrilo para la desorción de los analitos, se
alcanzaron eficacias de extracción entre 31 y 97% y LODs comprendidos entre
2,5 y 10 ng L-1.
En esta Tesis doctoral, se propone, como substituyente de las barras
agitadoras, el uso de siliconas en un formato no comercial, para la extracción y
concentración de filtros solares en muestras de agua. Es una modalidad de
microextracción en fase sólida con fundamento idéntico a la SPME
convencional, utilizando un absorbente líquido, en la que, en lugar de utilizar el
dispositivo comercial de SPME (Supelco) o de SBSE (Gerstel), se emplean
siliconas de grado técnico, en distintos formatos: láminas, tubo, cuerda… El
polímero, en estos formatos, puede ser cortado por el usuario en unas
dimensiones adecuadas para optimizar el rendimiento de la extracción para
cada analito, volumen de muestra y volumen de disolvente empleado en la
etapa de desorción.
Debido a su bajo coste (normalmente inferior a 0,1 €), la fase extractante
(silicona) puede ser utilizada como material desechable, con lo que se evitan los
problemas de contaminación cruzada, o el “efecto memoria” de las fibras de
SPME y los Twisters usados en SBSE.
Introducción-Filtros solares
55
Los primeros trabajos en los que se utilizaron las siliconas, en formato de
cuerda, fueron publicados por el grupo de investigación de Popp y col.
[Montero, 2004] [Popp, 2004] para la determinación de PCBs y PAHS en muestras
de agua. Posteriormente, se han desarrollado aplicaciones a otros compuestos
[Paschke, 2006] [Bicchi, 2007] [Shellin, 2007] [van Pinxteren, 2010]. Sin embargo, no
existen trabajos en los que usen siliconas de grado técnico, en formatos no
comerciales, para la determinación de filtros solares, salvo el presentado en esta
Tesis.
3.1.1.2.3. Microextracción mediante sorbentes empaquetados (MEPS)
Moeder y col. [Moeder, 2010] describieron el uso de MEPS para la extracción
automatizada de 4 filtros solares (BP-3, 4-MBC, EHMC y OCR) en muestras de
agua. En esta técnica, el material adsorbente (C18) se encuentra empaquetado
entre la aguja y el cuerpo de una jeringa de 100 µL, instalada en el autosampler
del equipo de GC-MS. La muestra (0,8 mL) se aspiró a través del material
adsorbente (1 mg de C18) y a continuación se pasó a desecho. Una vez secado el
adsorbente mediante la aspiración de aire, los analitos se desorben con dos
fracciones de 25 µL de acetato de etilo que se tranfirieron directamente al
inyector de grandes volúmenes del sistema GC-MS. Los LODs alcanzados
estuvieron en el rango de 34 a 87 ng L-1 y las recuperaciones obtenidas variaron
de 61 a 114%.
En la Tabla 13 se resumen las condiciones experimentales, las eficacias de
extracción, y los LOQ de las técnicas de microextracción en fase sólida
empleadas en la bibliografía para la determinación de filtros solares.
Introducción-Filtros solares
56
Tabla 13: Resumen de aplicaciones de las técnicas de microextracción en fase sólida a la
determinación de filtros UV.
Matriz Condiciones EF.
(%) Det.
LOQ
(ng L-1) Referencia
SPME
Orina CW-DVB, Vm: 4 mL - GC-MS 1667-
10000 [Felix, 1998]
Agua de piscina
y baño
PDMS, PA,
directa (0% NaCl)
y HS (30% NaCl),
Vm: 5 mL, 45min
- GC-MS 730-
4430
[Lambropoulou,
2002]
Agua de río
PDMS, inmersión,
Vm: 3 mL (pH 7, 10%
NaCl), 24ºC, 90 min
- GC-MS 0,7-6,7 [Liu, 2010]
SBSE
Agua de río SBSE (24 µL),
10 mL, 120 min
98-
115
106-
128
TD-GC-
MS
2-5
2-10
[Kawaguchi,
2006]
[Kawaguchi,
2008-A]
Agua de río,
lago y residual
SBSE (24 µL), 20 mL (pH
2, 10%metanol), 180 min
63-
122
TD-GC-
MS 0,6-26 [Rodil, 2008-B]
Agua de río,
efluente e
influente
SBSE (24 µL),
50 mL (pH 5), 180 min
Desorción: 1 mL ACN,
30ºC, 15 min
31-
97
UPLC-
MS/MS 8-33 [Pedrouzo, 2010]
MEPS
Agua de lago
y efluente
Vm: 0,8 mL
Desorción: 2 x 25 µL,
acetato de etilo
61-
114
PTV-GC-
MS 113-290 [Moeder, 2010]
Introducción-Filtros solares
57
3.1.1.3. Técnicas basadas en la microextracción en fase líquida
Tanto las modalidades de microextracción líquido-líquido directa,
(microextracción con gota suspendida, SDME y microextracción líquido-líquido
dispersiva, DLLME), como aquellas en las que existe una membrana entre
ambas fases (microextracción líquido-líquido con membranas no porosas,
MALLE y microextracción en fase líquida con fibra hueca, HF-LPME) han sido
evaluadas para la concentración de filtros solares en matrices acuosas. A
continuación, se resumen las aplicaciones de éstas técnicas a la determinación
de filtros solares, prestando especial atención a la DLLME, que es la modalidad
utilizada en esta Tesis Doctoral [Negreira, 2010-A].
3.1.1.3.1. Microextracción con gota suspendida (SDME)
Vidal y col. [Vidal, 2007] presentaron la primera aplicación de SDME a la
determinación de BP-3 en muestras de orina. En SDME, la fase extractante es
una gota de unos pocos microlitros de un disolvente inmiscible con la muestra,
suspendida en el extremo de una microjeringa, y normalmente expuesta
directamente a la muestra. Como alternativa, para analitos volátiles, la gota
puede ser suspendida en espacio de cabeza del vial que contiene la muestra. En
la aplicación desarrollada por Vidal y col. [Vidal, 2007] emplearon un líquido
iónico (IL) en lugar de un disolvente orgánico. Después del proceso de
microextracción, la fase extractante fue inyectada en un equipo de HPLC. Las
condiciones experimentales óptimas encontradas fueron: NaCl 13% (p/v), 25
min de tiempo de extracción, 900 rpm como velocidad de agitación y pH 2. El
volumen de gota fue de 5 µl y el de muestra de 10 mL. El método se usó para
determinar las concentraciones de BP-3 presentes en orina humana, después de
la aplicación de protectores solares que contienen este compuesto. El límite de
detección fue de 1300 ng L-1 y la repetibilidad del método, expresada como RSD
fue del 6% (n=8). Posteriormente, la técnica de SDME fue usada también para la
Introducción-Filtros solares
58
determinación de benzofenonas en muestras de aguas por Okanouchi y col.
[Okanouchi, 2008] y Vidal y col. [Vidal, 2010].
3.1.1.3.2. Microextracción líquido-líquido con membranas
La microextracción líquido-líquido con membranas no porosas, conocida
por las siglas MALLE correspondientes a “non-porous membrane-assisted liquid-
liquid extraction”, fue aplicada por Rodil y col. [Rodil, 2009-C] para la
determinación de filtros solares (BP-3, IAMC, 4-MBC, OCR, EHPABA, EHMC,
EHS y HMS) en muestras de agua mediante LC-(APPI)MS/MS. Utilizaron
membranas de 2 cm de polietileno de baja densidad (LDPE), rellenas con 100 µL
de propanol. La membrana fue sumergida en 15 mL de muestra, conteniendo
un 10% de metanol, con agitación durante 120 min a 40 ºC. A continuación, la
fase orgánica fue retirada de la membrana y transferida a un inserto. El método
optimizado proporcionó recuperaciones de 60% (BP-3) a 104% (EHS) y límites
de cuantificación entre 3 ng L-1 (EHPABA) y 53 ng L-1 (EHMC).
3.1.1.3.3. Microextracción en fase líquida con fibra hueca (HF-LPME)
Kawaguchi y col. [Kawaguchi, 2010] describieron una aplicación de la
denominada microextracción en fase líquida con fibra hueca (HF-LPME), para
la determinación de benzofenonas, entre ellas BP y BP-3, en orina humana. En
su trabajo, emplearon una fibra porosa de polipropileno que impregnaron con
tolueno y conectaron a una jeringa. La punta de la aguja junto con la fibra hueca
se sumerge en la muestra y después de la extracción, agitando durante 15 min a
temperatura ambiente, se inyectan 2 µL de extracto en el sistema GC-MS. Se
obtuvieron recuperaciones de 89 y 99% y LODs de 10 y 5 ng L-1 para BP y BP-3,
respectivamente.
Introducción-Filtros solares
59
3.1.1.3.4. Microextracción líquido-líquido dispersiva (DLLME)
La microextracción líquido-líquido dispersiva (DLLME) fue desarrollada
en 2006 por Assadi y col. [Rezaee, 2006] para la extracción de compuestos
orgánicos e inorgánicos de muestras acuosas. DLLME presenta principios
comunes a LLE pero evita el elevado consumo de disolventes orgánicos que
ésta conlleva y permite la extracción y concentración simultánea de los analitos.
DLLME se basa en el uso de un sistema ternario de disolventes, constituido por
la fase acuosa (la muestra de la que los analitos pretenden ser extraídos) y una
mezcla de dos disolventes orgánicos: uno miscible con agua, que funciona como
agente dispersante, y otro altamente inmiscible con agua y miscible con el
agente dispersante. Este último se denomina extractante y debe poseer una
densidad diferente a la de la mezcla muestra: dispersante. En la modalidad más
habitual de DLLME se usan disolventes halogenados de elevada densidad
como extractantes.
La mezcla dispersante-extractante se pone en contacto con la fase acuosa
en el interior de un tubo cónico formándose una emulsión. De esta forma se
incrementa al máximo la superficie de contacto entre las fases, favoreciendo así
la rápida transferencia del analito entre la muestra y el extractante.
Posteriormente, se procede a la centrifugación y se obtiene la separación de
fases, por un lado, el agente extractante, que se deposita en el fondo del tubo
cónico debido a su mayor densidad y que contendrá disueltos los analitos y,
por otro lado, la muestra con el agente dispersante (Fig. 8).
Introducción-Filtros solares
60
1ª 2ª
Figura 8: Etapas en DLLME: 1ª Formación de una emulsión de la mezcla dispersante-
extractante-muestra; 2ª Separación de la fase extractante en el fondo del vial cónico.
Fundamento termodinámico
El coeficiente de distribución, K, se define como el cociente entre la
concentración de analito en la fase orgánica y en la fase acuosa. La DLLME sólo
es aplicable para analitos en forma neutra, que tengan un elevado carácter
hidrofóbico, y es poco viable para la extracción de especies hidrofílicas (K<500).
No obstante, para compuestos ácidos o básicos, pueden realizarse
modificaciones del pH del medio, con el fin de desplazar el equilibrio ácido-
base hacia la forma neutra y con ello, incrementar su afinidad por la fase
orgánica.
Los parámetros que, normalmente, se utilizan a la hora de caracterizar la
eficacia de la DLLME son el factor de enriquecimiento (EF) y la recuperación
(R). El factor de enriquecimiento se define como el cociente entre la
concentración del analito en la fase sedimentada (Csed) y la concentración inicial
en la muestra (C0) (Ecuación 13). La recuperación se calcula multiplicando el
factor de enriquecimiento por el cociente entre el volumen de la fase
sedimentada (Vsed) y el de la muestra (V0) (Ecuación 14).
0C
CEF sed (13)
1000
V
VEFR sed (14)
Introducción-Filtros solares
61
Parámetros que afectan a la eficacia de la extracción
Selección del disolvente extractante
En DLLME, el agente extractante debe satisfacer las siguientes
condiciones indispensables:
- tener una elevada densidad y baja solubilidad en agua, lo que
posibilitará el depósito de la fase sedimentada en el fondo del vial
cónico mediante centrifugación;
- tener capacidad para extraer los compuestos de interés;
- un buen comportamiento cromatográfico, ya que va a ser
inyectado directamente en el sistema de cromatografía de gases.
Los disolventes más utilizados como extractantes son los hidrocarburos
halogenados, tales como cloroformo CHCl3, tetracloruro de carbono CCl4,
tricloroetano CH3Cl3, clorobenzeno ClBz, etc, aunque también hay aplicaciones
en las que se ha empleado el disulfuro de carbono CS2 [Rahnana, 2007] y los
líquidos iónicos [Zhou, 2008]. Estos últimos suelen utilizarse cuando el método
de determinación es la cromatografía líquida. La cantidad de extractante que se
utiliza suele ser inferior a 200 µL, con lo que la fase sedimentada que se obtiene
es de tamaño muy pequeño, consiguiéndose así unos elevados factores de
enriquecimiento. En la Tabla 14, se muestran algunas propiedades de los
disolventes más utilizados como extractantes en DLLME.
Tabla 14: Propiedades físico-químicas de los extractantes más utilizados en DLLME.
Propiedad Disolvente
CCl4 ClBz CH3CCl3 CHCl3 CS2
Densidad (g mL-1) 1,59 1,10 1,34 1,48 1,26
Solubilidad en agua (10-3M) 2,0 0,79 7,3 16 5
Log Kow 2,89 2,81 2,10 1,76 1,94
Selección del disolvente dispersante
El disolvente que se utiliza como dispersante debe ser miscible con el
agente extractante y con la muestra acuosa, lo que permite la formación de una
Introducción-Filtros solares
62
emulsión cuando la mezcla dispersante-extractante es añadida sobre la muestra,
este fenómeno provoca un gran aumento de la superficie de contacto entre la
muestra y el extractante, favoreciendo el paso de los analitos a la fase orgánica e
incrementando, por lo tanto, la eficacia y la cinética de extracción. Los
disolventes más utilizados para este fin son etanol, metanol, acetona,
acetonitrilo y tetrahidrofurano.
Volúmenes de extractante y dispersante
La cantidad de agente extractante utilizada tiene una gran influencia sobre
el factor de enriquecimiento. Un aumento en el volumen añadido provoca un
aumento en el tamaño de la fase sedimentada, lo que conlleva una disminución
de la concentración del analito en esta fase (Csed). Según la ecuación 13, esta
disminución producirá una disminución en el factor de enriquecimiento (EF),
ya que la concentración inicial C0 permanece constante. La cantidad óptima será
aquella que genere un elevado factor de enriquecimiento, y que sea lo
suficientemente elevada como para que la fase sedimentada pueda ser
fácilmente manipulable. Normalmente, se emplean volúmenes entre 20 y 100
µL.
Por otra parte, el volumen de agente dispersante afecta principalmente a
la formación de la emulsión. Cuanto mayor sea el grado de dispersión, mejor
será el contacto entre fases, y mayor será la eficacia de la extracción. Este
volumen puede afectar también, aunque en menor medida, al tamaño de fase
sedimentada, por lo que ambas contribuciones deben tenerse en cuenta. Suelen
seleccionarse volúmenes comprendidos entre 0,5 y 1,5 mL para muestras de 10
mL.
Efecto de la centrifugación
La centrifugación es necesaria para que la fase orgánica se deposite en el
fondo del tubo de fondo cónico. Sin embargo, no se han encontrado evidencias
de que sean necesarios tiempos elevados de centrifugación, por lo que la
mayoría de los autores utilizan valores inferiores a 5 minutos.
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63
Tiempo de extracción
En DLLME, se entiende por tiempo de extracción el que transcurre entre la
adición de la mezcla extractante-dispersante sobre la muestra y la
centrifugación. Este parámetro apenas tiene influencia sobre la eficacia de la
extracción ya que la transferencia de los analitos entre la muestra y el
extractante ocurre de manera inmediata tras la dispersión y el equilibrio entre
fases se alcanza rápidamente [Montes, 2009]. Esta característica representa una
de las ventajas más relevantes de la DLLME frente a otras técnicas de
microextracción.
Fuerza iónica y pH
La adición de sal sobre la muestra suele disminuir la solubilidad de los
analitos en esta fase, favoreciendo su paso hacia la fase orgánica (extractante) y
aumentando, por tanto, la eficacia de la extracción. Por otro lado, la adición de
sal provoca un aumento del volumen de fase sedimentada, con lo que
disminuye la concentración de los analitos en esta fase y por lo tanto, el factor
de enriquecimiento se verá afectado negativamente. Por todo ello, es necesario
mostrar especial atención a ambos efectos y elegir la fuerza iónica adecuada.
La variación del pH es especialmente importante cuando se trata de
analitos con características ácidas o básicas, ya que ajustando el pH puede
desplazarse el equilibrio de los mismos hacia su forma neutra consiguiendo la
extracción de especies que a priori, no podrían ser extraídas.
Aplicaciones
A pesar de que la DLLME es una técnica bastante reciente (año 2006), se
han desarrrollado numerosas aplicaciones en estos últimos años [Rezaee, 2010].
Siguiendo el trabajo inicial de Assadi y col. [Rezaee, 2006], para la determinación
de PAHS en agua, se usó la DLLME para la determinación de otras familias de
compuestos: pesticidas organofosforados [Berijani, 2006], trihalometanos
[Rahnama, 2007], clorobencenos [Kozani, 2007], bifenilos policlorados [Rezaei,
Introducción-Filtros solares
64
2008], ésteres del ácido ftálico [Farahani, 2007] y retardantes de llama
organofosforados [García-López, 2007].
La DLLME también permite la incorporación de una etapa adicional de
derivatización para aquellos compuestos que no son adecuados para su
determinación directa mediante cromatografía de gases. Así, Montes y col.
[Montes, 2009] realizaron de manera simultánea la derivatización, extracción y
concentración de triclosán en muestras de agua incluyendo el agente
derivatizante (MTBSTFA) en la mezcla dispersante (acetona)-extractante
(tricloroetano).
Se debe resaltar también la eficacia de la DLLME como etapa de
purificación y concentración para extractos obtenidos mediante otras técnicas
como la SPE. Normalmente, el extracto orgánico se emplea como agente
dispersante que, combinado con un agente extractante y añadido sobre la
disolución acuosa, da lugar a fases sedimentadas libres de interferencias. La
combinación SPE-DLLME fue aplicada por Assadi y col. a la extracción y
derivatización de clorofenoles en muestras acuosas [Fattahi, 2007] y
posteriormente por otros autores para la determinación de éteres difenil
polibromados [Liu, 2009] y herbicidas [Zhao, 2009] en agua.
Por último, decir que la técnica de DLLME también ha experimentado
adaptaciones. Regueiro y col. [Regueiro, 2008] proponen una técnica de
microextracción, conocida como microextracción–emulsificación asistida por
ultrasonidos (USAEME), para la determinación de fragancias y ftalatos en
muestras de agua. Esta técnica consiste en la formación de la emulsión con un
agente extractante (cloroformo) en una matriz acuosa por acción de
ultrasonidos, prescindiendo del agente dispersante. Durante este proceso tiene
lugar la transferencia de los analitos desde la muestra acuosa a las microgotas
de disolvente orgánico que se encuentran dispersas en la misma. Ambas fases
son posteriormente separadas mediante centrifugación y el extracto orgánico
resultante es recogido para su análisis. Otras tendencias claras en DLLME son la
Introducción-Filtros solares
65
utilización de disolventes no halogenados como extractantes y el uso de micelas
como dispersantes [Herrera-Herrera, 2010].
Tan sólo se descirben dos aplicaciones de la DLLME a la determinación de
filtros solares. Ambas se centran en el análisis de muestras acuosas y emplean
GC-MS como técnica de determinación. Tarazona y col. [Tarazona, 2010] la
aplican a la determinación de cuatro benzofenonas (BP-3, BP-8, BP-1 y 2,3,4-
trihidroxibenzofenona) en agua de mar. Utilizan 1 mL de acetona y 60 µL de
cloroformo como agentes dispersante y extractante, respectivamente. El
volumen de muestra fue de 5 mL, ajustada a pH 4 y conteniendo NaCl (10%).
Los extractos fueron evaporados y reconstituidos con N,O-bis-
(trimetilsilil)trifluoroacetamida (BSTFA) para su derivatización antes de la
inyección en el sistema GC-MS.
La otra aplicación es la desarrollada en esta Tesis para la determinación
de EHS, HMS, BzS, IAMC, EHMC, 4-MBC, BP-3, EHPABA y OCR en muestras
de agua de río, de piscina y residual [Negreira, 2010-A].
En la Tabla 15 se resumen las condiciones experimentales, las
recuperaciones y los LOQs de las técnicas de microextracción en fase líquida
empleadas en la bibliografía para la determinación de filtros solares.
Introducción-Filtros solares
66
Tabla 15: Resumen de las aplicaciones de las técnicas de microextracción en fase líquida
a la determinación de filtros UV.
Matriz Condiciones EF.
(%) Det.
LOQ
(ng L-1) Referencia
SDME
Orina Vm: 10 mL (pH 2, 13% NaCl),
5 µL [C6MIM][PF6], agit. 25 min - LC-UV 4400 [Vidal, 2007]
Agua de río Vm: 2 mL, 3 µL tolueno,
agit. 15 min
99-
100 GC-MS 50
[Okanouchi,
2008]
Agua Vm: 20 mL (pH 2, 1% etanol),
10 µL [C6MIM][PF6], agit. 37 min 8-98 LC-UV
200-
10000 [Vidal, 2010]
MALLE
Lago y
residual
Vm: 15 mL (1,5 mL metanol)
100 µL propanol, 40ºC, 120 min
60-
104
LC-APPI-
MS/MS 3-53 [Rodil, 2009]
HF-LPME
Orina
HF con tolueno, sumergida
en 1 mL de muestra,
15 min con agitación
89-
99 GC-MS 17-33
[Kawaguchi,
2009]
DLLME
Agua de
mar
Vm: 5 mL (pH 4, 0,5 g NaCl)
Disolvente: 1 mL acetona
(60 µL cloroformo)
65-
169 GC-MS 108-110
[Tarazona,
2010]
Como colofón a la revisión realizada en relación con la determinación de
filtros solares en muestras de agua, es preciso destacar que para determinados
compuestos (ej. EHMC, OCR y, en ocasiones, BP-3) los LOQs alcanzados vienen
condicionados por la señal de los blancos [Balmer, 2005] [Rodil, 2008-A] [Rodil,
2008-B] y no por la eficacia de la técnica de extracción o la sensibilidad de la
técnica de determinación.
Introducción-Filtros solares
67
3.1.2. Muestras sólidas
La extracción y recuperación de compuestos orgánicos de matrices
sólidas tales como sedimentos, suelos y lodos puede resumirse en cuatro etapas:
- Desorción de los analitos de los sitios activos de la matriz.
- Difusión de los analitos a través de la matriz.
- Solubilización de los analitos en el extractante.
- Recolección de los extractos con los analitos de interés.
La interacción entre los analitos y las muestras sólidas suele ser mucho
más intensa que en el caso de las matrices líquidas, de modo que es necesario
emplear técnicas de extracción más enérgicas, lo cual disminuye la selectividad
del proceso y, normalmente, hace necesario la consideración de etapas
posteriores de limpieza.
Entre las técnicas empleadas en esta Tesis Doctoral para la determinación
de filtros solares en matrices sólidas ambientales se encuentran: la dispersión de
la matriz en fase sólida (MSPD) y la extracción con disolventes presurizados
(PLE), cuyos fundamentos y aplicaciones se presentan a continuación.
3.1.2.1. Dispersión de la matriz en fase sólida (MSPD)
La dispersión de la matriz en fase sólida (MSPD) fue desarrollada en
1989 por Barker [Barker, 1989]. La MSPD se basa en homogeneizar y dispersar
una pequeña cantidad de muestra con ayuda de un mortero sobre un soporte
sólido. A continuación, la matriz homogeneizada se introduce en un cartucho,
normalmente de propileno, y los analitos son eluidos con un disolvente
apropiado. MSPD permite integrar la purificación en la misma etapa que la
extracción, introduciendo un co-adsorbente en el fondo del cartucho que
retenga los compuestos interferentes. La distribución de los analitos entre la
muestra dispersada y el disolvente de elución depende de las constantes que
regulan la partición (fase estacionaria líquida) o el equilibrio de adsorción (fase
Introducción-Filtros solares
68
estacionaria sólida). La eficacia y la selectividad en MSPD depende de varios
factores entre los cuales destacan la selección del material dispersante, la
utilización de co-adsorbentes, el tipo de disolvente y la secuencia de elución.
Naturaleza del soporte sólido y de la fase enlazada.
Los materiales derivados de la sílica son los más empleados para la
disrupción de la matriz en MSPD, ya que presentan la ventaja de poseer grupos
silanoles no enlazados, tanto en la superficie de las partículas como en los
poros, que interaccionan con el agua de la muestra, actuando a su vez como
agentes desecantes en el caso de matrices semi-sólidas. Dentro de los
adsorbentes denominados de fase reversa, el más empleado es el octadecilsilano
(C18), frente a otros como el C8 y el C30. Las partículas de sílica actúan como
dispersantes, mientras que el C18 solubiliza los componentes de la matriz sobre
su superficie. Con este tipo de fases es posible obtener extractos relativamente
libres de grasas para muestras de músculo [Kubala-Drinic, 2003] [Le Boulaire,
1997], hígado [Crescenzi, 2001] [Horne, 1998], riñón [Ruiz, 2005] y pescado con
alto contenido lipídico [Tolls, 1999] [García-Reyes, 2007] [Canosa, 2008], usando
acetonitrilo como disolvente. En la actualidad, las fases reversas basadas en
cadenas hidrocarbonadas se están sustituyendo por amino‐propil‐sílica o por
aminas primarias‐secundarias (PSA) [García-Reyes, 2007] [Ferrer, 2005] [Cunha,
2007] para reducir todavía más el contenido lipídico en los extractos
procedentes de muestras biológicas de origen animal. El Florisil (MgSiO3), la
alúmina (Al2O3) y la sílica (SiO2) son adsorbentes denominados de fase normal
utilizados también para la extracción de pesticidas, herbicidas y contaminantes
prioritarios en matrices biológicas [Gómez-Ariza, 2002] [Martínez, 2005] y frutas
[Hu, 2006] mediante MSPD. Además, se han usado como dispersantes de
muestras ambientales (lodos de depuradora, sedimentos, polvo de
aspiradora,…) para la extracción de contaminantes tanto prioritarios como
emergentes en MSPD [Pena, 2007] [Blanco, 2006] [Shen, 2006]. Estos adsorbentes
pueden ser utilizados tal y como se encuentran comercialmente o modificados
mediante la adición de agua, ácidos o bases en función de las características y el
Introducción-Filtros solares
69
comportamiento que queramos conferirles [Carro, 2005].
Naturaleza del co‐adsorbente
Generalmente debe ser de distinta naturaleza que el sólido utilizado para
dispersar la muestra. El co-adsorbente actúa reteniendo las interferencias,
normalmente las de mayor contenido lipídico [Pensado, 2005], o también puede
destruirlas cuando presenta algún tipo de modificación química [Carro, 2005]
[Canosa, 2008].
Naturaleza de la matriz de la muestra
Los componentes de la matriz dispersada se mueven a través de la fase
cromatográfica, contenida en el cartucho de MSPD, lo que teóricamente, hace
posible un fraccionamiento en función de la naturaleza de estos compuestos, así
como de las sustancias interferentes.
Tipo de disolvente y secuencia de elución
Al igual que en cromatografía, o en SPE, la polaridad del disolvente es de
gran importancia a la hora de determinar qué analitos eluyen del cartucho de
MSPD y en qué orden lo hacen. La correcta elección del disolvente y el diseño
del perfil de elución permite obtener extractos libres de impurezas en base a la
retención de las mismas en la fase estacionaria [Pensado, 2005], o mediante una
primera elución para retirarlas del cartucho de MSPD [Canosa, 2007] [García,
2007], previamente a la extracción de los analitos.
En esta Tesis doctoral, MSPD ha sido la técnica elegida para la extracción
de filtros solares en muestras de polvo [Negreira, 2009-C]. De acuerdo con la
revisión bibliográfica realizada, este trabajo constituye la primera referencia
describiendo la presencia de filtros UV en atmósferas interiores, así como una
de las primeras aplicaciones de MSPD a la determinación de estos compuestos
en matrices ambientales. En trabajos previos del grupo de investigación se ha
demostrado la aplicabilidad de MSPD a la determinación de productos de
Introducción-Filtros solares
70
cuidado personal [Canosa, 2007] y retardantes de llama [García, 2007] en
muestras de polvo.
3.1.2.2. Extracción con disolventes presurizados (PLE)
La extracción con disolventes presurizados (PLE), también denominada
extracción acelerada con disolventes (ASE), es una técnica de preparación de
muestra que combina el uso de temperaturas (50-200 ºC) y presiones (1500-2000
psi) elevadas, para extraer rápida y eficazmente los analitos de matrices sólidas
o semisólidas [Garrido-López, 2005]. Los parámetros fundamentales que afectan
a la eficacia de extracción, y que deben tenerse en cuenta a la hora de
desarrollar un método de PLE, son los siguientes:
Temperatura
Tiene que ser suficientemente elevada como para aumentar las
recuperaciones y favorecer la cinética de extracción, sin degradar a los
compuestos objeto de estudio [ConchaGraña, 2004]. Normalmente, es superior
al punto de ebullición del disolvente, pero ligeramente inferior a su punto
crítico. Al aumentar la temperatura, el disolvente disminuye su viscosidad y
penetra con mayor facilidad en los poros de la matriz, favoreciendo la difusión
de los analitos, es decir la cinética de extracción. De este modo, la eficacia de
extracción se incrementa, minimizando el volumen de disolvente empleado
[Richter, 1996].
Disolvente
Debe ser capaz de solubilizar los analitos sin arrastrar el resto de los
componentes de la matriz. Las mezclas de disolventes de diferentes polaridades
pueden ser usadas para la extracción de un amplio rango de familias de
compuestos. La mayoría de disolventes pueden emplearse en PLE, incluidos
agua y mezclas acuosas tamponadas. No se recomiendan ácidos fuertes debido
a su reacción con el acero, pero se pueden usar ácidos débiles, como ácido
Introducción-Filtros solares
71
acético o fosfórico, añadidos a agua o a disolventes polares en porcentajes de
hasta un 10% (v/v).
Presión
Ésta debe ser elevada para mantener el disolvente en estado líquido a la
temperatura de trabajo. Sin embargo, no es un factor que afecte en gran medida
a la eficacia de extracción, de modo que puede ser fijado de antemano [Camel,
2001]. El modo más usual de hacer la extracción es el modo estático, en el que el
disolvente es introducido en la celda y ésta se mantiene a presión constante un
tiempo determinado. Tras esta etapa, la celda se vacía recogiendo todo el
extracto en un vial colector. En el modo dinámico, el disolvente está pasando
continuamente a un flujo constante a través de la celda presurizada. En este
caso, la extracción es más efectiva, pero presenta el inconveniente de
incrementar el volumen del extracto [Camel, 2001] y además, el equipo necesita
una válvula restrictora que permita operar en ese modo.
Tiempo
El aumento del tiempo estático a elevadas temperaturas favorece la
difusión de los analitos al disolvente de extracción evitando su retención en la
matriz.
Número de ciclos
El uso de varios ciclos de extracción estáticos fue desarrollado para
introducir fracciones nuevas de disolvente durante el proceso de extracción. El
uso de varios ciclos de extracción es útil para muestras difíciles de penetrar o
que presenten muy alta concentración de analito.
Porcentaje de flush
Después de cada ciclo de extracción en modo estático, se hace pasar un
volumen de disolvente a través de la celda, expresado en porcentaje (% flush)
de su volumen interior, para arrastrar posibles trazas de los analitos que
Introducción-Filtros solares
72
pudiesen quedar en ella. Ese volumen, en el caso de que se usen dos o más
ciclos de extracción, es dividido entre el número de los mismos [Richter, 1996].
Equipamiento para PLE
La técnica de PLE fue comercializada en 1995 por la compañía Dionex
[Fidalgo-Used, 2007] y el modelo más utilizado es el extractor ASE 200, cuya
fotografía se muestra en la Fig. 9.
Figura 9: Fotografía del equipo de extracción ASE 200
Independientemente del modelo y la casa comercial, los equipos de PLE
constan de:
- una bomba para impulsar el disolvente,
- un horno donde se introducen las celdas de acero para
mantenerlas a la temperatura seleccionada,
- un vial colector en donde se recoge el extracto líquido,
- nitrógeno para purgar la celda una vez terminado el proceso de
extracción.
El esquema de un equipo de extracción con disolventes presurizados se
muestra en la siguiente figura, Fig. 10.
Introducción-Filtros solares
73
Figura 10. Equipo de extracción con disolventes presurizados.
Etapas del proceso de extracción
Las etapas principales del proceso de extracción son:
- Preparación de muestra. Con esta etapa se pretende reducir el
tamaño de partícula para aumentar la superficie de contacto entre
la muestra y el disolvente, evitando también la agregación de las
partículas de la muestra. Para ello, se utilizan agentes dispersantes
como la arena o la tierra de diatomeas. La muestra debe secarse
para evitar la presencia de agua que dificulte la penetración de
disolvente en sus poros, sobre todo cuando se trata de extraer
analitos poco polares con disolventes apolares. Para ello, se
aconseja la liofilización de la muestra, el secado en horno o la
adición de agentes desecantes.
- Preparación de celda. Se colocan filtros de celulosa y/o fibra de
vidrio en los extremos de la celda con el fin de evitar la obturación
de los conductos del extractor de PLE. La celda se rellena con la
muestra y con una matriz inerte (tierra de diatomeas, arena…)
para ocupar el volumen muerto. Además se pueden introducir co-
adsorbentes para purificar el extracto en la misma etapa.
- Extracción estática. La celda se introduce en el horno a la
temperatura de extracción, se rellena con disolvente y se presuriza.
También se puede precalentar la celda previamente y mantenerla a
esa temperatura un tiempo determinado. Después de recoger el
Introducción-Filtros solares
74
disolvente con los analitos en el vial colector, se bombea disolvente
fresco a través de los conectores y celda (% de flush). Finalmente, se
purga la celda con nitrógeno para arrastrar el disolvente que
permanece impregnando la muestra, el dispersante y, en su caso, el
co-adsorbente en la celda.
Aplicaciones
PLE es equivalente a otras metodologías de extracción, como Soxhlet,
siendo utilizada incluso como método de referencia por la Agencia de
Protección Medioambiental estadounidense (EPA) en uno de sus métodos
(método 3545A)3 para la extracción de compuestos volátiles y semivolátiles en
matrices medioambientales sólidas. La aplicación de PLE a la extracción de
filtros solares en muestras sólidas se ha centrado en sedimentos y lodos de
depuradora. En relación con la primera matriz, Rodil y col. [Rodil, 2008-C]
extrajeron EHS, HMS, IAMC, EHMC, 4-MBC, BP-3, EHPABA y OCR
mezclando 4 g de muestra con 1 g de sulfato sódico anhidro que introdujeron
en una celda conteniendo 2 g de sílica gel y la misma cantidad de cobre en
polvo. La extracción se llevó a cabo en 4 ciclos de 5 min, a 160 ºC y 100 bar
usando acetato de etilo:hexano (80:20) como disolvente de extracción. Después
de concentrar el extracto, añadieron N,O-bis(trimetilsilil)trifluoroacetamida
(BSTFA) para la derivatización de los salicilatos y la BP-3, antes de su
determinación mediante GC-MS. Las recuperaciones obtenidas oscilaron del 73
al 128% y los LODs variaron entre 2 y 20 ng g-1 sin derivatizar y de 2 a 5 ng g-1
derivatizando los analitos.
Nieto y col. [Nieto, 2009] también emplearon PLE como técnica de
extracción y purificación para la determinación de filtros solares mediante
UPLC-MS/MS en lodos de depuradora. La celda fue cargada con 1 g de óxido
de aluminio, 1 g de muestra y, otra vez, óxido de aluminio para rellenar el
3 U.S.Environmental Protection Agency (EPA), disponible en:
http://www.epa.gov/ (acceso en diciembre de 2009).
Introducción-Filtros solares
75
volumen muerto restante. Las condiciones de la extracción fueron de 2 ciclos de
5 min con metanol, seguidos de 2 ciclos con agua (pH 7): metanol (1:1) durante
un tiempo de 5 min, a 100ºC y 140 bar. Las recuperaciones oscilaron desde el
30% (BP-1) al 108% (EHPABA).
Rodil y col. [Rodil, 2009-D] desarrollaron un método para la
determinación de filtros UV en muestras de lodos mediante LC-(APPI)MS/MS.
Su principal novedad radica en el uso de membranas poliméricas no porosas en
combinación con PLE. La muestra (0,5 g lodo) y 1 mL de disolvente de
extracción se encerraron en una membrana preparada con polietileno de baja
densidad (LDPE) y se introdujeron en un equipo convencional de PLE. La
extracción se lleva a cabo a 70ºC, 4 ciclos de 5 min con acetato de etilo:hexano,
(1:3). El extracto evaporado se reconstituyó con metanol: agua (1:1) antes de su
inyección en LC-MS. La principal ventaja de este proceso es la reducción de
tiempo y disolvente al combinar extracción y limpieza en un único paso. Sin
embargo, la extracción no fue cuantitativa, siendo necesario el uso de adiciones
estándar sobre la muestra para poder determinar los niveles de analitos en
muestras de lodos.
Wick y col. [Wick, 2010] mezclaron 0,2 g de lodo liofilizado con arena de
mar purificada. La extracción se llevó a cabo en 4 ciclos de 10 min a 80 ºC con
agua:metanol (1:1). Los autores emplearon una etapa posterior de purificación
mediante SPE para la cual, el extracto resultante se diluyó a 800 mL con agua
ajustada a pH 6 antes de ser pasado a través de un cartucho Oasis HLB (200
mg). La elución se realizó con 4 x 2 mL de una mezcla de metanol:acetona
(60:40) y la determinación tuvo lugar por LC-MS/MS, obteniéndose valores de
recuperación comprendidos entre 18% y 196% y LOQs entre 2,5 y 25 ng L-1. El
método desarrollado sólo incluye 4 benzofenonas (BP-1, BP-2, BP-3 y BP-4),
junto con otros compuestos de la familia de los benzotriazoles y, probablemente
no proporciona recuperaciones cuantitativas para compuestos más lipofílicos
como 4-MBC, EHMC y OCR.
Introducción-Filtros solares
76
En esta tesis doctoral también se desarrolló una metodología para la
determinación de filtros solares en lodos de depuradora empleando PLE como
técnica de extracción. Los objetivos del estudio realizado ha sido la obtención de
recuperaciones cuantitativas con un nivel de selectividad adecuado para la
determinación de los compuestos mediante GC-MS.
3.1.2.3. Otras técnicas de extracción
La preparación de muestra para la determinación de filtros solares en
sedimentos y, sobre todo, en lodos implica la combinación de metodologías
eficaces de extracción, con un grado adecuado de selectividad, teniendo en
cuenta la complejidad de la matriz de partida (sobre todo en el caso de lodos), y
la técnica empleada en la etapa de determinación. A continuación, se resume la
bibliografía encontrada para la determinación de filtros solares en sedimentos,
suelos y lodos, excluyendo los trabajos de PLE que han sido citados en la
página anterior.
Jeon y col. [Jeon, 2006] determinaron filtros UV en suelo extrayendo 10 g
de muestra, mezclada con sulfato sódico (10 g), con 20 mL de metanol, durante
20 min. A continuación, el extracto se concentra (ca. 3 mL) se agita junto con 1
mL de una disolución acuosa de NaCl (5%) y 5 mL de acetato de etilo y se
congela a -30ºC para la separación de la fase orgánica que se evapora para ser
derivatizada e inyectada en GC-MS. Las recuperaciones obtenidas variaron de
60% a 115% y el LOQ alcanzado fue de 500 ng g-1 para todos los compuestos.
Plagellat y col. [Plagellat, 2006] determinaron tres filtros solares (EHMC,
OCR y 4-MBC) en muestras de lodos de depuradora utilizando extracción
líquido-líquido. Para ello, los lodos espesados, y sin liofilizar, se mezclaron con
3 g de NaCl y 20 mL de pentano:acetona (1:1, v/v), se agitaron durante 30 min y
luego se llevaron a cabo otras 2 extracciones sucesivas con 20 mL pentano:dietil
Introducción-Filtros solares
77
éter (1:1, v/v) y dietil éter:diclorometano (4:1, v/v). Después de centrifugar, las
diferentes fracciones se concentraron a sequedad, se redisolvió el extracto con 1
mL de hexano y se transfirió a una columna con 5 g de sílica gel para llevar a
cabo el clean-up. Los analitos fueron recogidos con 50 mL de hexano:dietil éter
(9:1, v/v) después de descartar 2 fracciones anteriores de 20 mL de hexano y 20
mL de hexano:dietil éter. Para su inyección en GC-EI-SIM-MS fue necesario la
evaporación del extracto y su reconstitución con 1 mL de acetato de etilo. Las
recuperaciones se situaron entre 88 y 101% y los LODs ente 2 y 6 ng g-1; sin
embargo, el método es difícil de automatizar y consume del orden de 150 mL de
disolventes orgánicos por muestra.
En la siguiente tabla, se hace un resumen de las condiciones de
extracción y purificación utilizadas en la determinación de filtros solares en
muestras de sedimentos y lodos, Tabla 16.
Tabla 16 : Resumen de la metodología analítica para la determinación de filtros solares
en sedimentos y lodo de depuradora.
Muestra Analitos Técnica
extracción Purificación
Técnica
Det. Referencia
Sedimento
BP, BP-1, BP-3, BP-8 LLE - GC-MS [Jeon, 2006]
EHS, HMS, 4-MBC,
BP-3, IAMC, EHMC,
EHPABA, OCR
PLE (extracción y
purificación) GC-MS
[Rodil,
2008-C]
Lodo
4-MBC, OCR, EHMC LLE Sílica gel GC-MS [Plagellat,
2006]
EHS, HMS, 4-MBC,
BP-3, IAMC, EHMC,
EHPABA, OCR
PLE
Difusión
membrana no
porosa
LC-
MS/MS
[Rodil,
2009-D]
BP-1, BP-3, BP-8,
OCR, EHPABA
PLE (extracción y
purificación)
UHPLC-
MS/MS
[Nieto,
2009]
BP-1, BP-2,BP-3, BP-4 PLE SPE LC-
MS/MS
[Wick,
2010]
Introducción-Filtros solares
78
3.1.3. Muestras de biota
En el caso de biota (Ej. músculo o órganos de pescado) la etapa más
importante en la preparación de muestra es la separación de los analitos y los
lípidos, sobre todo en el caso de emplear cromatografía de gases como técnica
de determinación. Para llevar a cabo ese fraccionamiento, se suele utilizar la
cromatografía de exclusión por tamaños (SEC). Aunque efectiva, SEC implica el
consumo de ingentes volúmenes de disolventes orgánicos.
Balmer y col. [Balmer, 2005] analizaron muestras de pescado para la
determinación de filtros solares. Estas muestras fueron homogeneizadas con
sulfato sódico y se extrajeron con diclorometano:ciclohexano (1:1). Para el
proceso de clean-up se utilizó cromatografía de exclusión por tamaños (SEC)
empleando una columna Biobeads S-X3 y diclorometano:ciclohexano (35:65)
como fase móvil. Para la determinación se utilizó GC-MS obteniendo
recuperaciones de 93% a 115% y LODs de 3 a 380 ng g-1. La misma metodología
fue empleada por Buser y col. [Buser, 2006] en un estudio posterior para evaluar
la acumulación de filtros UV en pescado de ríos y lagos en Suiza.
Meinerling y col. [Meinerling, 2006] determinaron BP-3, 4-MBC, OCR y
EHMC en tejidos de pescado. La extracción se llevó a cabo mediante Soxhlet
empleando 200 mL de n-hexano:acetona (9:1, v/v) durante aprox. 3 h. El
extracto frío se redujo a sequedad usando un rotavapor y fue disuelto en 20 mL
de n-hexano:acetona (9:1, v/v). A continuación se purificó empleando SEC y
SPE. La cuantificación fue desarrollada por LC-MS. Las recuperaciones
obtenidas se encontraron en el rango de 86% a 108%. El límite de cuantificación
determinado para cada analito fue de 8 ng g-1 de muestra fresca.
Zenker y col. [Zenker, 2008] desarrollaron una metodología para el análisis
de músculo de pescado agitándolo vigorosamente con una mezcla de acetato de
etilo, n-heptano y agua (1:1:1). Después de centrifugar, el sobrenadante fue
Introducción-Filtros solares
79
concentrado y disuelto con etanol para posteriormente llevar a cabo la limpieza
mediante HPLC con una columna en fase reversa. La elución de los analitos se
realizó con una mezcla de metanol y agua (70:30) recogiendo dos fracciones: la
primera contenía los compuestos más polares (BP-1, BP-2, BP-8, BP-4, Et-PABA)
que fueron inyectados en LC-MS, y en la segunda se recuperaban los más
lipofílicos (BP-3, EHMC, 4-MBC) que fueron determinados mediante GC-MS.
Las recuperaciones obtenidas variaron desde 76% a 99% y los LODs de 11 a 36
ng g-1 para BP-3, EHMC y 4-MBC y de 78 a 205 ng g-1 para el resto de
compuestos. Fent y col. [Fent, 2010-B] aplicaron el mismo método para el
análisis de pescado en Suiza con resultados similares.
Mottaleb y col. [Mottaleb, 2009] mezclaron tejido de pescado con 10 mL de
acetona, concentraron y reconstituyeron el extracto con hexano:acetona (65:35,
v/v). A continuación, este extracto se purificó empleando un adsorbente en fase
normal (sílica) y SEC. La determinación se llevó a cabo utilizando GC-MS/MS
obteniéndose recuperaciones de 57% a 87% y LODs de 16 a 120 ng g-1. Para
muestras con bajo contenido lipídico, los autores prescinden de la purificación
mediante SEC, obteniendo en este caso recuperaciones entre 98% y 101%.
Kwon y col. [Kwon, 2009] emplearon LLE y SPE para extraer y purificar
muestras de hígado de pescado antes de la determinación de BP-3, junto con
otros contaminantes emergentes, mediante LC-MS/MS con ionización mediante
electrospray en modo negativo y positivo. El extracto primario en n-hexano fue
llevado a sequedad, reconstituido con acetonitrilo y diluido con 50 mL de agua.
Esta disolución se concentró con un cartucho de SPE (Oasis HLB) que
posteriormente fue eluido con metanol. La determinación mediante LC-MS
proporcionó un LOD de 8 ng g-1 con recuperaciones entre 72% y 77% para BP-3.
En la siguiente tabla, se hace un resumen de las condiciones de
extracción y purificación para la determinación de filtros solares en muestras de
biota, Tabla 17. En general, los métodos propuestos en la bibliografía para esta
Introducción-Filtros solares
80
matriz son multietapa y presentan un bajo grado de automatización, así como
un elevado consumo de disolventes orgánicos.
Tabla 17: Resumen de la metodología analítica para la determinación de filtros solares
en pescado.
Analitos Técnica
extracción Purificación
Técnica
Det. Referencia
BP-3, 4-MBC, EHMC, OCR ESL SEC GC-MS
[Balmer, 2005]
4-MBC, OCR [Buser, 2006]
BP-3, 4-MBC, OCR, EHMC Soxhlet SEC + SPE LC-MS [Meinerling, 2006]
BP-1, BP-2, BP-8, BP-4, Et-PABA ESL RP-HPLC
LC-MS [Zenker, 2008],
[Fent, 2010-B] BP-3, 4-MBC, EHMC GC-MS
BP, 4-MBC, OCR ESL SEC + SPE GC-MS [Mottaleb, 2009]
BP-3 ESL SPE LC-MS [Kwon, 2009]
ESL, extracción sólido-líquido
RP, fase reversa
Introducción-Filtros solares
81
3.2. Técnicas de determinación
La determinación de niveles traza de filtros solares en muestras
ambientales complejas se basa normalmente en la utilización de cromatografía
de gases (GC) o líquidos (LC) combinadas con espectrometría de masas simple
(MS) o en tándem (MS/MS). Otras técnicas consideradas para la determinación
de estos compuestos en productos de cuidado personal, ej. LC con detección
UV-visible, no ofrecen prestaciones adecuadas en términos de sensibilidad y
selectividad para poder abordar el estudio de matrices ambientales.
La elección de la técnica de separación (GC o LC) viene condicionada
fundamentalmente por los pesos moleculares y los grupos funcionales
presentes en la estructura de los analitos considerados. A modo de ejemplo, la
BP-3 y sus metabolitos (ej. BP-1), presentan grupos fenólicos en su estructura lo
que provoca un ensanchamiento de los correspondientes picos en GC cuando se
aborda su determinación directa. Este mismo comentario es también válido
para los salicilatos. Otros filtros UV que contienen grupos más polares en sus
estructuras, tales como la BP-4, son sólo cuantificables mediante LC. La
afirmación anterior es también aplicable a filtros de mayor tamaño molecular,
tales como los derivados de la triazona y del benzotriazol, así como las
polisiliconas.
Puesto que LC se emplea normalmente en combinación con MS/MS es
necesario tener en cuenta, no sólo la capacidad de separación de la etapa
cromatográfica sino también el rendimiento de la etapa de ionización, que en
ocasiones, controla los LOQs alcanzables. A modo de ejemplo, los filtros de la
familia de salicilatos presentan eficacias de ionización muy bajas en sistemas
equipados con interfases de electrospray (ESI) [Rodil, 2009-E].
A continuación se comentan las aplicaciones de las técnicas anteriores a
la determinación de filtros UV en matrices medioambientales.
Introducción-Filtros solares
82
3.2.1. Cromatografía de gases acoplada a espectrometría de
masas simple (GC-MS) y en tándem (GC-MS/MS)
Sin duda, GC ha sido la técnica más empleada para la determinación de
filtros solares en muestras medioambientales dada su disponibilidad en la
mayoría de laboratorios. Las diferencias básicas entre los estudios realizados
radican en el tipo de analizador de masas empleado, las características de la
columna cromatográfica, el modo de inyección y la utilización o no de
reacciones de derivatización para mejorar la detectabilidad de ciertos analitos.
A continuación, se describe de forma resumida el fundamento del acoplamiento
GC-MS, así como GC-MS/MS, para luego comentar algunas aplicaciones
representativas a la determinación de filtros UV en matrices medioambientales.
El acoplamento GC-MS data de la década de los 70, y revolucionó el
análisis de mezclas complejas de compuestos orgánicos debido a que combina
el elevado poder de resolución que proporciona la cromatografía de gases, con
la alta sensibilidad y la información estructural de la espectrometría de masas.
Mediante esta técnica se obtienen registros tridimensionales, es decir, para cada
tiempo de retención, obtenemos un espectro de masas de las especies que
emergen de la columna cromatográfica.
En GC, los compuestos en estado vapor se separan en la columna
cromatográfica en función de su distribución entre la fase estacionaria y la fase
móvil. Debido al bajo flujo de fase móvil empleado en las columnas capilares, es
posible la conexión directa de la columna cromatográfica con la fuente de
ionización del espectrómetro de masas. Posteriormente, los iones pasan al
analizador de masas, el cual está sometido a un alto vacío para evitar la colisión
y reconstrucción de los fragmentos cargados generados tras la ionización. Una
vez son seleccionados por el analizador de masas, se registran las intensidades
correspondientes a cada relación m/z.
Para la determinación de filtros UV mediante GC-MS se hace uso de la
ionización por impacto electrónico (EI). Las fuentes de EI consisten en un
filamento de wolframio, que emite electrones acelerados hacia un ánodo. Estos
Introducción-Filtros solares
83
electrones chocan con las moléculas del efluente cromatográfico dado que sus
trayectorias son perpendiculares. A continuación, los iones generados pasan al
analizador de masas. En esta Tesis Doctoral se emplearon sistemas GC-MS
equipados con dos tipos de analizadores de masas diferentes (trampa de iones y
cuadrupolo):
- Trampa de iones
Una trampa de iones, Fig. 11, consta de tres electrodos que forman una
cavidad en la que tiene lugar el proceso completo de ionización, fragmentación,
almacenamiento y separación de los iones. Al electrodo central se le aplica un
potencial de radiofrecuencia (RF) que crea un campo eléctrico hiperbólico
tridimensional, en el que los iones son atrapados en órbitas estables. A medida
que se aumenta el voltaje de RF, las trayectorias de los iones se hacen inestables
en el sentido en que aumenta su relación masa/carga, y son expulsados de la
trampa hacia el multiplicador de electrones.
Para evitar las reacciones ión-ión e ión-molécula, causa de espectros mal
resueltos, se aplica un voltaje adicional de RF (voltaje de modulación axial) que
aumenta la resolución entre masas. Mediante este potencial, los iones ocupan
mayor volumen en el interior de la trampa.
Figura 11: Esquema de una trampa de iones.
Introducción-Filtros solares
84
- Cuadrupolo
El analizador de masas tipo cuadrupolo, Fig. 12, consta de dos pares de
cilindros o barras a cada uno de los cuales se le aplica una combinación de
potenciales, de radiofrecuencia (RF) y de corriente continua (DC), que se van
variando de forma que sólo los iones con determinada relación m/z sean
capaces de atravesar completamente el filtro de masas. Los potenciales
aplicados a los dos pares de cilindros o barras son iguales pero de signo
opuesto.
Figura 12: Esquema de un cuadrupolo. Una de las ventajas de estos analizadores es la posibilidad de trabajar en
modo SIM (selected ion monitoring), cuando el objetivo principal del análisis es
maximizar la sensibilidad. Por su parte, las trampas de iones permiten trabajar
en modo MS/MS sin coste adicional de la instrumentación, frente al empleo de
triples cuadrupolos cuando se pretende trabajar con espectrometría de masas en
tándem.
En las tablas siguientes (Tablas 18 y 19) se recogen las aplicaciones más
relevantes en las que se describe el uso de GC-MS para la determinación de
filtros solares en muestras medioambientales.
Introducción-Filtros solares
85
Tab
la 1
8: R
esum
en d
e la
s ap
licac
ione
s de
GC
-MS
para
la d
eter
min
ació
n de
filt
ros
sola
res
en m
uest
ras
acuo
sas.
Mat
riz
acu
osa
An
alit
os
Téc
nic
a d
e ex
trac
ción
C
olu
mn
a em
ple
ada
LO
D (n
g L
-1)
Rec
up
. (%
) R
efer
enci
a
Pis
cina
y b
año
BP
-3, E
HP
AB
A
SPM
E
DB
-5-M
S
(30
m x
0,3
2 m
m, 1
µm
)
0,2-
1 82
-99
[Lam
brop
oulo
u, 2
002]
SP
E
2 93
-97
Río
y b
año
BP
-3, E
HS,
EH
MC
,
4-M
BC
, OC
R
SPE
SE
54
(25
m x
0,3
2 m
m)
2 57
-90
[Poi
ger,
200
4]
Pis
cina
y d
uch
a B
P-3
, 4-M
BC
SP
E
DB
-5-M
S
(30
m x
0,2
5 m
m, 0
,25
µm)
0,2-
0,4
95-9
9 [G
ioka
s, 2
004]
Río
y b
año
BP
-3, E
HM
C, 4
-MB
C
LL
E
DB
-5-M
S
(30
m x
0,2
5 m
m, 0
,25
µm)
2-30
95
-102
[G
ioka
s, 2
005]
Sup
erfi
cial
y re
sid
ual
BP-
3, E
HM
C,
4-M
BC
, OC
R
SPE
SE54
(25
m x
0,3
2 m
m, 0
,25
µm)
BG
B-5
(30
m x
0,2
5 m
m, 0
,25
µm)
2 (s
upe
rfic
ial)
10 (r
esid
ual
)
78-1
29
[Bal
mer
, 200
5]
Res
idua
l E
HM
C, 4
-MB
C, O
CR
L
LE
D
B-5
-MS
(50
m x
0,2
0 m
m, 0
,33
µm)
3-14
75
-91
[Ku
pper
, 200
6]
Río
, lag
o
y re
sid
ual
B
P-3,
BP
-1, B
P-8
L
LE
U
ltra
2
(30
m x
0,2
mm
, 0,3
3µm
) 5
76-1
13
[Jeo
n, 2
006]
Río
B
P-3
SB
SE
DB
-5-M
S
(30
m x
0,2
5 m
m, 0
,25
µm)
1 11
0-11
4 [K
awag
uch
i, 20
06]
Río
, mar
y pi
scin
a
BP-
3, H
MS,
EH
MC
,
4-M
BC
, EH
PA
BA
, OC
R
SPE
H
P-5
-MS
(30
m x
0,2
5mm
, 0,2
5 µm
) 13
-266
50
-98
[Cu
der
man
, 200
7]
Introducción-Filtros solares
86
Tab
la 1
8 co
nt: R
esum
en d
e la
s ap
licac
ione
s de
GC
-MS
para
la d
eter
min
ació
n de
filt
ros
sola
res
en m
uest
ras
acuo
sas.
Mat
riz
acu
osa
An
alit
os
Téc
nic
a d
e ex
trac
ción
C
olu
mn
a em
ple
ada
LO
D (n
g L
-1)
Rec
up
. (%
) R
efer
enci
a
Res
idua
l B
P-3
, EH
MC
, 4-M
BC
, OC
R
SPE
D
B-5
-MS
(30
m x
0,2
5 m
m, 0
,25
µm)
10
67-1
18
[Li,
2007
]
Río
B
P-3
, BP-
1 SB
SE
DB
-5-M
S
(30
m x
0,2
5 m
m, 0
,25
µm)
0,5-
1 10
2-12
8 [K
agaw
uch
i, 20
08-A
]
Río
B
P-3
L
PM
E
DB
-5-M
S
(30
m x
0,2
5 m
m, 0
,25
µm)
10
100
[Oka
nou
chi,
2008
]
Río
, lag
o y
resi
du
al
BP
-3, E
HS,
HM
S, 4
-MB
C,
IAM
C, E
HM
C, E
HP
AB
A, O
CR
SB
SE
HP-
5-M
S
(30
m x
0,2
5 m
m, 0
,25
µm)
0,2-
16
63-1
22
[Rod
il, 2
008-
B]
Lag
o 4-
MB
C, O
CR
SB
SE
HP
-5m
s
(30
m x
0,2
5 m
m, 0
,25
µm)
0,3-
4 50
-100
[H
auns
chm
idt,
2010
]
Río
E
HS,
BP
-3, 4
-MB
C, O
CR
SP
ME
R
tx-5
-MS
(30
m x
0,2
5 m
m, 0
,25
µm)
0,2-
2 72
,5-1
14
[Liu
, 201
0]
Mar
B
P-1,
BP
-3, B
P-8
D
LL
ME
T
R-5
-MS
(30
m x
0,2
5 m
m, 0
,25
µm)
32-3
3 65
-169
[T
araz
ona,
201
0]
Lag
o y
eflu
ente
B
P-3
, 4-M
BC
, EH
MC
, OC
R
ME
PS
HP-
5-M
S
(30
m x
0,2
5 m
m, 0
,25
µm)
34-8
7 61
-114
[M
oed
er, 2
010]
Introducción-Filtros solares
87
Tab
la 1
9: R
esum
en d
e la
s ap
licac
ione
s de
GC
-MS
para
la d
eter
min
ació
n de
filt
ros
sola
res
en m
uest
ras
sólid
as.
Mat
riz
An
alit
os
Téc
nic
a d
e ex
trac
ción
C
olu
mn
a em
ple
ada
LO
D (n
g g-
1 )
Rec
up
. (%
) R
efer
enci
a
Sed
imen
to
BP
-3, B
P-1
, BP
-8
LL
E
Ult
ra 2
(30
m x
0,2
mm
, 0,3
3µm
) 0,
1 60
-84
[Jeo
n, 2
006]
Sed
imen
to
BP
-3, E
HS,
HM
S, 4
-MB
C, I
AM
C,
EH
MC
, EH
PAB
A, O
CR
P
LE
H
P-5
MS
(30
m x
250
mm
, 0,2
5 µm
) 2-
6 73
-128
[R
odil,
200
8-C
]
Lod
o E
HM
C, 4
-MB
C, O
CR
L
LE
D
B-5
(60
m x
0,2
5 m
m, 0
,25
µm)
3-6
88-1
01
[Pla
gella
t, 20
06]
Pes
cad
o B
P-3
, 4-M
BC
, EH
MC
, OC
R
ESL
, SE
C
SE54
(25
m x
0,3
2 m
m, 0
,25
µm)
BG
B-5
(30
m x
0,2
5 m
m, 0
,25
µm)
3-38
0 93
-115
[B
alm
er, 2
005]
Pesc
ado
4-M
BC
, OC
R
ESL
, SE
C
- 3-
60
- [B
user
, 200
6]
Pes
cad
o B
P-3
, 4-M
BC
, EH
MC
E
SL
OPT
IMA
-5-M
S
(50
m x
0,2
mm
, 0,5
µm
)
11-3
6 72
-102
[Z
enke
r, 2
008]
6-50
70
-105
[F
ent,
2010
-B]
Pes
cad
o 4-
MB
C, O
CR
E
SL, S
EC
V
F-5
(30
m x
0,2
5mm
, 0,2
5 µm
)
5-17
98
-99
[Mot
tale
b, 2
009]
36-1
20
57-7
9 [M
otta
leb,
200
9]
Introducción-Filtros solares
88
En la determinación de compuestos polares mediante GC es habitual
recurrir al uso de reacciones de derivatización con objeto de: mejorar la
estabilidad térmica de los analitos, la resolución entre picos y, en ocasiones,
también la detectabilidad. Entre las reacciones de derivatización más frecuentes
utilizadas en combinación con cromatografía de gases figuran las sililaciones y
las acetilaciones. Las primeras consisten en la sustitución de un hidrógeno
activo (perteneciente a un grupo hidroxilo, carboxílico, amida, etc.) por un
grupo sililo, así se reduce la polaridad del analito y la posibilidad de formación
de enlaces de hidrógeno intermoleculares, además aumenta su volatilidad y
estabilidad térmica. Los agentes sililantes más utilizados son el bis‐(trimetilsilil)
trifluoroacetamida (BSTFA), el N‐metil‐N‐(trimetilsilil)trifuoroacetamida
(MSTFA) y el N‐(tert‐butildimetilsilil)‐N‐metiltrifluoroacetamida (MTBSTFA)
(Fig. 13). Las sililaciones transcurren en medio anhidro, por lo que la
derivatización va a realizarse después del proceso de extracción.
Figura 13: Estructura de los principales agentes sililantes utilizados para la
derivatización de los compuestos estudiados.
Por su parte, las reacciones de acetilación, usando anhídrido acético
como derivatizante, son adecuadas para disminuir la polaridad de especies
fenólicas y pueden llevarse a cabo tanto en medio acuoso como orgánico,
empleando en ambos casos una base como catalizador del proceso.
A continuación se resumen las aplicaciones en las que se ha usado GC en
combinación con reacciones de derivatización para la determinación de filtros
solares en matrices ambientales.
Jeon y col. [Jeon, 2006] determinaron diferentes benzofenonas por GC-MS
usando MSTFA como agente derivatizante. Se alcanzaron recuperaciones de
Introducción-Filtros solares
89
76% a 113% en muestras acuosas y de 60% a 84% en muestras de suelo. Los
límites de detección fueron de 5 ng L-1 para las muestras acuosas y de 0,1 ng g-1
para el suelo y los de cuantificación de 25 ng L-1 para agua y de 0,5 ng g-1 para
suelo.
Cuderman y col. [Cuderman, 2007] usaron el MSTFA como agente
derivatizante en la determinación de HMS y BP-3 en muestras acuosas para ser
analizadas por GC-MS. Los límites de detección obtenidos se encontraron en el
rango de 13 a 266 ng L-1.
Rodil y col. [Rodil, 2008-C] emplean BSTFA para la derivatización de EHS,
HMS y BP-3 después de su extracción de muestras de sedimentos y antes de la
determinación mediante GC-MS. Las recuperaciones obtenidas variaron del
73% al 128% y los LOD oscilaron entre 2 y 20 ng g-1 sin derivatizar y de 2 a 6 ng
g-1 derivatizando.
Tarazona y col. [Tarazona, 2010] utilizan BSTFA para la derivatización de
benzofenonas (BP-1, BP-3 y BP-8) en extractos de agua de mar antes de su
determinación mediante GC-MS. Los límites de cuantificación alcanzados
fueron de 108 a 110 ng L-1.
Kawaguchi y col. [Kawaguchi, 2008-A] desarrollaron un procedimiento para
la determinación de BP-1 y BP-3 en muestras de agua consistente en la
acetilación de los compuestos con anhídrido acético en medio básico, seguido
de su concentración y extracción con SBSE y posterior determinación mediante
GC-MS. La acetilación permitió mejorar la detectabilidad de los compuestos y
también la eficacia de extracción sobre los “Twister”. Ito y col. [Ito, 2009],
aplicaron la misma estrategia a la determinación de benzofenonas en muestras
de orina.
Introducción-Filtros solares
90
En la siguiente tabla, se indican las condiciones de derivatización y los
agentes derivatizantes empleados para la determinación de algunos filtros UV
considerados en esta Tesis doctoral, Tabla 20.
Tabla 20: Resumen de aplicaciones empleando agentes sililantes y anhídrido acético para
la derivatización de benzofenonas y salicilatos.
Analitos Matriz Condiciones de derivatización Autor
Derivatización durante la extracción
Anhídrido acético
BP-1, BP-3 Agua de río 100 µL + 1 mL carbonato potásico
(1M)
[Kawaguchi, 2008-
A]
BP-1, BP-3 Orina 20 µL + 50 µL carbonato potásico
(1 M) [Ito, 2009]
Derivatización después de la extracción
MSTFA
BP-1, BP-3, BP-
8
Agua río, lago y
suelo 50 µL, 80ºC, 30 min [Jeon, 2006]
HMS, BP-3 Agua residual
y recreacional 100 µL, 60ºC, 60 min [Cuderman, 2007]
BSTFA
BP-3, BP-1, BP-
8 Agua de mar 60 µL, 75 ºC, 30 min [Tarazona, 2010]
EHS, HMS,
BP-3 Sedimentos 50 µL, Tªamb, 1 h [Rodil, 2008-C]
En esta tesis se ha considerado la derivatización de varios filtros solares
(salicilatos y benzofenonas), usando N-(tert-butildimetilsilil)-N-
metiltrifluoroacetamida (MTBSTFA) en combinación con SPE [Negreira, 2008] y
N-metil-N-(trimetilsilil)trifluoroacetamida (MSTFA) cuando los analitos habían
sido concentrados previamente en una fibra de SPME [Negreira, 2009-A]. El
MTBSTFA presenta un grupo tert-butilo con gran impedimento estérico,
protegiendo el enlace silicio-oxígeno del ataque hidrolítico del agua y de ahí
Introducción-Filtros solares
91
que los derivados formados sean muy estables. Por su parte, el MSTFA presenta
una mayor reactividad pero da lugar a derivados menos estables.
3.2.2. Cromatografía líquida acoplada a espectrometría de
masas (LC-MS)
La cromatografía líquida de alta resolución (HPLC) nace en la década de
los sesenta y desde entonces ha desempeñado un papel fundamental en los
laboratorios analíticos debido a su enorme aplicabilidad a compuestos no
volátiles, polares, termosensibles, de alto peso molecular (cromatografía de
exclusión por tamaños), e incluso, iónicos (cromatografía de intercambio iónico)
[Rouessac, 2003] [Cela, 2002].
Su éxito se debe a la posibilidad de actuar de forma muy precisa sobre la
selectividad de la separación a través de la elección de la columna y de la
composición del eluyente, es decir, a sacar partido de las interacciones
analito/fase móvil/fase estacionaria. El acoplamiento LC-MS se empieza a
estudiar en los años 70, centrándose los veinte años posteriores en
compatibilizar las condiciones de operación de ambas técnicas y en la
innovación tecnológica de diferentes interfases. Cada interfase aplica una
aproximación diferente para resolver los dos problemas principales que plantea
este tipo de conexión:
- Eliminar la gran cantidad de gas y vapor procedente de la fase
móvil, antes de entrar a la región de alto vacío del espectrómetro de
masas.
- Transformar las moléculas en disolución en iones en fase gaseosa,
sin que se produzca su degradación térmica.
Actualmente, los fabricantes de instrumentación de LC-MS han optado
por ofertar sus equipos con varias interfases. Los más populares se incluyen
dentro de la modalidad de ionización a presión atmosférica (API), denominadas
Introducción-Filtros solares
92
electrospray (ESI), ionización química a presión atmosférica (APCI) e ionización
fotoinducida a presión atmosférica (APPI).
Las aplicaciones de LC-MS a la determinación de filtros solares en
muestras complejas suele basarse en la utilización de separaciones en fase
reversa, combinadas con un espectrómetro de masas equipado con un triple
cuadrupolo [Trenholm, 2008] [Kasprzyk-Hordern, 2008] [Rodil, 2008-B; Rodil, 2009-
C; Rodil, 2009-D] [Pedrouzo, 2009; Pedrouzo, 2010] [Nieto, 2009] [Wick, 2010]. Una
dificultad importante a la hora de generalizar la aplicación de los sistemas LC-
MS a la determinación de filtros UV es que no existe un único sistema de
ionización que proporcione eficacias elevadas para los analitos más
ampliamente estudiados en matrices ambientales. Rodil y col. [Rodil, 2009-E]
han comparado las interfases de ESI y APPI, observando que APPI ofrece
mejores límites de detección para los filtros solares menos polares y más
lipofílicos: 4-MBC, OCR, EHPABA y los salicilatos. Estos últimos presentan
muy bajo rendimiento por ESI; mientras que las benzofenonas (BP-3 y BP-4) son
ionizadas de forma más efectiva empleando ESI como fuente de ionización. Por
otro lado, Wick y col. [Wick, 2010] comparan ESI y APCI para la determinación
de benzofenonas. Aunque la eficacia de ionización en APCI se ve menos
afectada por otros componentes de la matriz, el mejor rendimiento es
proporcionado por ESI.
Otra limitación para la determinación simultánea de un número amplio
de compuestos mediante LC-MS es la mayor o menor tendencia de los
diferentes analitos a ionizarse en modo positivo o negativo. A modo de
ejemplo, la eficacia de ionización de la BP-3 en ESI es mucho mayor en modo
positivo que en negativo; por el contrario, sus metabolitos y la BP-4 presentan
una mayor tendencia a ionizarse en modo negativo [Rodil, 2008-A] [Nieto, 2009]
[Zenker, 2008] [Pedrouzo, 2009; Pedrouzo, 2010]. Dentro de esta tesis se ha
desarrollado un método para la determinación simultánea de BP-4, BP-3 y otras
benzofenonas hidroxiladas, relacionadas estructuralmente con la BP-3 (BP-1,
Introducción-Filtros solares
93
BP-2, BP-6 y BP-8), mediante LC-MS/MS empleando de forma simultánea los
modos de ionización positivo y negativo en ESI [Negreira, 2009-B].
A continuación se resumen en tablas las características de los métodos de LC-
MS y LC-MS/MS aplicados a la determinación de filtros solares en muestras
ambientales acuosas (Tabla 21) y sólidas (Tabla 22).
Introducción-Filtros solares
94
Tab
la 2
1: R
esum
en d
e la
s ap
licac
ione
s de
LC
-MS
y LC
-MS/
MS
a la
det
erm
inac
ión
de fi
ltro
s so
lare
s en
mue
stra
s ac
uosa
s.
Mat
riz
acu
osa
An
alit
os
Téc
nic
a
det
ecci
ón
Inte
rfas
e C
olu
mn
a em
ple
ada
Fase
móv
il
(mod
ific
ador
)
LO
D
(ng
L-1
) R
efer
enci
a
Río
, mar
y
resi
dua
l
BP
-4
LC
-
MS/
MS
ESI
(-)
Sym
met
rySc
hiel
d R
P-1
8
(150
x 2
,1 m
m, 3
,5 µ
m)
Agu
a:M
etan
ol
(5 m
M a
ceta
to a
món
ico)
7-46
1-30
[Rod
il, 2
008-
A]
[Rod
il, 2
009-
B]
BP
-3, 4
-MB
C,
IAM
C, E
HM
C,
EH
PA
BA
, OC
R
ESI
(+)
Res
idua
l B
P-3
LC
-
MS/
MS
ESI
(+)
Sine
rgy
Max
-RP
(250
x 4
,6 m
m)
Agu
a:M
etan
ol
(0,1
% á
cid
o fó
rmic
o)
1,0
[Tre
nhol
m, 2
008]
Supe
rfic
ial
y re
sid
ual
BP
-1, B
P-2
,
BP
-3, B
P-4
LC
-
MS/
MS
ESI
(-)
AC
QU
ITY
UP
LC
BE
H C
18
(100
x 1
mm
, 1,7
µm
)
Agu
a:M
etan
ol
(0,5
% H
Ac,
0,5
% N
H3)
0,
1-5
[Kas
przy
k-H
ord
ern,
200
8]
Lag
o y
resi
dua
l
BP-
3, B
P-4
, 4-M
BC
,
HM
S, IA
MC
, EH
MC
,
OC
R, E
HPA
BA
LC
-
MS/
MS
AP
PI
C8
Ecl
ipse
XD
B
(150
x 4
,6 m
m, 5
µm
) A
gua:
Met
anol
0,
4-16
[R
odil,
200
9-C
]
Río
y
resi
dua
l
BP
-1, B
P-8
BP
-3, E
HPA
BA
, OC
R
UP
LC
-
MS/
MS
ESI
(-)
ESI
(+)
C18
Ecl
ipse
XD
B
(50
x 4,
6 m
m, 1
,8 µ
m)
Agu
a:M
etan
ol
(áci
do
acét
ico,
pH
2,8
)
1-4
(río
)
1-10
(efl
u.)
5-20
(inf
lu.)
[Ped
rouz
o, 2
009]
2,5
(río
)
5-10
(res
.) [P
edro
uzo,
201
0]
Introducción-Filtros solares
95
Tab
la 2
2: R
esum
en d
e la
s ap
licac
ione
s de
LC
-MS
y LC
-MS/
MS
para
la d
eter
min
ació
n de
filt
ros
sola
res
en m
atri
ces
sólid
as.
Mat
riz
An
alit
os
Téc
nic
a
det
ecci
ón
Inte
rfas
e C
olu
mn
a em
ple
ada
Fase
móv
il
(mod
ific
ador
)
LO
D
(ng
g-1 )
R
efer
enci
a
Pes
cad
o B
P-3
, EH
MC
,
4-M
BC
, OC
R
LC
-
MS/
MS
ESI
(+)
Perf
ectS
il 12
0 O
DS-
2
(125
x 3
mm
)
Agu
a:M
etan
ol
(0,1
% á
cid
o ac
étic
o)
2,4
[Mei
nerl
ing,
200
6]
Pes
cad
o
BP
-3, 4
-MB
C, E
HM
C
BP-
1, B
P-2
, BP
-4
LC
-
MS/
MS
ESI
(+)
ESI
(-)
Zor
bax
SB-C
18
(150
x 3
mm
, 3,5
µm
)
Agu
a:A
ceto
nitr
ilo
(0,1
% á
cid
o fó
rmic
o)
78-2
05
[Zen
ker,
200
8]
6-50
[F
ent,
2010
-B]
Pes
cad
o B
P-3
L
C-M
S E
SI (+
) A
lltim
a
(250
x 2
,1 m
m, 5
µm
)
Agu
a:A
ceto
nitr
ilo
(0,1
% á
cid
o fó
rmic
o)
2 [K
won
, 200
9]
Lod
o B
P-3
, EH
S, H
MS,
IA
MC
, EH
MC
, OC
R,
4-M
BC
, EH
PA
BA
,
LC
-
MS/
MS
AP
PI
Ecl
ipse
XD
B C
8,
(150
x 4
,6 m
m, 5
µm
) A
gua:
Met
anol
0,
3-18
[R
odil,
200
9-D
]
Lod
o B
P-3
, OC
R, E
HP
AB
A
UP
LC
-
MS/
MS
ESI
(+)
Zor
bax
(50
× 4,
6 m
m, 1
,8 µ
m)
Agu
a:M
etan
ol
(áci
do
acét
ico)
1,
5-3,
5 [N
ieto
, 200
9]
BP
-1, B
P-8
E
SI (-
)
Lod
o
BP
-3
LC
-
MS/
MS
ESI
(+)
Syne
rgi F
usi
on-R
P 80
(150
× 3
mm
, 4 µ
m)
Agu
a (1
0 m
M fo
rmat
o
amón
ico)
:Ace
toni
trilo
(0,1
% á
cid
o fó
rmic
o)
0,75
-7,5
[W
ick,
201
0]
BP-
1, B
P-2
, BP
-4
ESI
(-)
Introducción-Fotoiniciadores
97
B. Fotoiniciadores
1. ASPECTOS GENERALES
El segundo grupo de compuestos considerados en este trabajo son los
llamados fotoiniciadores, compuestos químicos que se añaden a las tintas
aplicadas sobre envases alimentarios con el fin de acelerar el proceso de secado
de las impresiones realizadas sobre los mismos. En el año 2005, en Italia, se
produjo una alarma social al detectar restos de un fotoiniciador de esos
componentes químicos en leche infantil distribuida en envases tipo brick. Éste
fue la Isopropil tioxantona (ITX). Esta preocupación condujo a que las
autoridades retiraran del mercado más de dos millones de litros de leche en
Italia, España, Francia y Portugal.
En el caso anterior, parece que el problema de contaminación se originó
al almacenar los envases impresos enrollados en grandes bobinas. Durante este
proceso, la cara externa impresa entra en contacto con la cara interna del
tetrabrick; de esa manera, parte del ITX utilizado en el sistema de impresión
puede depositarse en el interior del envase, desde donde se produjo su
posterior migración a la leche.
El mecanismo de actuación de los fotoiniciadores se basa en su
descomposición en presencia de luz ultravioleta, lo que genera radicales libres
que activan la polimerización de los componentes de la tinta. Tradicionalmente,
las tintas incluían en su formulación disolventes orgánicos o agua, que tenían
que ser eliminados mediante un proceso de secado relativamente lento.
Actualmente, la mayoría de los procesos de secado están basados en la
exposición a radiación ultravioleta. Este proceso es iniciado por un
fotoiniciador, que es altamente propenso a absorber un fotón de luz y crear
especies activas (radicales, cationes o aniones) que inician y completan dicho
proceso de polimerización y secado de la tinta [Sanches-Silva, 2008-B] [Bradley,
2006].
Introducción-Fotoiniciadores
98
2. ESTRUCTURA Y PROPIEDADES
Generalmente, los fotoiniciadores presentan uno o dos anillos aromáticos
en sus moléculas con estructuras parecidas a los filtros UV empleados en
protectores solares. De hecho, el 2-etil-4-dimetilaminobenzoato (EDMAB) es un
derivado del PABA. Los analitos incluidos en este estudio han sido:
benzofenona (BP), 1-hidroxiciclohexil-fenilcetona (CPK), 4-metilbenzofenona (4-
MBP), 2-isopropiltioxantona (ITX), 2,2’-dimetoxi-2-fenilacetofenona (DMPA),
EDMAB y el EHPABA, usado también como filtro solar y de estructura muy
similar al EDMAB. A continuación, se muestran sus estructuras químicas
(Figura 14) y sus propiedades físico-químicas más relevantes (Tabla 23). Los
datos correspondientes al EHPABA han sido presentados en páginas anteriores
de esta memoria.
4-MBPCPK
ITX DMPA
BP
EDMAB
Figura 14: Estructuras de los fotoiniciadores objeto de estudio.
Introducción-Fotoiniciadores
99
Tabla 23: Propiedades físico-químicas de los fotoiniciadores estudiados.
Propiedad Analito
BP CPK 4-MBP ITX DMPA EDMAB
Nº CAS 119-61-9 947-19-3 134-84-9 5495-84-1 24650-42-8 10287-53-3
Peso molecular
(g mol-1) 182,20 204,26 196,24 254,35 256,30 193,24
Densidad
(g cm-3) 1,09 ± 0,06 1,14 ± 0,06 1,07 ± 0,06 1,20 ± 0,06 1,12 ± 0,06 1,06 ± 0,06
pKa - 13,2 ± 0,2 - - - 2,6 ± 0,1
Entalpía de
vaporización (KJ mol-1) 55 ± 3 61 ± 3 57 ± 3 65 ± 3 62 ± 3 54 ± 3
logKow 3,2 ± 0,3 2,3 ± 0,3 3,6 ± 0,3 5,3 ± 0,3 4,8 ± 0,5 3,1 ± 0,2
Presión de vapor
(mTorr) 0,823 0,0368 0,194 0,043 0,0106 1,43
Punto de ebullición
(ºC) 305,4 ± 0,0 339 ± 25 328 ± 11 399 ± 32 371 ± 42 297 ± 23
Punto de fusión
(ºC) 124 ± 14 144 ± 16 141 ± 14 217 ± 11 169 ± 14 115 ± 14
Solubilidad molar
(mol L-1), pH 1 7,1E-4 2,4E-3 3,4E-4 2,0E-6 2,4E-4 0,037
Solubilidad molar
(mol L-1), pH 4-10 7,1E-4 2,4E-3 3,4E-4 2,0E-6 2,4E-4 1,1E-3
3. PRESENCIA EN ALIMENTOS
Después de la alarma social provocada por la detección de ITX en leche,
aparecieron publicaciones en las que determinan este compuesto no sólo en
leche [Sagratini, 2008] [Gil-Vergara, 2007] [Bagnati, 2007] [Sun, 2007] [Allegrone,
2008] [Sanches-Silva, 2008-A] [Shen, 2009] sino también en otros alimentos como
yogures [Sun, 2007], zumo [Sagratini, 2008] e incluso té [Sun, 2007] y vino
[Sagratini, 2008]. El ITX aparece como una mezcla de dos isómeros pero el más
frecuentes es el 2-ITX encontrando niveles de hasta 346 ± 102 µg L-1 en leches
distribuidas en envases tipo brick [Bagnati, 2007].
Además del ITX, se encontraron trazas de otros fotoiniciadores como
EHPABA en leche [Gil-Vergara, 2007] [Sagratini, 2008] [Sanches-Silva, 2008-A] y
Introducción-Fotoiniciadores
100
zumo [Sagratini, 2008]; BP en leche [Sagratini, 2008] [Shen, 2009], zumo
[Sagratini, 2008] y vino [Sagratini, 2008] y CPK en vino [Sagratini, 2008]. En la
siguiente tabla (Tabla 24), se recogen las concentraciones encontradas para los
compuestos anteriores en alimentos envasados.
Tabla 24: Niveles de fotoiniciadores en muestras de leche, zumo, yogur, té y vino.
Tipo de
muestra
Lugar del
estudio
Compuestos
detectados
Concentración
(µg L-1) Referencia
Leche España ITX
EHPABA
2,5-189
8-101 [Gil-Vergara, 2007]
Leche Italia
2-ITX
4-ITX
249-346
17-9
[Bagnati, 2007]
Leche
China ITX
<0,5-4,78
[Sun, 2007] Zumo <0,5-84,30
Yogur <0,5
Té <0,5
Leche Italia ITX 4-53 [Allegrone, 2008]
Leche
Italia
EHPABA
BP
0,13-0,8
5,25-39
[Sagratini, 2008]
Zumo
EHPABA
BP
2-ITX
0,14-0,8
5-90
0,2
Vino
BP
2-ITX
CPK
5,5-217
0,2-0,24
1,2
Leche España EHPABA
ITX
78-214
37-213 [Sanches-Silva, 2008-A]
Leche China
2-ITX 0,81-8,87
[Shen, 2009] BP 2,84-18,35
EHPABA 0,32
Introducción-Fotoiniciadores
101
4. METODOLOGÍA ANALÍTICA PARA LA DETERMINACIÓN DE
FOTOINICIADORES EN ALIMENTOS
Dentro de este apartado se presentan las características básicas de los
métodos propuestos en la bibliografía para la determinación de fotoiniciadores
en muestras de alimentos envasados. La mayoría de los trabajos recopilados en
la bibliografía se centran en alimentos líquidos (leche, zumos, bebidas
alcoholicas), semisólidos (yogures, margarinas) y sólidos (fundamentalmente
leches infantiles en polvo). De forma genérica, se estima que el contenido
lípidico o etanólico (caso del vino) de los alimentos anteriores podría favorecer
la liberación de los fotoiniciadores que pudiesen estar adheridos a la cara
interna de los envases multicapa, debido a problemas de contaminación
durante su almacenamiento, o a una potencial migración desde la cara externa
impresa, hacia la interna en contacto con el alimento. El interés por la
determinación de fotoiniciadores en leche se debe por un lado a la frecuente
distribución de este alimento en envases tipo brick y a su consideración como
nutriente imprescindible en la dieta de bebés y niños.
La complejidad de las muestras de alimentos hace necesaria una etapa
previa para separar los analitos de otros componentes de la matriz y, en la
mayoría de los casos, un paso adicional de limpieza para mejorar la selectividad
de la extracción antes de la etapa de determinación. La técnica más utilizada
para la extracción de fotoiniciadores en leche [Morlock, 2006] [Bagnati, 2007]
[Sanches-Silva, 2008-A] y otras matrices, como yogur [Morlock, 2006] [Benetti,
2008] es la LLE. En la mayoría de los trabajos se emplea acetonitrilo como
disolvente de extracción, permitiendo la recuperación de los compuestos con un
bajo nivel de lípidos, posibilitando la inyección del extracto en un equipo de
HPLC [Bagnati, 2007] [Sanches-Silva, 2008-B] [Benetti, 2008]. Otros autores pasan
los extractos a través de cartuchos de sílica [Sagratini, 2008], C18 (1 g, 6 mL)
[Allegrone, 2008] o Oasis HLB (60 mg, 3 mL) [Sun, 2007] [Gallart-Ayala, 2008]
[Shen, 2009] antes de su inyección en el sistema cromatográfico. En algunos
trabajos, Morlock y col. [Morlock, 2006] y Gil-Vergara y col. [Gil-Vergara, 2007]
Introducción-Fotoiniciadores
102
usan PLE para la extracción de ITX y EHDAB de muestras de leche y yogur. El
disolvente de extracción fue ciclohexano:acetato de etilo (1:1) [Morlock, 2006] y
acetato de etilo [Gil-Vergara, 2007]. En ambos casos, la temperatura de
extracción fue de 100 ºC y la técnica de determinación fue LC-MS. La
determinación mediante GC-MS sólo se llevó a cabo en dos trabajos [Sagratini,
2008] [Allegrone, 2008] y en ambos casos emplean SPE para la extracción y
limpieza de los extractos.
A continuación se presenta una tabla resumen donde se recogen los
analitos considerados, las técnicas empleadas y los LOQ alcanzados en la
bibliografía en relación a la determinación de fotoionizasores en alimentos
(Tabla 25 y Tabla 25 cont.).
En la mayoría de las aplicaciones se consumen volúmenes considerables
de disolvente orgánico además de necesitar una etapa previa de limpieza antes
de su inyección en GC-MS [Allegrone, 2008]. En esta Tesis Doctoral se evaluó,
por primera vez, la SPME como técnica de extracción y concentración de siete
fotoiniciadores: BP, CPK, 4-MBP, EDMAB, EHPABA, 2,2-DMPA y ITX en
muestras de leche [Negreira, 2010-B].
Introducción-Fotoiniciadores
103
T
abla
25:
Res
umen
de
met
odol
ogía
s em
plea
das
para
la d
eter
min
ació
n de
foto
inic
iado
res
en a
limen
tos.
M
atri
z A
nal
itos
E
xtra
cció
n
Pu
rifi
caci
ón
T. D
et.
LO
D
(µg
L-1
)
Ref
eren
cia
Lec
he,
yogu
r IT
X
PL
E: 4
mL
o 4
g m
uest
ra,
4 g
hyd
rom
atri
x,
cicl
ohex
ano:
ace
tato
de
etilo
(1:1
), 10
0 ºC
- L
C-(
ESI
)-M
S 0,
13
[M
orlo
ck, 2
006]
Mar
gari
na,
soja
IT
X
1 g
de
mue
stra
,
1 m
L a
ceto
nitr
ilo, 4
0ºC
, 30
min
-
LC
-(E
SI)-
MS
1
[Mor
lock
, 200
6]
Lec
he,
yogu
r,
té, z
umo
2-IT
X
4-IT
X
10 g
de
mu
estr
a,
100
mL
ace
toni
trilo
: agu
a (1
:1),
1% r
eact
ivo
Car
rez,
agit
ació
n y
cent
rifu
gaci
ón
SPE
: Oas
is H
LB
(60
mg)
,
4 m
L a
ceto
nitr
ilo
LC
-(E
SI)-
MS/
MS
0,15
[Sun
, 200
7]
Lec
he
2-IT
X
4-IT
X
100
µL d
e m
uest
ra,
300
µL a
ceto
nitr
ilo,
agit
ació
n y
cent
rifu
gaci
ón
- L
C-(
ESI
)-
MS/
MS
0,75
[Bag
nati
, 200
7]
Lec
he
infa
ntil
ITX
, BP
,
CP
K, D
MP
A,
EH
DA
B
10 m
L d
e m
uest
ra,
1,5
mL
NH
3 y 2
x 2
0 m
L h
exan
o -
LC
-MS
20- 3
0
[San
ches
-Silv
a, 2
008-
A]
Lec
he
2-IT
X
4-IT
X
EH
DA
B
PL
E: 1
mL
de
mu
estr
a,
2 g
aren
a, 9
g N
a 2SO
4,
acet
ato
de
etilo
a 1
00ºC
- L
C-M
S/M
S 0,
1-0,
3
[G
il-V
erga
ra, 2
007]
-, no
dis
poni
ble
Introducción-Fotoiniciadores
104
Tab
la 2
5 co
nt.:
Res
umen
de
met
odol
ogía
s em
plea
das
para
la d
eter
min
ació
n de
foto
inic
iado
res
en a
limen
tos.
M
atri
z A
nal
itos
E
xtra
cció
n
Pu
rifi
caci
ón
T. D
et.
LO
D
(µg
L-1
) R
efer
enci
a
Zu
mo
de
nara
nja
ITX
, EH
PAB
A
DM
PA
, CP
K
10 m
L d
e m
uest
ra, c
once
ntra
n a
5 m
L
y ex
trae
n co
n 5
mL
ace
toni
trilo
-
LC
-UV
20
-30
[San
ches
-Silv
a, 2
008-
B]
Com
ida
infa
ntil,
zum
os, l
eche
2-IT
X
4-IT
X
2,5
g d
e m
ues
tra,
10 m
L a
ceto
nitr
ilo, r
eact
ivo
Car
rez,
agit
ar y
cen
trif
ugar
SPE
: Oas
is H
LB
(60
mg)
,
6 m
L a
ceto
nitr
ilo
LC
-(E
SI)-
MS/
MS
0,00
2-
0,01
3
[Gal
lart
-Aya
la, 2
008]
Lec
he, z
umos
,
vino
2-IT
X, E
HPA
BA
,
EH
DA
B, B
P, C
PK
EL
L: 2
5 m
L d
e m
ues
tra
3 x
30 m
L n
-hex
ano
SPE
: Síli
ca (1
g),
2 x
2 m
L h
exan
o:
acet
ato
de e
tilo
(30:
70)
LC
-(E
SI)-
MS/
MS
2-10
0
[Sag
rati
ni, 2
008]
GC
-MS
0,2-
1
Yog
ur
2-IT
X
4-IT
X
10 g
de
mue
stra
,
40 m
L a
ceto
nitr
ilo,
agit
ació
n y
cent
rifu
gaci
ón
-
LC
-(E
SI)-
MS
- [B
enet
ti, 2
008]
Lec
he
ITX
SPE
: 1 m
L d
e m
ues
tra
+ 9
mL
agu
a:
met
anol
(8:1
), ca
rtu
chos
Isol
ute
C18
(1 g
), 5
mL
ace
toni
trilo
- G
C-
MS/
MS
0,1
[Alle
gron
e, 2
008]
Lec
he
BP
, IT
X, C
PK,
EH
PA
BA
2 g
de
mue
stra
, 10
mL
ace
toni
trilo
,
soni
caci
ón y
cen
trif
uga
ción
SPE
: Oas
is H
LB
(60
mg)
,
6 m
L a
ceto
nitr
ilo
LC
-(E
SI)-
MS/
MS
0,1-
0,74
[S
hen,
200
9]
-, no
dis
pon
ible
Metodología desarrollada-Muestras acuosas
107
III. METODOLOGÍA DESARROLLADA
A. Filtros solares
1. MUESTRAS ACUOSAS
1.1. Introducción
La SPE es la técnica más utilizada para la concentración de filtros solares
en muestras acuosas; sin embargo, requiere el consumo de volúmenes elevados
de muestra y moderados de disolventes orgánicos. Las técnicas de
microextracción permiten subsanar los inconvenientes anteriores,
proporcionando una excelente sensibilidad. No obstante, algunas de la
modalidades empleadas, tales como SPME, SBSE, MEPS y ciertas versiones de
LPME, presentan cinéticas de extracción lentas, utilizan dispositivos frágiles
con un coste considerable y/o pueden requerir importantes adaptaciones en los
sistemas cromatográficos para transferir los analitos desde el dispositivo de
extracción a la columna cromatográfica.
Una alternativa a las técnicas de microextracción anteriormente citadas
es la denominada DLLME que permite la rápida extracción de especies
orgánicas de muestras acuosas. Aunque existen numerosas aplicaciones de
DLLME a diferentes especies orgánicas, incluso fue usada para la
determinación de BP-3 y otras 3 benzofenonas hidroxiladas [Tarazona, 2010], las
condiciones de extracción no han sido optimizadas para filtros UV menos
polares, detectados con frecuencia en el medio acuático. Así pues, en esta Tesis
se ha desarrollado un método de preparación de muestra rápido y sensible para
la determinación de 8 filtros solares y el BzS en matrices acuosas basado en
DLLME [Negreira, 2010-A].
La segunda técnica considerada en esta parte de la memoria ha sido la
microextracción en fase sólida sobre siliconas de grado técnico, en formato
disco. El bajo coste de este material permite desecharlo después de cada ciclo de
Metodología desarrollada-Muestras acuosas
108
extracción, evitando los problemas de contaminación cruzada entre muestras,
típicos de las técnicas de SPME y SBSE. Además de optimizar las condiciones
de extracción para un grupo amplio de filtros UV se ha comparado las eficacias
de extracción proporcionadas por siliconas, en diferentes formatos, y los
correspondientes a los Twister recubiertos con PDMS, empleados en SBSE.
La tercera aplicación presentada en esta sección de la memoria se ha
centrado en la combinación de SPME con GC-MS para la determinación de
varias benzofenonas (BP-1, BP-3, BP-8) y dos salicilatos (EHS y HMS), a niveles
de los ng L-1, en muestras de agua. Con objeto de mejorar los LOQs
previamente obtenidos en la bibliografía [Felix, 1998], en esta Tesis se ha
desarrollado un método de SPME combinado con derivatización on-fibre,
empleando reacciones de sililación para la determinación de EHS, HMS, BP-3 y
otras 2 benzofenonas hidroxiladas (BP-1 y BP-8) [Negreira, 2009-A].
Dentro de la familia de las benzofenonas, existen compuestos que no
pueden ser determinados mediante cromatografía de gases (ej. BP-4, presente a
niveles importantes en muestras de agua residual y poco eliminada en las
estaciones depuradoras convencionales) y otros cuya derivatización y
determinación mediante cromatografía de gases presenta muchas dificultades,
ej. BP-2 y BP-6. Por ello, se ha desarrollado un método de LC-MS/MS para la
determinación de varias benzofenonas, empleando SPE como técnica de
extracción. Las condiciones de trabajo han sido optimizadas con objeto de
compatibilizar la extracción y determinación simultánea de los analitos
considerados. [Negreira, 2009-B].
Adicionalmente a la optimización de métodos de preparación de
muestra, es importante conocer la estabilidad de los filtros UV en el medio
acuático. Dado que algunos presentan grupos fenólicos y aminos en sus
estructuras parece factible que puedan dar lugar a la formación de derivados
halogenados al reaccionar con el cloro empleado en la desinfección del agua. De
hecho, Sakkas y col. [Sakkas, 2003] detectaron la formación de derivados
clorados del EHPABA en agua de piscina, aunque no aportaron datos de la
velocidad de reacción ni de la estabilidad de los productos generados. En esta
Metodología desarrollada-Muestras acuosas
109
Tesis, se llevaron a cabo estudios de degradación química de EHPABA, EHS y
BP-3 en presencia de cloro libre y se identificaron varios productos de
transformación [Negreira, 2008]. Las muestras fueron concentradas mediante
SPE, empleando GC-MS para seguir la evolución temporal de los compuestos
de interés e identificar los productos de reacción formados.
Además del desarrollo de metodología analítica, en los trabajos que se
presentan a continuación se aporta un número relevante de datos relativos a la
presencia de diversos filtros solares en muestras de agua superficial, de piscina
y aguas residuales urbanas, discutiendo también su estabilidad en las
estaciones depuradoras de aguas residuales.
Metodología desarrollada-Muestras acuosas
110
1.2. Esquemas de los métodos desarrollados para muestras
acuosas
Separación de fases
Muestra agua10 mL
Adición 1 mL acetona (dispersante) conteniendo
60 µL clorobenceno (extractante)
Agitación durante 1 min
Inyección 2 µL GC-MS (modo SIM)
Centrifugación3 min, 3200 rpm
Formación de la emulsión
Recogida de la gota sedimentada
Figura 15: Esquema seguido para la determinación de filtros solares en agua mediante
DLLME y GC-MS.
Metodología desarrollada-Muestras acuosas
111
Muestra agua100 mL (10% metanol)
EXTRACCIÓN: inmersión,
14 horas a Tª ambiente
Inyección 9 µL PTV GC-MS
DESORCIÓN:30 min a Tª ambiente con
200 µL acetato de etilo
Retirada del disco de silicona y
secado con papel
Figura 16: Esquema empleado en la determinación de filtros solares en agua mediante
extracción con siliconas, en formato no comercial, y GC-MS.
Metodología desarrollada-Muestras acuosas
112
Derivatización on-fibre20 µL MSTFA10 min a 45 ºC
Muestra agua10 mL, pH 3
Inmersión, fibra PDMS-DVB, 30 min a Tª ambiente
con agitación
Retirada de la fibra, secar con papel
Extracción de los analitos
Derivatización
Desorción de la fibra2 min a 270 ºC GC-MS/MS
Figura 17: Esquema para la determinación de filtros solares en agua mediante SPME,
derivatización y GC-MS/MS.
Metodología desarrollada-Muestras acuosas
113
•Flujo: 10 mL/min.
•Secado cartuchos: corriente N2; 30 min.
•Ajustar pH 3.
•Acondicionamiento cartuchos.
5 mL metanol-acetato amónico (5 mM)
5 mL agua milli-Q
Inyección
LC—MS/MS (15 μL)
Oas
is H
LB
, 60
mg
Muestra agua
Elución 3 mL metanol-acetato amónico
(5 mM)
..
Evaporación a sequedad
Re-disolución con 1 mL metanol:agua milli-Q (1:1)2,5 mM acetato amónico
Figura 18: Esquema de trabajo para la determinación de filtros solares en agua mediante
SPE y LC-MS/MS.
Metodología desarrollada-Muestras acuosas
114
• Adicción de cloro libre (0 3 mg/L)
Medida de HClO/ClO- con fotómetro
• Adicción de analitos (aprox. 10 50 ng/mL)
Muestra agua tamponada (rango pH 6,2 8,2)
Reacción a Tª ambiente(0 180 min)
Parada de la reacción con tiosulfato sódico
Acidificar a pH 4,5 y concentrar mediante SPE (ver Figura 20)
Figura 19: Metodología empleada en los estudios de halogenación de filtros solares.
Metodología desarrollada-Muestras acuosas
115
Elución 2 mL
•Flujo: 10 mL/min.
•Secado cartuchos: corriente N2; 30 min.
•Elución con acetato de etilo o CH2Cl2.
•Ajustar pH 4,5.
•Acondicionamiento cartuchos.
3 mL acetato de etilo o CH2Cl2
3 mL metanol
3 mL agua Milli-Q
InyecciónGC-MS (2 μL)
Oas
is H
LB
, 60
mg
Muestra agua
..
PREPARACIÓN MUESTRA
DERIVATIZACIÓDERIVATIZACIÓDERIVATIZACIÓDERIVATIZACIÓDERIVATIZACIÓDERIVATIZACIÓDERIVATIZACIÓ
20 µL de MTBSTFA
(Tª ambiente, 5 min.)
DERIVATIZACIÓN (BP-3 y BP-1)
500 µLEXTRACTO
+Inyección
GC-MS (2 μL)
Figura 20: Condiciones de extracción y determinación empleadas en los estudios de
halogenación.
1.3. Publicación:
DISPERSIVE LIQUID-LIQUID
MICROEXTRACTION FOLLOWED BY GAS
CHROMATOGRAPHY-MASS SPECTROMETRY
FOR THE RAPID AND SENSITIVE DETERMINATION
OF UV FILTERS IN ENVIRONMENTAL
WATER SAMPLES
N. Negreira, I. Rodríguez, E. Rubí, R. Cela
Analytical and Bioanalytical Chemistry 398 (2010) 995
(doi:10.1007/s00216-010-4009-9)
Analytical and Bioanalytical Chemistry 398 (2010) 995
119
Dispersive liquid-liquid microextraction followed by gas chromatography-
mass spectrometry for the rapid and sensitive determination of UV filters in
environmental water samples
N. Negreira, I. Rodríguez*, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto
de Investigación y Análisis Alimentario (IIAA), Universidad de Santiago de
Compostela, Santiago de Compostela 15782, Spain.
Abstract
The performance of the dispersive liquid-liquid microextraction
(DLLME) technique for the determination of eight UV filters and a structurally
related personal care species, benzyl salicylate (BzS), in environmental water
samples is evaluated. After extraction, analytes were determined by gas
chromatography combined with mass spectrometry detection (GC-MS).
Parameters potentially affecting the performance of the sample preparation
method (sample pH, ionic strength, type and volume of dispersant and
extractant solvents) were systematically investigated using both, multi- and
univariant optimization strategies. Under final working conditions, analytes
were extracted from 10 mL water samples by addition of 1 mL of acetone
(dispersant) containing 60 L of chlorobenzene (extractant), without modifying
either the pH, or the ionic strength of the sample. Limits of quantification
(LOQs) between 2 and 14 ng L-1, inter-day variability (evaluated with relative
standard deviations, RSDs) from 9 to 14% and good linearity up to
concentrations of 10000 ng L-1 were obtained. Moreover, the efficiency of the
extraction was scarcely affected by the type of water sample. With the only
exception of 2-ethylhexyl-p-dimethylaminobenzoate (EHPABA), compounds
were found in environmental water samples at concentrations between 6 ± 1 ng
L-1 and 26 ± 2 ng mL-1.
Metodología desarrollada-Muestras acuosas
120
Keywords: UV filters; Dispersive liquid-liquid microextraction; GC-MS;
water samples.
1. Introduction
Reduction of the ozone layer increases the amount of UV radiation
reaching the surface of the earth, and thus the concern about its harmful effects
on human health. The so-called organic UV filters are compounds designed to
absorb radiation in the UV region, protecting human skin against direct
exposure to deleterious wavelengths of sunlight [1-2]. These compounds are
incorporated in sunscreen products, as well as in shampoos, lipsticks, facial
day-creams, after-shave products, make-up formulations and even in plastics,
varnishes and clothes to enhance their light stability [1-4]. These uses have led
to the appearance of significant concentrations of several UV filters in the
aquatic environment [5-6]. Particularly, compounds belonging to
benzophenone and salicylate classes, as well as 3-(4-methylbenzylidene)
camphor (4-MBC), 2-ethylhexyl-p-methoxycinnamate (EHMC) and octocrylene
(OCR) have been systematically detected in surface, bathing and municipal
wastewater samples, at concentrations ranging from the low ng L-1 up to several
ng mL-1 [7-12]. The latter three species and 2-hydroxy-4-methoxybenzophenone
(BP-3) have been also reported in river and lake sediments [13], sludge [14-15]
and aquatic organisms [3,16]. Moreover, the results of several studies suggest
that certain UV filters, which have been found in the above refereed matrices,
behave as endocrine disrupters [17-19]. Together, the above findings indicate
the need to evaluate the fate of the UV filters in the aquatic media in order to
assess their mobility, bio-accumulation potential and possible effects on aquatic
organisms.
The determination of UV filters in water samples is normally
accomplished with solid-phase extraction (SPE) followed by gas
chromatography (GC) [3,8-9] or liquid chromatography (LC) [7,10-11], both
combined with mass spectrometry (MS). Although SPE offers considerable
Analytical and Bioanalytical Chemistry 398 (2010) 995
121
advantages over liquid-liquid extraction (LLE), it still requires large sample
volumes (from 0.3 to 1 L), a moderate consumption (10-15 mL) of organic
solvents for analytes desorption, and further clean-up to compensate for its
limited selectivity when applied to wastewater [3,9]. Miniaturization of solid-
and liquid-phase extraction methodologies overcomes some of the above
limitations offering a similar performance to SPE. In this way, solid-phase
microextraction (SPME) [20-22], stir-bar sorptive extraction (SBSE) [23-24],
microextraction using packed sorbents (MEPS) [25] and liquid-phase
microextraction (LPME), with non-porous membranes [26] as well as in the
single drop modality [27-28], have been applied to the extraction of UV filters
from aqueous matrices. In most of the above applications an excellent
sensitivity has been achieved; however, relatively long extraction steps, fragile
extraction devices and/or dedicated equipment are required.
An alternative to the above microextraction techniques is the so-called
dispersive liquid-liquid microextraction (DLLME) [29]. This approach allows
the rapid extraction of organic species from aqueous solutions by addition of a
binary mixture of an extractant and a dispersant. The first is a high-density,
water insoluble solvent; whereas, acetone, methanol or acetonitrile are normally
used as dispersants. When this mixture comes in contact with the water sample
a cloudy stage, consisting of fine particles of the extractant dispersed into the
aqueous phase, is formed. After centrifugation, the high-density solvent settles
at the bottom of the extraction tube. Then, a fraction of the sedimented phase is
injected in the chromatographic system [29]. Following this initial study on
DLLME [29], more than one hundred applications have been published. Most of
them have been compiled in two recent reviews [30-31]. To the best of our
knowledge, DLLME has been only applied to the determination of BP-3 and
some hydroxylated by-products [32]; however, extraction conditions have not
been optimized for less polar UV filters, often found in the aquatic
environment.
Metodología desarrollada-Muestras acuosas
122
The aim of this study is to develop a rapid and sensitive sample
preparation method for the extraction of eight UV filters, commonly included in
the formulation of sunscreen products and belonging to different chemical
classes, as well as a structurally related compound (benzyl salicylate, BzS) from
environmental water samples. GC-MS was used as the determination
technique. Parameters affecting the efficiency of the extraction process have
been evaluated systematically and the performance of the developed method
compared with previously reported approaches involving other
microextraction techniques and with SPE. An overview of the concentrations
for selected UV filters in a significant number of water samples is also provided.
2. Experimental
2.1. Standards and material
HPLC-grade methanol and acetonitrile were supplied by Merck
(Darmstadt, Germany). Trace analysis grade carbon tetrachloride (CCl4), 1,1,1-
trichloroethane (CH3CCl3), chlorobenzene (C6H5Cl) and acetone were obtained
from Aldrich (Milwaukee, WI, USA). Sodium chloride was also provided by
Aldrich. Ultrapure water was obtained from a Milli-Q system (Millipore,
Billerica, MA, USA). Standards of 2-ethylhexyl salicylate (EHS), 3,3,5-
trimethylcyclohexyl salicylate (Homosalate, HMS), 2-ethylhexyl-p-
dimethylaminobenzoate (EHPABA), BzS, BP-3, 4-MBC, EHMC and OCR were
acquired from Aldrich (Milwaukee, WI, USA) and Merck. Isoamyl-p-
methoxycinnamate (IAMC) was kindly provided by Dr. R. Rodil (University of
La Coruña, Spain). Individual solutions (ca. 1000 g mL-1) and mixtures of the
above analytes were prepared in methanol. Further dilutions were also made in
methanol and used to prepare the aqueous standards employed during
optimization of extraction conditions. Another series of standards was made in
chlorobenzene.
Glass tubes (12 mL volume) with a conical bottom and a screw cap,
furnished with a polytetrafluoroethylene (PTFE)-lined septum, were acquired
from Afora (Barcelona, Spain).
Analytical and Bioanalytical Chemistry 398 (2010) 995
123
2.2. Samples
River, swimming-pool and wastewater samples, obtained from the inlet
and outlet streams of an urban sewage treatment plant (STP) equipped with
primary and activated sludge units, were processed throughout this study.
Samples were collected in amber glass bottles (250 mL), previously rinsed with
acetone and ultrapure water, and transported immediately to the laboratory.
Then, they were passed through glass fibre filters followed by cellulose acetate
ones (0.45 m pore size) and stored in the dark, at 4º C, for a maximum of 48
hours before being concentrated. In the case of swimming-pool water samples,
200 mg of sodium thiosulphate were added immediately after collection in
order to remove free chlorine. During sampling and sample preparation
operations gloves were used to minimize contamination problems arising from
the usage of UV filters and BzS in hand creams, bath gels and other personal
care products.
2.3. Sample preparation
Optimization of DLLME conditions was performed with aqueous
solutions of target species in ultrapure water. Unless otherwise is stated, the
concentration of UV filters in these aqueous standards was 5 ng mL-1. Aliquots
of 10 mL were poured into conical bottom glass tubes, previously wrapped
with aluminium foil to prevent photo-isomerization of UV filters during
extraction. The extraction mixture was added using a micropipette. Both, uni-
and multivariant optimization strategies, based on the use of experimental
factorial designs, were considered in order to assess the effect of different
experimental variables on the performance of the extraction process. In the
second case, the Statgraphics Centurion software (Manugistics, Rockville, MD,
USA) was used for data analysis.
Under optimized conditions, the binary extraction mixture consisted of 1
mL of acetone (dispersant) containing 60 L of chlorobenzene (extractant). After
addition to the water sample (or aqueous standard), the conical tube was
closed, shaken manually for 1 min and centrifuged at 3200 rpm (ca. 900 g) for 3
Metodología desarrollada-Muestras acuosas
124
min. During centrifugation, the dispersed droplets of chlorobenzene settled at
the bottom of the vessel, building-up a drop (45 L volume) which was further
recovered and transferred to a conical insert with a 25 µL micro-syringe fitted
with a bevel-tip needle. A conventional autosampler for liquid samples was
used to inject a fraction (2 L) of this extract in the GC-MS system.
2.4. GC-MS equipment
Analytes were determined with a GC-MS system consisting of an Agilent
(Wilmington, DE, USA) 7890A gas chromatograph connected to a quadrupole
mass spectrometer (Agilent MS 5975C), operated in the SIM mode. Separations
were carried out in an HP-5ms type capillary column (30 m x 0.25 mm i.d., df:
0.25 m) supplied by Agilent. Helium (99.999 %) was used as the carrier gas at a
constant flow of 1.2 mL min-1. The GC oven was programmed as follows: 125 ºC
(held for 1 min), increased at 9 ºC min-1 to 280 ºC (held for 10 min).
Temperatures of the electron impact ionization source and the quadrupole mass
analyzer were set at 230 ºC and 150 ºC, respectively. Standards and sample
extracts in chlorobenzene (2 L) were injected in the splitless mode (splitless
time 1 min), with the injector port at 280 ºC. Retention times and m/z ratios of
selected ions for each compound are summarized in Table 1.
Table 1. Retention times and monitored ions for target species.
Compound Retention time (min) aSelected ions (m/z)
EHS
HMS
BzS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OCR
9.77
10.46,10.71
10.54
b12.28
12.28
b12.57
14.57
b15.01
18.11
120,138
120,138
91
178,161
227,151
254,239
165,277
178,161
360, 232, 249
aUnderlined ions were used for quantification purposes.
bRetention times corresponding to (E) forms.
Analytical and Bioanalytical Chemistry 398 (2010) 995
125
2.5. Method characterization and samples quantification
The performance of the optimized methodology was characterized in
terms of extraction efficiencies (%) and enrichment factors (EFs). EFs were
defined as the ratio between the concentration of each species in the
chlorobenzene extract (Cs), which was determined against a calibration curve
obtained for standards in chlorobenzene, and that added to the water sample
(C0). Efficiencies (%) were calculated as the ratio between the mass of each
species in the settled phase (Cs x Vs) and that added to the sample (C0 x Va),
multiplied by 100 [29-31]. Being Vs and Va the volumes of the settled
chlorobenzene extract and the water sample, respectively.
Potential changes in the performance of the DLLME process among
ultrapure and environmental water samples (matrix effects) were assessed
using relative recoveries (%). They were calculated as the difference between
responses measured for spiked and non-spiked aliquots of different
environmental water samples divided by those obtained for ultrapure water
standards with the same addition level and multiplied by 100 [26]. Obtained
values stayed around 100%; therefore, the levels of UV filters in real-life
samples were quantified by comparison with the responses obtained for
aqueous standards in ultrapure water, containing increased amounts of target
analytes (up to 10000 ng L-1) and subjected to the optimized DLLME process.
Procedural blanks (ultrapure water) were tested with each series of
samples in order to assess the significance of contamination problems.
Responses measured for real samples were corrected with those obtained for
procedural blanks.
3. Results and discussion
3.1. Optimization of DLLME conditions
3.1.1. Effects of extractant, sample pH, salt addition and extraction time
The type of extractant is one of the most important parameters during
optimization of DLLME methods. In a first series of assays (n=3 replicates),
mixtures of 0.9 mL of acetone with 0.1 mL of 3 different extractants (CCl4,
Metodología desarrollada-Muestras acuosas
126
CH3CCl3 and C6H5Cl) were added to 10 mL aliquots of aqueous standards
prepared in ultrapure water. The lowest responses for all compounds
corresponded to CCl4 (data not shown) which was therefore rejected as
extractant.
The effects of sample pH, salt (NaCl) addition, extraction time (defined
as the period during which the sample is shaken after addition of the binary
extraction mixture and before centrifugation) and type of extractant on the
performance of the method were simultaneously investigated using a two
levels 24-1 type experimental fractional factorial design, with four replicates of
the central point. Low and high levels for each variable are shown in Table 2. In
all experiments, a binary mixture of acetone (0.9 mL) with 0.1 mL of the
corresponding extractant was used. The experimental conditions explored
within the domain of this design exerted a limited effect on the volume of the
sedimented phase (85 ± 4 L); thus, the peak areas for each compound were
directly considered as variables response.
Analytical and Bioanalytical Chemistry 398 (2010) 995
127
Tab
le 2
. Exp
erim
enta
l dom
ain
and
stan
dard
ized
mai
n ef
fect
s of
fact
ors
cons
ider
ed in
the
24-1
exp
erim
enta
l fra
ctio
nal f
acto
rial
des
ign.
Fact
or
Lev
el
Stan
dar
diz
ed m
ain
effe
ct v
alue
s
L
ow
Hig
h E
HS
HM
S B
zS
IAM
C
BP-
3 4-
MB
C
EH
PA
BA
E
HM
C
OC
R
NaC
l (%
) 0
5 -0
.58
-0.8
3 -0
.53
-1.1
0.
36
-0.0
10
-0.6
0 -0
.64
-1.1
Sam
ple
pH
2 6
0.17
0.
54
0.29
0.
64
1.1
0.78
0.
001
0.14
-0
.07
Ext
ract
ant
CH
3CC
l 3
C6H
5Cl
7.0
a 8.
5 a
8.9
a 15
a
17 a
11
a
10 a
6.
6 a
5.3
a
Ext
ract
ion
tim
e (m
in)
1 10
-0
.32
-0.2
0 0.
054
-0.0
67
0.18
0.
54
-0.2
9 -0
.09
-0.1
0
a St
atis
tica
lly s
igni
fica
nt fa
ctor
s at
the
95%
con
fid
ence
leve
l.
Metodología desarrollada-Muestras acuosas
128
Standardized main effects summarized in Table 2 indicate the same trend
for all species. The type of extractant was the only variable with a significant
influence (95% confidence level) on the performance of the extraction, with the
highest responses corresponding to C6H5Cl. On the other hand, the addition of
a 5% of NaCl to water samples and the extraction time played negligible effects
on the obtained responses; thus, they were maintained at their lower levels (0%
of NaCl and 1 min of shaking, respectively), to simplify sample handling and
also to speed up the extraction process. Finally, between pH 2 and 6 the
efficiency of the extraction also remained unaffected. For this latter factor,
considering that (1) environmental water samples can present pH values above
6 units and that (2) salicylates and BP-3 show a slightly acidic behaviour (their
pKa are comprised between 7.5 and 8 units), an additional series of extractions
was carried out with aqueous standards adjusted to pHs 6 and 9. Again, no
significant differences were observed in the responses of target analytes (Fig. 1);
therefore, the decision was to process the real water samples without any pH
adjustment. Data shown in Fig. 1 demonstrated that the ionized forms of EHS,
HMS and BP-3 are also effectively extracted by chlorobenzene. Likely, the
capability of this solvent to establish - interactions with the aromatic region
of target compounds significantly contributes to enhance the efficiency of the
extraction process. Two-factor interactions, obtained from the experimental
factorial study, also showed very low standardized values, far below the
statistically significant threshold (data not shown).
Analytical and Bioanalytical Chemistry 398 (2010) 995
129
0.0E+00
1.0E+05
2.0E+05
3.0E+05
4.0E+05
5.0E+05
6.0E+05
7.0E+05
8.0E+05
9.0E+05
EHS HMS BzS IAMC BP‐3 4‐MBC EHPABA EHMC OCR
Compound
Peak area
pH 6 pH 9
Fig. 1. Responses obtained for ultrapure water samples (spiked concentration 5 ng mL-1)
adjusted at two different pHs, n=3 replicates.
3.1.2. Selection of dispersant
Fig. 2 shows the responses obtained for spiked aliquots of ultrapure
water using the above optimized conditions and considering methanol or
acetonitrile as alternative to acetone. Overall, the lowest responses
corresponded to methanol, the most polar of the 3 tested dispersants, whereas
no differences were noticed between acetone and acetonitrile. Taking into
account its lower cost and larger commercial availability, acetone was
maintained as dispersant.
Metodología desarrollada-Muestras acuosas
130
0.0E+00
2.0E+05
4.0E+05
6.0E+05
8.0E+05
1.0E+06
EHS HMS BzS IAMC BP-3 4-MBC EHPABA EHMC OCR
Compound
Pe
ak
are
a
Methanol Acetone Acetonitrile
Fig. 2. Comparison of responses (peak areas) obtained with 0.1 mL of chlorobenzene as
extractant and 0.9 mL of three different dispersants: methanol, acetone and acetonitrile,
n= 3 replicates.
3.1.3. Volumes of extractant and dispersant
The optimal volumes of dispersant (acetone) and extractant
(chlorobenzene) were evaluated with a multilevel experimental factorial design.
The investigated levels were 0.5, 1, 1.5 and 2 mL for acetone and 0.06 and 0.1
mL of chlorobenzene. Each experiment was performed in duplicate. Analysis of
variance (ANOVA) was used in order to determine the contribution of both
factors, and their first order interaction, to responses (peak areas) measured for
target species. P-values for extractant and dispersant volumes remained under
0.05 for all the analytes, indicating a significant influence, at the 95% confidence
level, on their responses. Furthermore, the interaction acetone-chlorobenzene
was also statistically significant for four (IAMC, BP-3, EHPABA and OCR) of
the nine considered species. Numerical results of ANOVA are provided as
supplementary data, Table S1. Fig. 3 shows the interaction plots corresponding
to the average responses of selected compounds. As observed, the volume of
extractant was the factor affecting in a higher extension to the obtained peak
areas, with the most favourable situation corresponding to the lower volume of
Analytical and Bioanalytical Chemistry 398 (2010) 995
131
C6H5Cl. Considering 0.1 mL of C6H5Cl, the analytes peak areas decreased
steadily with the volume of acetone; however, for 0.06 mL of C6H5Cl the
maximum responses were obtained using 1 mL of acetone, Fig. 3.
Consequently, 1 mL of acetone and 60 L of chlorobenzene were fixed as
optimal values of both variables. Under these conditions, the volume of the
settled phase (45 L) was large enough to be transferred to a conical insert and
loaded in the autosampler of the GC-MS instrument.
Table S1. F-ratios and P-values resulting from ANOVA of data obtained in the
multilevel experimental factorial design of extractant and dispersant volumes.
Compound Main effects Two-factor interaction
Acetone volume Chlorobenzene volume Acetone-chlorobenzene
F-Ratio P-Value F-Ratio P-Value F-Ratio P-Value
EHS
HMS
BzS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OCR
9.5
9.7
10.4
13.2
18.1
10.20
10.8
11.07
23.7
0.0052
0.0049
0.0039
0.0018
0.0006
0.0041
0.0034
0.0032
0.0002
299
381
272
427
304
279
370
318
466
0.0000
0.0000
0.0000
0.0000
0.0000
0.0000
0.0000
0.0000
0.0000
3.2
3.4
4.0
5.1
4.4
4.0
4.4
2.9
6.8
0.086
0.072
0.052
0.029
0.041
0.051
0.042
0.10
0.014
Metodología desarrollada-Muestras acuosas
132
Acetone (mL)
47
67
87
107
127
Pea
k ar
ea (
x 10
4 )
0.5 1 1.5 2
EHS BP-3
18
28
38
48
58
0.5 1 1.5 2
Pea
k ar
ea (
x 10
4 )
Acetone (mL)
4-MBC
15
18
21
24
27
30
33
0.5 1 1.5 2
Acetone (mL)
Pea
k ar
ea (
x 10
4 )
EHPABA
65
85
105
125
145
0.5 1 1.5 2
Acetone (mL)
Pea
k ar
ea (
x 10
4 )
OCR
37
47
57
67
77
87
97
0.5 1 1.5 2
Acetone (mL)
Pea
k ar
ea (
x 10
4 )
EHMC
62
82
102
122
142
0.5 1 1.5 2
Acetone (mL)
Pea
k ar
ea (
x 10
4)
60 l of C6H5Cl
100 l of C6H5Cl
Fig. 3. Interaction plots obtained from the multilevel factorial experimental design for
selected compounds.
3.2. Performance of the method
Table 3 shows the extraction efficiencies and the EFs obtained for
ultrapure water standards under optimized conditions. Efficiencies varied from
74 to 95% and the averaged EFs stayed between 170 and 200 times. Likely, a
further reduction in the volume of extractant would provide even higher EFs.
Consequently, lower LOQs may be achieved. Nevertheless, this possibility was
Analytical and Bioanalytical Chemistry 398 (2010) 995
133
not investigated since extracts (settled drops) with a significant lower volume
have to be manually injected.
Table 3. Extraction efficiencies (%) and average enrichment factors (EFs) provided by
the DLLME method for ultrapure water standards, n=4 replicates.
Compound Extraction efficiencies (%) ± SD
Average EFs a1000 ng L-1 a200 ng L-1
EHS
HMS
BzS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OCR
90 ± 4
84 ± 4
79 ± 7
84 ± 4
81 ± 7
83 ± 4
92 ± 5
83 ± 5
85 ± 6
90 ± 4
85 ± 4
74 ± 2
95 ± 5
82 ± 5
81 ± 2
83 ± 6
84 ± 9
79 ± 6
200 ± 9
187 ± 9
170 ± 10
199 ± 10
181 ± 13
182 ± 7
194 ± 12
186 ± 16
182 ± 13
aConcentration level
The precision of the method was evaluated with aqueous standards of
different concentrations and extracted within the same day (repeatability) or on
different days (reproducibility). In the first case, relative standard deviations
(RSDs) from 2 to 11% were observed for triplicate extractions of aqueous
solutions fortified at four different concentrations: 50, 100, 1000 and 5000 ng L-1,
Table 4. Reproducibility was assessed with an aqueous standard (500 ng L-1)
processed in triplicate during 5 consecutive days. In this case, RSDs between 9
and 14% were attained. The linearity of the overall method was investigated
with standards prepared at eight different concentrations from 10 to 10000 ng L-
1. Fig. 4 shows the GC-MS chromatogram corresponding to a 50 ng L-1 aqueous
calibration solution. Within the above range, the plot of peak areas versus the
concentration of each analyte fitted a linear model with R2 values higher than
0.998, Table 4.
Metodología desarrollada-Muestras acuosas
134
Table 4. Repeatability (n=3 extractions in the same day), reproducibility (n=15
extractions in 5 different days), linearity and limits of quantification (LOQs) of the
method.
Compound
Repeatability,
(RSDs, %)
Reproducibility
(RSDs, %)
Linearity, R2
a (10- 10000,
8 levels)
LOQs
(ng L-1) a50 a100 a1000 a5000 a500
EHS
HMS
BzS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OCR
5
3
11
10
11
9
9
7
3
2
4
2
3
4
2
5
3
4
2
4
2
3
4
2
5
3
4
3
3
2
2
3
2
3
3
4
11
12
10
10
9
12
10
10
14
0.9995
0.9996
0.9997
0.9996
0.9993
0.9993
0.9994
0.9991
0.9982
2
9
3
4
7
3
3
14
10
aConcentration level (ng L-1)
EHS
HMS
(E) 4-MBC
EHPABABzS
OCR
(E) EHMC
BP-3, (E) IAMC
HMS
10.00 11.00 12.00 13.00 14.00 15.00 16.00 17.000
200
400
600
800
1000
1200
1400
1600
1800
2000
2200
2400
Time (min)
Abundance (counts)
Fig. 4. Total ionic current (TIC) chromatogram corresponding to a 50 ng L-1 aqueous
standard.
Analytical and Bioanalytical Chemistry 398 (2010) 995
135
Procedural blanks, performed with samples of ultrapure water, showed
the absence of significant contamination problems for most species. However,
traces of EHMC and OCR were systematically detected in blanks extracts, even
when different ultrapure water samples were tested. The exact source of this
contamination could not be identified; however, our finding is concordant with
the previous observations of Rodil et al. using SBSE as concentration technique
[23]. For those analytes not affected by contamination processes, the LOQs were
estimated as the concentration providing a peak with a signal to noise ratio
(S/N) 10 times higher than the baseline noise. In the case of EHMC and OCR,
LOQs were calculated as 10 times the standard deviations of blanks signals
(n=5 replicates) divided by the slope of their calibration curves. Achieved LOQs
remained below 14 ng L-1 for all species, Table 4. Globally, these values are
similar to those reported for SBSE followed by GC-MS determination (LOQs
from 2 to 26 ng L-1) [23-24], and liquid-phase microextraction (LPME) through
non-porous membranes in combination with LC-MS/MS detection (LOQs from
3 to 45 ng L-1) [25]. The combination of MEPS with GC-MS offers slightly higher
limits of detection (from 35 to 87 ng L-1) [26] and the same comment is also valid
for some SPME based methods [20]. SPE followed by GC-MS detection [9] and
LC-MS/MS [7] renders LOQs in the same range of values than those
summarized in Table 4.
Fig. 5 shows the relative recoveries of the optimized DLLME method for
river water, swimming-pool water (Fig. 5A) and wastewater samples (Fig. 5B)
fortified at different concentration levels. In the case of swimming-pool and
river water relative recoveries between 87% and 109% were achieved (Fig. 5A).
Similar results were attained for treated wastewater; whereas, values from 80%
to 117% were obtained for raw urban wastewater (Fig. 5B). In all cases, the
relative standard deviations remained around or below 10%. In view of these
data, it is evident that the efficiency of the sample preparation method is
scarcely affected by the characteristics of the matrix; therefore, external
calibration, against fortified aliquots of ultrapure water, was selected as
quantification technique.
Metodología desarrollada-Muestras acuosas
136
0
20
40
60
80
100
120
EHS HMS BzS IAMC BP-3 4-MBC EHPABA EHMC OCR
Compound
Rel
ativ
e re
cove
ry (
%)
Treated wastewater, 500 ng L-1 Treated wastewater, 1000 ng L-1
Raw wastewater, 2000 ng L-1
0
20
40
60
80
100
120
EHS HMS BzS IAMC BP-3 4-MBC EHPABA EHMC OCR
Compound
Rel
ativ
e re
cove
ry (
%)
Swimming pool, 250 ng L-1 River water, 100 ng L-1
River water, 500 ng L-1 River water, 1000 ng L-1
A
B
Fig. 5. Relative recoveries (%), with their standard deviations, obtained for river,
swimming-pool water (A) and wastewater samples (B), spiked at different concentration
levels, n= 3 replicates.
Analytical and Bioanalytical Chemistry 398 (2010) 995
137
3.3. Application to environmental water samples
The proposed method was applied to the analysis of 21 samples
corresponding to wastewater (codes 1-6), rivers without (codes 7-12) and with
bathing areas (codes 13-17) and public swimming-pools (codes 18-21), Table 5.
Most samples were collected in the summer of 2009 and all were processed
within 48 hours, after received. Except EHPABA, which remained below the
LOQ of the method in all samples, the rest of species were quantified in some of
the processed samples, with the highest detection frequency and the maximum
concentrations corresponding to OCR, followed by 4-MBC and EHMC. As
shown in Fig. 6, 4-MBC and EHMC were found in environmental samples as a
mixture of isomers (E and Z forms). The sum of their peak areas was compared
with the calibration curve obtained for the E form, which is the commercial
isomer included in the formulation of personal care products.
Metodología desarrollada-Muestras acuosas
138
Tab
le 5
. Con
cent
rati
ons
(ng
L-1 )
, wit
h th
eir
stan
dard
dev
iati
ons,
mea
sure
d in
env
iron
men
tal w
ater
sam
ples
, n=
3
repl
icat
es.
Cod
e T
ype
Sam
plin
g d
ate
Con
cent
rati
on (n
g L
-1) ±
SD
EH
S H
MS
BzS
IA
MC
B
P-3
4-
MB
C
EH
MC
O
CR
1
R.W
. 4/
05/0
9 6
± 1
N.D
. 28
± 1
N
.D.
17 ±
2
7 ±
1 N
.D.
66 ±
6
2 T
. W
4/05
/09
N.D
. N
.D.
N.D
. N
.D.
N.D
. N
.D.
N.D
. 15
± 1
3
R.W
. 9/
07/0
9 18
± 3
11
± 2
18
9 ±
21
N.D
. 11
1 ±
8 10
3 ±
9 51
± 6
15
8 ±
9 4
T. W
9/
07/0
9 N
.D.
N.D
. 8
± 1
N.D
. 10
± 1
66
±3
N.D
. 40
± 5
5
R.W
. 27
/7/0
9 32
± 1
20
± 1
13
7 ±
2 N
.D.
227
± 10
15
3 ±
3 12
4 ±
2 44
0 ±
34
6 T
. W
27/7
/09
N.D
. N
.D.
N.D
. N
.D.
58 ±
3
94 ±
5
N.D
. 59
± 4
0 7
Riv
er
9/07
/09
N.D
. N
.D.
6 ±
1 N
.D.
N.D
. N
.D.
N.D
. N
.D.
8 R
iver
9/
07/0
9 N
.D.
N.D
. 7
± 1
N.D
. 8
± 1
7 ±
1 N
.D.
18 ±
1
9 R
iver
9/
07/0
9 N
.D.
N.D
. 79
± 2
N
.D.
42 ±
3
7 ±
1 N
.D.
46 ±
4
10
Riv
er
9/07
/09
60 ±
5
N.D
. 31
± 4
N
.D.
N.D
. N
.D.
N.D
. N
.D.
11
Riv
er
26/0
7/09
12
± 1
N
.D.
12 ±
1
N.D
. 12
± 2
13
4 ±
4 81
3 ±
7 80
2 ±
33
12
Riv
er
26/0
7/09
N
.D.
N.D
. N
.D.
N.D
. N
.D.
N.D
. N
.D.
79 ±
19
13
Riv
er
11/0
7/09
N
.D.
N.D
. N
.D.
N.D
. N
.D.
N.D
. 10
4 ±
16
209
± 43
14
R
iver
11
/07/
09
N.D
. N
.D.
N.D
. N
.D.
N.D
. N
.D.
20 ±
1
48 ±
6
15
Riv
er
11/0
7/09
N
.D.
N.D
. N
.D.
N.D
. N
.D.
N.D
. 19
± 2
36
± 3
16
R
iver
25
/07/
09
31 ±
5
18 ±
3
N.D
. N
.D.
N.D
. 73
± 1
2 12
7± 3
42
56 ±
311
17
R
iver
25
/07/
09
62 ±
8
124
± 6
N.D
. N
.D.
15 ±
2
1132
± 1
43
362
± 41
33
90 ±
459
18
S.
P.
11/0
7/09
19
± 1
N
.D.
N.D
. 65
5 ±
48
N.D
. 10
70 ±
93
1274
± 1
40
1421
± 2
61
19
S.P.
25
/07/
09
17 ±
1
13 ±
1
14 ±
1
537
± 8
N.D
. 12
50 ±
16
1462
± 5
3 27
52 ±
167
20
S.
P.
25/0
7/09
17
8 ±
7 45
0 ±
20
N.D
. N
.D.
2326
± 1
18
4035
± 9
2 20
7 ±
7 25
967
± 16
22
21
S.P.
25
/07/
09
12 ±
1
N.D
. N
.D.
33 ±
2
N.D
. 43
± 3
10
7 ±
12
2997
± 3
26
Det
ecti
on fr
eque
ncy
52%
29
%
48%
14
%
43%
67
%
62%
90
%
R
.W. r
aw w
aste
wat
er
T.W
. tre
ated
was
tew
ater
S.
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wim
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ater
N
.D. b
elow
det
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s
Analytical and Bioanalytical Chemistry 398 (2010) 995
139
Code 17Code 11Procedural blank
9.60 9.80 10.00 10.20 10.40 10.60 10.80 11.00
406080
100120140160180200220240260280
Time (min)
Abundance m/z 120
EHS HMS
Time (min)11.50 11.70 11.90 12.10 12.30
405060708090
100110120130140150160170180
Abundancem/z 227
BP-3
13.40 13.80 14.20 14.60 15.00 15.400
500100015002000250030003500400045005000550060006500
Abundance
Time (min)
(Z) EHMC
(E) EHMC
m/z 178
11.80 12.20 12.60 13.00 13.400
200
400
600
800
1000
1200
1400
1600
1800
2000
2200
Abundance
Time (min)
(Z) 4-MBC
(E) 4-MBC
m/z 254
18.6017.60 17.80 18.00 18.20 18.400
400
800
1200
1600
2000
2400
2800
3200
3600
4000
4400
Abundance
Time (min)
m/z 360
OCR
Fig. 6. Selected ion chromatograms for a procedural blank and samples code 11 and 17
(Table 5).
Samples 1 to 6 were collected in the same STP, from a non-costal city of
100,000 inhabitants, equipped with primary and activated sludge treatments.
Although data summarized in Table 5 correspond to grab samples, obtained
Metodología desarrollada-Muestras acuosas
140
without considering the residence time of the STP, all species appear to be
removed to a considerable extent in the plant, with 4-MBC displaying the
lowest degradation efficiency. This trend is consistent with the results reported
for time-average samples from several STPs in Switzerland [3].
Analytes were also found in rivers without bathing or recreational areas
(codes 7-12). In some cases, these rivers received treated municipal wastewaters
(codes 8-9); however, the highest concentrations of 4-MBC, EHMC and OCR
were measured in a small stream (code 11) flowing through a highly
industrialized area and a 400,000 inhabitants city in the Northwest of Spain.
Some of the river water samples, collected in bathing areas (codes 13-17),
and particularly all swimming-pool water samples (codes 18-21), presented
very high concentrations of several UV filters (4-MBC, EHMC and OCR),
whereas the salicylates remained at lower levels. In fact, the concentration of
OCR in sample code 20 surpassed the linear response range of the method.
Thus, this sample was processed twice. First directly to quantify those species
presented at low levels, and then after a 5-fold dilution to establish the
concentration of OCR. It is worth noting that IAMC was detected only in
swimming-pool water but not in sewage or in river water. This finding suggest
a limited usage of this UV filter in sunscreen products in comparison with
EHMC, the other cinnamate type UV filter involved in this study.
4. Conclusions
DLLME constitutes an advantageous, rapid and inexpensive alternative
for the sensitive extraction of selected UV filters from environmental water
samples. The proposed method requires a small volume of sample, extraction is
completed in a few minutes (using a four positions centrifuge sample
preparation remains around 2-3 min per sample), dedicated instrumentation
and sorbents are not required and the extractant is compatible with GC-based
determination techniques. Moreover, the extraction efficiency is barely affected
by the type of water sample, allowing the comparison against spiked aliquots of
ultrapure water (aqueous standards) as quantification technique. The volume
Analytical and Bioanalytical Chemistry 398 (2010) 995
141
and the type of extractant and dispersant solvents were the variables with the
highest influence on the efficiency of the extraction, whereas the ionic strength
and the pH of the sample did not affect the extraction process. Data obtained
for different environmental water samples revealed the presence of eight of the
nine investigated species in the aquatic environment, with the highest
occurrence frequency corresponding to OCR, followed by 4-MBC and EHMC.
Acknowledgements
Financial support from the Spanish Government and E.U. FEDER funds
(project CTQ2009-08377) is acknowledged. N.N. is grateful for an FPU grant
from the Spanish Ministry of Education and Science.
Metodología desarrollada-Muestras acuosas
142
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1.4. Publicación:
SILICONE DISCS AS DISPOSABLE
ENRICHMENT PROBES FOR GAS
CHROMATOGRAPHY-MASS SPECTROMETRY
DETERMINATION OF UV FILTERS
IN WATER SAMPLES
N. Negreira, I. Rodríguez, E. Rubí, R. Cela
Analytical and Bioanalytical Chemistry, submitted
Analytical and Bioanalytical Chemistry, submitted
147
Silicone discs as disposable enrichment probes for gas chromatography-mass
spectrometry determination of UV filters in water samples
N. Negreira, I. Rodríguez*, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto
de Investigación y Análisis Alimentario, Universidad de Santiago de
Compostela, Santiago de Compostela 15782, Spain.
Abstract
This work describes an effective, low solvent consumption and
affordable sample preparation approach for the determination of eight UV
filters in surface and wastewater samples. It involves sorptive extraction of
target analytes in a disposable, technical grade silicone disc (5 mm diameter x
0.6 mm thickness) followed by organic solvent desorption, large volume
injection (LVI) and gas chromatography-mass spectrometry determination (GC-
MS). Parameters affecting the performance of sampling and desorption steps
are systematically investigated and the observed trends are related with the
polarity of UV filters. Final working conditions involved overnight extraction of
100 mL samples, containing a 10% of methanol, followed by analytes
desorption with 0.2 mL of ethyl acetate. The method provides linear responses
between the limits of quantification (from 0.003 to 0.040 ng mL-1) and 10 ng mL-
1, an intra-day precision below 13%, and low matrix effects for surface,
swimming pool and treated sewage water samples. Except in the case of
benzophenone-3 (BP-3), extraction efficiencies are in reasonable agreement with
values predicted from octanol-water partition coefficients of UV filters, sample
and sorbent volumes. Globally, no differences are noticed between the
extraction yields provided by the bulk, technical grade silicone sorbent and
those corresponding to polydimethylsiloxane covered stirring bars. Several UV
filters were found in surface and sewage water samples, with the maximum
concentrations corresponding to octocrylene (OC).
Metodología desarrollada-Muestras acuosas
148
Keywords: sample preparation, sorptive extraction, silicone sorbents, UV
filters, water analysis
1. Introduction
Organic UV filters belong to the group of personal care chemicals
considered as emerging environmental pollutants [1]. Direct release in bathing
areas and diffuse discharges through domestic wastewater represent the main
inputs of these species in the aquatic media [2-3]. The available information
related with removal efficiencies of UV filters in sewage treatment plants (STPs)
[4], occurrence in surface water samples, [5-6], accumulation in solid matrices
and biota [7-11] and eco-toxicological effects [12] indicate the need of long term
monitoring studies in order (1) to detect seasonal variations in surface and
sewage water, (2) to fully understand their removal and/or accumulation
routes and (3) to assess the impact of UV filters in the environment. The
feasibility of such long term monitoring studies, involving the analysis of many
samples, depends on the availability of cost affordable analytical
methodologies.
Trace level determination of UV filters in water samples is normally
accomplished by gas (GC) or liquid chromatography (LC) followed by mass
spectrometry (MS), after an effective sample preparation step. Sample
preparation contributes significantly not only to the performance of the overall
analytical method but also to the cost, complexity and automation possibilities
of the analysis. The above comments justify the plethora of extraction and
microextraction approaches proposed for the concentration of UV filters in
water samples. Among solid-phase microextraction modalities, stir bar sorptive
extraction (SBSE) combines some interesting features, such as high extraction
yields for medium and low polarity compounds (as it is the case of many UV
filters), simple set up, unattended operation and suitability of the
Analytical and Bioanalytical Chemistry, submitted
149
polydimethylsiloxane (PDMS) coated bars (Twisters) for thermal and organic
solvent desorption [13-14]. So far, several research groups have developed
successful applications of SBSE, in combination with GC-MS and LC-MS/MS,
for the determination of UV filters in water samples [15-18]. Despite the above
commented features, two relevant limitations of the SBSE technique are (1) the
relatively high cost of Twisters and (2) the risk of cross-contamination when
they are re-used. As an alternative to the commercial version of SBSE, in 2004,
Popp and co-workers [19-20] reported the use of disposable silicone sorbents
(DSS) as high capacity probes for the extraction and concentration of organic
compounds from water samples. After they early works, bulk silicone sorbents
have been used for direct and headspace (HS) sorptive extraction of organic
pollutants and natural compounds from water and food samples [21-24], as
well as for time-average passive sampling [25].
Technical grade silicone sorbents are inexpensive; thus, they are
normally considered as single use devices, avoiding carry-over problems
related with the incomplete desorption of target analytes. Moreover, the
amount and the format (tube, rods, sheets) of the sorbent can be adjusted for
each particular application. On the other hand, they may contain other
polymers, in addition to pure PDMS, as well as variable percentage of
additives, which can modify the efficiency of the extraction and/or interfere
with the determination of target analytes [26-27].
In this study, an effective, low cost sample preparation method for the
sensitive determination of eight UV filters in water samples is proposed. It
involves sorptive extraction of target analytes with a DSS followed by solvent
desorption, large volume injection (LVI) and GC-MS determination. Extraction
conditions were optimized using a small silicone disc (5 mm diameter x 0.6 mm
thickness), which can be easily desorbed in an autosampler vessel (1.5 mL
volume) with a few microlitres of a suitable organic solvent. Thereafter, the
efficiency of the extraction process was compared with (1) the theoretical values
Metodología desarrollada-Muestras acuosas
150
predicted from octanol-water partition coefficients (Kow) of UV-filters, and (2)
with the extraction yields provided by another silicone sorbent, in a rod format,
and with those achieved with the commercial Twisters. Finally, the applicability
of the method was demonstrated with river and sewage water samples.
2. Experimental
2.1. Solvents, standards and sorbents
HPLC-grade methanol, ethyl acetate and dichloromethane for trace
analysis were supplied by Merck (Darmstadt, Germany). UV filters standards
were purchased from Aldrich (Milwaukee, WI, USA) and Merck. Table 1
summarizes the analytes considered in this study and some properties of
relevance to optimize the sorptive extraction process. Individual solutions of
each species (ca. 1000 g mL-1) were prepared in methanol. Further dilutions
and mixtures of the UV filters were also dissolved in methanol and used to
fortify the water samples employed during optimization of extraction
conditions. Another set of standards (from 1 to 1000 ng mL-1) was made in ethyl
acetate and used to determine the amount of each analyte in the organic extract
obtained from silicone sorbents.
Technical grade silicone sorbents were acquired from Goodfellow (Bad
Nauheim, Germany) in two different formats: cord with a diameter of 1 mm
and sheets with a 0.6 mm thickness. Accordingly to the supplier, the
composition of the sorbent corresponds to phenyl-vinyl-methyl polysiloxane;
however, the relative percentage of the above substituents in the polysiloxane
skeleton is not provided. Rods (15 mm length x 1 mm diameter) were cut using
a sharp blade; discs (5 mm diameter x 0.6 mm thickness) were obtained by
pressing the silicone sheet with a sharp hollow punch (internal diameter 5 mm).
Discs and rods contained the same volume of silicone (ca. 12 L) and their cost
stayed below 0.1 Euro per unit. Twisters covered with 24 L of PDMS were from
Gerstel (Mühlheim, Germany). Twisters and silicone sorbents were soaked twice
with a methanol: acetone (1:1) solution, for 15 min, and then conditioned
overnight at 200 ºC, before being used for first time. Silicone discs and rods
Analytical and Bioanalytical Chemistry, submitted
151
were discarded after each extraction-desorption cycle. Twisters were
additionally soaked with methanol:acetone (1:1) after each use.
Oasis HLB (60 mg) solid-phase extraction (SPE) cartridges were acquired
from Waters (Milford, MA, USA).
Table 1. Abbreviated names, octanol-water partition coefficients (Kow), pKa values, GC-
MS retention times and quantification ions of target analytes.
Analyte Abbreviation aLog
Kow apKa
Retention
time (min)
Quantification
Ions (m/z)
2-Ethylhexyl salicylate EHS 5.77 8.13 15.41 120+138
Homosalate HMS 5.82 8.10 15.93,16.14 120+138
Isoamyl-p-methoxycinnamate IAMC 4.06 - 17.35b 161+178
2-Hydroxy-4-
methoxybenzophenone BP-3 3.64 7.56 17.36 151+227
3-(4-Methylbenzylidene)
camphor 4-MBC 4.95
- 17.57b 254
2-Ethylhexyl-p-
dimethylaminobenzoate EHPABA 6.15 2.39 19.10 161+178
2-Ethylhexyl-p-
methoxycinnamate EHMC 5.66 - 19.42b 165+277
Octocrylene OC 7.53 - 21.80 232+249+360
aValues compiled from SciFinder Scholar Database
bRetention time values for the E isomers
2.2. Samples and sample preparation
Ultrapure (Milli-Q), river, swimming pool and sewage water were used
in this study. Except ultrapure water, the rest of samples were passed through
0.45 m pore size filters before extraction. Unless otherwise stated, the
enrichment step was performed at room temperature (20 ± 2 ºC), with the
sorbent dipped into the water sample. Extractions were carried out in vessels
with two different nominal volumes (20 and 100 mL), furnished with a PTFE-
coated silicone septum and crimped with an aluminium cap. Silicone discs and
rods were pierced in a stainless steel pin, passed through the septum of the
vessel. During extraction, water samples were stirred with a PTFE covered
Metodología desarrollada-Muestras acuosas
152
magnetic bar. Extractions with Twisters were carried out under same
experimental conditions, using the PDMS covered bar as stirring device. After
finishing the sampling step, Twisters and silicone rubber sorbents were rinsed
with ultrapure water, dried using a soft tissue and desorbed with a small
volume of organic solvent.
Under final working conditions, extractions were carried out with a
silicone disc (5 mm x 0.6 mm) exposed overnight (14 hours) to 100 mL samples,
which contained a 10% of methanol. Thereafter, the disc was introduced in a
GC vial (1.5 mL volume) and desorbed with 0.2 mL of ethyl acetate. An aliquot
(9 L) of this extract was injected in the GC-MS system.
2.3. GC-MS determination
Analytes were determined by GC-MS, using a Varian (Walnut Creek,
CA, USA) 450 GC instrument connected to an ion-trap Varian 240 mass
spectrometer (MS) and equipped with an electron impact (EI) ionization source,
in the external configuration mode. Separations were carried out in an Agilent
(Wilmington, DE, USA) HP-5ms type capillary column (30 m x 0.25 mm i.d., df:
0.25 µm), operated at a constant helium flow of 1.2 mL min-1. The GC oven was
programmed as follows: 70 ºC (held for 4 min), first rate at 12 ºC min-1 to 280 ºC
(held for 5 min). The injector was furnished with Siltek fritted liner. The
injection volume was 9 L. The temperature of the injector was maintained at 60
ºC, for 1 min and then increased to 280 ºC with a rate of 200 ºC min-1. During the
first 0.2 min, the solenoid valve was maintained in the split position (split flow
20 ml min-1) to remove the excess of solvent and then, it was switched to the
splitless mode until 4 min. Thereafter, it was turned again to the split mode
and, a flow rate of 80 mL min-1 used to sweep the liner. Transfer line, ion source
and trap temperatures were set at 280, 200 and 150 ºC, respectively. The helium
damping gas flow in the mass analyzer was set at 2.5 mL min-1.
The mass spectrometer was operated in the electron impact ionization
mode (70 eV), with a filament emission current of 50 A. MS spectra were
acquired in the m/z range between 80 and 400 a.m.u. The electromultiplier
Analytical and Bioanalytical Chemistry, submitted
153
voltage was set at 1800 V. The m/z ratios summarized in Table 1 were used to
monitor the extracted ion chromatograms for target analytes.
2.4. Extraction efficiencies and sample quantification
The concentrations of UV filters in the ethyl acetate extracts from Twisters
and silicone sorbents were established by comparison with the responses (peak
areas) measured for calibration standards in the same solvent. The extraction
efficiency (EE) of the sample preparation method (extraction plus desorption)
was defined as the ratio between the mass of each analyte in the ethyl acetate
extract and that added to the water sample, multiplied by 100. Theoretical
extraction efficiencies (TE) were estimated with the following equation:
100
1
1x
K
TE
OW
, being the ratio between sample and sorbent volumes,
and Kow the octanol-water partition coefficients of UV filters summarized in
Table 1, [14,28].
The levels of UV filters in river and treated sewage water samples were
determined by comparison with aqueous standards, prepared in ultrapure
water, submitted to whole sample preparation procedure. Raw sewage water
samples were quantified with the standard addition methodology.
3. Results and discussion
3.1. Sample preparation conditions
Sample preparation (extraction and desorption) parameters were
optimized with fortified (10 ng mL-1) aliquots of ultrapure water. Unless
different conditions are specified, the extraction step was performed overnight
(12-14 hours), with the silicone disc exposed directly to the spiked samples
placed on a multi-position magnetic stirrer.
Metodología desarrollada-Muestras acuosas
154
3.1.1. Desorption parameters
Dichloromethane and ethyl acetate were considered to recover the
analytes previously concentrated in the silicone discs. Both solvents have been
recommended for the elution of UV filters from SPE cartridges [5,29] and, they
are compatible with GC-based determinations. In the initial experiments, 0.5
mL of solvent and 30 min of desorption were employed. Under these
conditions, ethyl acetate provided between 15% and 20% higher desorption
efficiencies than dichloromethane; thus, the former solvent was selected. Fig. 1
shows the normalized responses obtained in the consecutive desorptions of
silicone discs (5 mm diameter x 0.6 mm thickness) with 0.2 mL aliquots of ethyl
acetate. Around 95% of the extracted amount corresponded to the first fraction,
Fig. 1. Taking into account that discs were employed as single use devices, 0.2
mL was adopted as the working value for this factor. In a further series of
assays, desorption times of 10 and 20 min were also investigated; however, the
obtained responses were slightly lower than those corresponding to 30 min,
which was maintained as the optimum desorption time.
85%
90%
95%
100%
EHS HMS IAMC BP-3 4-MBC EHPABA EHMC OC
No
rma
lize
d re
spo
nse
Fraction 1 Fraction 2 Fraction 3
Fig. 1. Normalized responses in the consecutive desorptions of silicone discs with 0.2
mL of ethyl acetate. Desorption time 30 min. Average values for triplicate experiments.
Analytical and Bioanalytical Chemistry, submitted
155
3.1.2. Optimization of extraction conditions
3.1.2.1. Sampling mode and temperature
Sorptive extractions can be performed in direct and headspace (HS)
modes. The major advantage of the latter is the higher selectivity of the process,
avoiding the co-extraction of non-volatile organic interferences. In this study,
we have compared the extraction yields obtained for the direct mode, at room
temperature, versus those corresponding to HS at room temperature, 55 ºC and
90 ºC. In all cases, the sampling time was 14 hours and the sample volume 100
mL. HS extractions at room temperature rendered negligible yields for all
species. Responses obtained under the rest of conditions are compared in Fig. 2.
Globally, the most favourable situation corresponded to direct sampling at
room temperature; nevertheless, some analytes (both salicylates, 4-MBC,
EHPABA and EHMC) were also extracted in a considerable extent in the HS
mode, Fig. 2. For the above UV filters, the combination of HS sampling with
high temperatures offers an interesting balance between efficiency and
selectivity. Obviously, in this study, direct sampling and room temperature
were the adopted conditions.
0%
20%
40%
60%
80%
100%
120%
140%
160%
180%
EHS HMS IAMC BP-3 4-MBC EHPABA EHMC OC
Nor
mal
ized
res
pons
e
Direct, room temperature HS, 55 ºC HS, 90 ºC
Fig. 2. Effect of the sampling mode and the temperature in the performance of the
sorptive extraction step. Normalized responses to those obtained in the direct mode, at
room temperature. Sample volume 100 mL, extraction time 14 hours, n=3 replicates.
Metodología desarrollada-Muestras acuosas
156
3.1.2.2. Effects of salt and methanol
The influence of these variables on the yield of the sorptive extraction
was investigated using a sequential approach, considering three levels for the
concentration of sodium chloride (0, 5 and 15%) and four for the percentage of
methanol (0, 2, 5 and 10%). The obtained data followed two different trends,
which can be related with the octanol-water (Kow) partition coefficients of UV
filters. Fig. 3 depicts both trends using IAMC and OC as model analytes. For
medium polarity species (IAMC, 4-MBC and BP-3, log Kow from 3.6 to 5 units),
the extraction yield decreased steady with the percentage of methanol added to
the sample, whereas it improved slightly with the addition of sodium chloride.
This behaviour is the result of the negative (sodium chloride) and positive
(methanol) effects of the above factors in their water solubility. On the other
hand, for the more lipophilic compounds (EHS, HMS, EHPABA, EHMC and
OC, log Kow above 5.6 units), the extraction efficiency improved with the
percentage of methanol, achieving a maximum between 5% and 10% of organic
modifier, and underwent a dramatic diminution with the concentration of
sodium chloride, Fig. 3. Salt addition reduced the solubility of the compounds,
increasing sorption losses on the walls of glass vessels; moreover, it increased
the viscosity of the sample, which limits the migration rate of the less polar
analytes from the sample to the interface with the silicone disc. Obviously, the
addition of methanol prevented competitive adsorption losses. On the basis of
the above comments, sodium chloride was not added to water samples,
whereas the percentage of methanol was adjusted to 10%. These conditions
matched with those recommended for the sorptive extraction of UV filters using
PDMS covered stir bars [17].
Analytical and Bioanalytical Chemistry, submitted
157
0%
20%
40%
60%
80%
100%
120%
IAMC OC
Ext
ract
ion
effic
ienc
y
Without methanol 3% Methanol
5% Methanol 10% Methanol
A
0%
20%
40%
60%
80%
100%
120%
IAMC OC
Ext
ract
ion
effic
ienc
y
Without salt 5% salt 15% salt
B
Fig. 3. Influence of methanol (A) and sodium chloride (B) on the extraction efficiencies
of IAMC and OC. Sample volume 100 mL, extraction time 14 hours, n=3 replicates.
3.1.2.3. Stirring and sample pH
The extraction efficiency was enhanced significantly for stirred versus
non-stirred samples; however, for a sampling step of 14 hours, no differences
were appreciated using stirring rates of 300, 600 and 1000 rpm. The effect of
sample pH was investigated at three levels (3, 6 and 9). Despite BP-3 and
salicylate type UV filters are slightly acidic compounds, within the above range,
the pH showed a negligible effect in the extraction yield for all compounds,
data not shown. Further extractions were performed without adjusting the pH
Metodología desarrollada-Muestras acuosas
158
of water samples, unless it falls outside of the above range, and using a stirring
speed of 600 rpm.
3.1.2.4. Time and sample volume
In solid-phase microextraction processes, for a given amount of sorbent,
an increase in the volume of the sample leads to longer equilibrium times and
lower extraction efficiencies [30]. Fig. 4 depicts the time course of the sorptive
extraction for HMS and OC (all compounds showed similar profiles) in 20 and
100 mL samples. The Y axis in the figure represents the total amount (mass, in
ng) of each analyte in the ethyl acetate extract (0.2 mL) from the silicone disc.
For 20 mL samples, equilibrium was achieved after 5 hours of extraction,
whereas more than 12 hours were required for 100 mL ones. It is also evident
that, for sampling steps longer than 5 hours, the extracted amount of all
compounds was higher for 100 mL samples than for 20 mL ones. In further
experiments, 100 mL samples and 14 hours extractions (overnight) were
employed.
Analytical and Bioanalytical Chemistry, submitted
159
HMS
0
250
500
750
1000
0 5 10 15 20 25
Time (h)
Ext
ract
ed m
ass
(ng)
100 mL 20 mL
OC
0
250
500
750
1000
0 5 10 15 20 25
Time (h)
Ext
ract
ed m
ass
(ng)
100 mL 20 mL
Fig. 4. Time course of sorptive extraction for HMS and OC considering two different
sample volumes, n=2 duplicates.
3.2. Performance of the method
The performance of the proposed method, sorptive extraction with
disposable silicone discs followed by ethyl acetate desorption and GC-MS
determination, was first characterized in terms of linearity, intra- and inter-day
precision and limits of quantification (LOQs), Table 2. Linearity was assessed
with samples fortified at seven concentration levels, prepared in duplicate,
comprised between 0.01 and 10 ng mL-1. The plots of peak area versus
concentration were adjusted to a linear model obtaining determination
Metodología desarrollada-Muestras acuosas
160
coefficients (R2) values between 0.992 and 0.999 (Table 2); moreover, the
ANOVA lack-of-fit test demonstrated the suitability of the linear model to
describe the dependence between responses (peak areas) and concentration, at a
95% confidence level (data not given). Intra- and inter-day precision stayed
below 10% and 13%, respectively. These relative standard deviations (RSDs) are
similar to those reported for PDMS covered stir bars used in combination with
thermal desorption and GC-MS determination [17].
Procedural blanks, corresponding to ultrapure water samples (Fig. 5),
showed some traces of OC and EHMC, as it has been usually reported in
previous works [4,17,31]. The achieved LOQs, defined as the concentration of
each UV filter providing a signal 10 higher than the baseline noise, or the
standard deviation of 5 consecutive blanks (case of EHMC and OC), stayed at
the low ng L-1 level, except in case of BP-3 (LOQ 0.040 ng mL-1), which was the
compound showing the lowest extraction efficiency. Overall, LOQs
summarized in Table 2 are similar to those reported using Twisters in
combination with thermal desorption and GC-MS determination for 20 mL
samples [17] and other effective microextraction techniques, such as dispersive
liquid-liquid microextraction [32]; moreover, they remained below those
reported for solid-phase extraction and GC-MS determination [5].
Analytical and Bioanalytical Chemistry, submitted
161
Tab
le 2
. Lin
eari
ty (
0.01
0-10
ng
mL-
1 ), r
epea
tabi
lity,
rep
rodu
cibi
lity
and
limit
s of
qua
ntifi
cati
on (
LOQ
s) o
f the
pro
pose
d m
etho
d. D
epic
ted
data
corr
espo
nd to
100
mL
sam
ples
usi
ng a
sili
cone
dis
c (5
mm
dia
met
er x
0.6
mm
thic
knes
s) a
s so
rben
t.
Ana
lyte
L
inea
rity
(0.0
1-10
ng
mL
-1, n
=7
leve
ls)a
R
epea
tabi
lity,
RSD
(%)
(n=5
rep
licat
es, s
ame
day
)
Rep
rodu
cibi
lity,
RSD
(%)
(n=
12 r
eplic
ates
, 4 d
ays)
L
OQ
s (n
g m
L-1 )
Sl
ope
(SD
) In
terc
ept (
SD)
R2
0.2
ng m
L-1
2
ng m
L-1
0.
5 ng
mL
-1
EH
S
HM
S
IAM
C
BP-
3
4-M
BC
EH
PA
BA
EH
MC
OC
8.9E
4 (3
.4E
3)
5.9E
4 (2
.4E
3)
6.7E
4 (2
.3E
3)
1.3E
4 (2
.1E
2)
1.1E
4 (2
.7E
2)
1.7E
5 (6
.2E
3)
1.3E
5 (4
.1E
3)
8.4E
4 (2
.2E
3)
542
(124
3)
-106
0 (9
220)
-128
0 (8
112)
-117
0 (7
80)
323
(106
3)
3800
(179
0)
1135
0 (1
4800
)
1410
0 (7
800)
0.99
3
0.99
2
0.99
4
0.99
9
0.99
7
0.99
4
0.99
5
0.99
7
5.6
6.2
8.4
9.9
2.0
1.0
1.0
3.8
4.7
6.0
2.8
4.4
1.4
5.5
6.2
1.0
8.4
11.4
11.4
8.5
12.5
7.6
12.2
8.5
0.00
3
0.00
8
0.00
6
0.04
0.01
0.00
4
0.01
5
0.01
1
a The
line
ar r
espo
nse
rang
e of
BP
-3 w
as 0
.04-
10 n
g m
L-1
.
Metodología desarrollada-Muestras acuosas
162
15 16 17 18 19 minutes
0
5
10
15
20
(x 103) Counts
m/z 120+138+161+178
EHS
HMS
IAMCEHMC
16.75 17.00 17.25 17.50 17.75 minutes0.00
0.25
0.50
0.75
1.00
1.25
(x 103) Counts
m/z 151+227+254
BP‐3
4‐MBC
19.0 19.5 20.0 20.5 21.0 21.5 22.0 minutes0
1
2
3
4
5
6
7
8
(x 103) Counts
m/z 165+277+360+232+249
EHPABA OC
Fig. 5. Extracted ion chromatograms for a procedural blank (dotted line) and a low level
(0.07 ng mL-1) spiked ultrapure water sample (solid line).
Analytical and Bioanalytical Chemistry, submitted
163
The potential influence of the matrix on the efficiency of the sorptive
extraction was evaluated comparing the responses obtained for different water
samples, fortified at the same concentration level. River, swimming-pool and
sewage water were first filtered and then divided in two fractions. Thereafter,
one of them was fortified with UV filters at the same level as the aliquots of
ultrapure water. The difference between responses measured for spiked and
non-spiked fractions of real water samples were normalized to those obtained
for ultrapure water. Relative recoveries over 75% were obtained for river,
swimming-pool and treated wastewater, Table 3. On the other hand, the
efficiency of the extraction underwent a reduction up to 50% for the most
lipophilic species (OC) in raw sewage water. Thus, external calibration against
aqueous standards, dissolved in ultrapure water, can be used to measure the
levels of UV filters in surface, bathing and treated sewage water; however, the
standard addition methodology is required for untreated wastewater.
Table 3. Influence of the sample matrix on the efficiency of the sorptive extraction using
silicone discs. Normalized responses to those measured for ultrapure water fortified at
the same concentration level (2 ng mL-1), n=3 replicates.
Analyte Normalized response (%) SD
River Swimming pool Treated wastewater Raw wastewater
EHS
HMS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OC
104 3
96 3
92 6
98 7
90 7
94 1
98 5
104 5
93 6
87 2
87 5
76 6
82 12
89 2
88 2
89 2
93 11
92 9
86 4
75 4
86 7
91 1
91 11
92 1
62 1
57 5
82 9
108 7
91 8
66 5
60 3
49 3
Metodología desarrollada-Muestras acuosas
164
3.3. Silicone discs versus rods and Twisters
Table 4 compiles the experimental extraction efficiencies (EE) obtained
with silicone discs, rods and Twisters for 20 and 100 mL water samples, spiked
at different concentration levels. Theoretical extraction efficiencies (TE),
calculated as described in section 2.4, are also given. In order to prevent
changes in the extraction efficiencies caused by incomplete desorptions, the
volume of ethyl acetate was increased up to 0.5 mL for rods and Twisters, which
were not completely covered, in the 1.5 mL vessels, with 0.2 mL of solvent.
Overall, an excellent agreement was observed among EE values provided by
silicone discs and rods containing 12 L of silicone for 20 and 100 mL samples.
The same comment is also valid for the experimental efficiencies achieved using
two discs (total silicone volume 24 L) and a single Twister (24 L of pure
PDMS), Table 4. For the most lipophilic UV filters (EHS, HMS, EHMC,
EHPABA and OC), EE values were similar to TE ones; moreover, the above UV
filters were recovered in an extension higher than 90% in all the tested
conditions. In the case of IAMC and 4-MBC (log Kow 4.06 and 4.95,
respectively), the EE were in agreement with TE for 20 mL samples and slightly
lower for the 100 mL ones. Finally, the EE for BP-3 always remained 2-3 times
lower than the TE value. Solid-phase extraction of the water sample [33], after
finishing the sorptive extraction step, confirmed that BP-3 remained un-
extracted in the water phase, without undergoing significant degradation
processes.
Analytical and Bioanalytical Chemistry, submitted
165
Table 4. Experimental extraction efficiencies (EE) and theoretical values (TE) for spiked
samples of ultrapure water (pH 6, 10% methanol) using different sorbents. Sampling
time 14 hours, n=3 replicates.
Analyte Sample volume 20 mLa
EE (%) SD TE (%)
Disc (12 L) Rod (12 L) Disc (2 x 12 L) Twister (24 L) (12 L) (24 L)
EHS
HMS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OC
105 3
101 6
79 3
18 5
88 4
95 3
99 1
104 2
104 2
109 6
86 10
20 6
87 5
99 2
105 7
106 2
95 4
91 2
87 8
28 2
92 2
94 2
93 2
94 5
104 8
104 7
98 4
36 3
102 1
108 10
105 3
102 7
100
100
87
72
98
100
100
100
100
100
93
84
99
100
100
100
aAddition level 1 ng mL-1.
Table 4 cont. Experimental extraction efficiencies (EE) and theoretical values (TE) for
spiked samples of ultrapure water (pH 6, 10% methanol) using different sorbents.
Sampling time 14 hours, n=3 replicates.
Analyte Sample volume 100 mLb
EE (%) SD TE (%)
Disc (12 L) Rod (12 L) Disc (2 x 12 L) Twister (24 L) (12 L) (24 L)
EHS
HMS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OC
92 8
101 5
47 6
8 2
54 5
93 2
94 6
104 9
95 8
96 6
44 64
7 2
50 3
91 2
85 6
91 7
96 4
102 1
54 2
19 3
76 3
96 7
98 1
101 1
106 4
106 1
65 8
26 5
92 1
102 1
99 2
94 5
99
99
58
34
91
99
98
100
99
99
73
51
96
100
99
100
bAddition level 0.2 ng mL-1.
Metodología desarrollada-Muestras acuosas
166
A potential explanation of the low extraction yields observed for BP-3 is
that, the methanolic content of the samples (10%) resulted in a significant
reduction of the partition coefficients for UV filters between the silicone (or
PDMS) sorbent and the matrix. Despite such reduction, the most lipophilic
species were still quantitatively transferred to silicone sorbents; however, in the
case of BP-3 the efficiency of the extraction was significantly reduced. Another
feasible hypothesis is that, Kow values are not valid to predict the distribution of
certain compounds between water samples and polysiloxane sorbents. In this
sense, Popp and co-workers [34] have also found important disagreements
between predicted and observed extraction yields for other phenolic species,
using technical grade silicone sorbents. Serodio and co-workers [35] reported
similar deviations, for medium polarity species, using PDMS covered stir bars.
Fig. 6 shows the total ion current (TIC) chromatograms corresponding to
desorption of a Twister and two silicone discs. In both cases, the corresponding
ethyl acetate extracts were adjusted to a final volume of 0.2 mL, and an aliquot
(9 L) was injected in the GC-MS system. As appreciated, any of both materials
released relevant levels of interfering species.
16 17 18 19 20 21 22 min.0
10
20
30
40
(x 103 Counts)
Silicone disc
Twister
Fig. 6. Total ion chromatograms (TIC) corresponding to ethyl acetate extracts (final
volume 0.2 mL) from a Twister and two silicone discs.
Analytical and Bioanalytical Chemistry, submitted
167
3.3. Application to water samples
The developed method was applied to a total of twelve grab samples of
river and sewage water collected in three different dates. Concentrations for
compounds found above their LOQs are provided in Table 5. Samples code 1 to
3 correspond to a river which receives the discharges from an urban STP,
equipped with primary and biological treatments, in the Northwest of Spain.
The highest concentrations in the above river water samples were found at the
end of summer and the beginning of autumn, as a consequence of the reduction
in the flow rate of the river. Samples 4 to 6 were obtained from bathing areas in
small streams, located in the same geographic area. The pairs of samples 7-8, 9-
10 and 11-12 were simultaneously collected in the inlet and outlet streams of the
urban STP. EHS and OC were found in the six samples at concentrations above
their LOQs. For the latter compound, the measured levels were slightly higher
than those previously reported by our group for samples taken in the same
plant in 2009 [32], and those found in other STP in Spain [18,31]. On the other
hand, they are similar to those reported for STPs in Switzerland [4]. Besides
EHS and OC, 4-MBC and EHMC were also found in a significant percentage of
the processed sewage water samples, achieving concentrations above 100 ng L-1,
Table 5. Finally, IAMC and EHPABA were never detected.
Metodología desarrollada-Muestras acuosas
168
Table 5. UV filter concentrations measured in river and wastewater samples. IAMC
and EHPABA were not detected in any sample, n= 3 replicates.
Code Type Sampling
date
Concentration ( ng L-1) ± SD
EHS HMS BP-3 4-MBC EHMC OC
1 River July 2010 - - - - - 27 ± 3
2 River September 2010 4.0 ± 0.3 8 ± 2 - 46 ± 1 39 ± 4 243 ± 18
3 River October 2010 15 ± 1 15 ± 2 - 48 ± 6 - 251 ± 30
4 River September 2010 - - - - 27 ± 1 19 ± 2
5 River September 2010 - - - - 24 ± 1 27 ± 11
6 River September 2010 - - 53 ± 6 - - -
7 R.W. July 2010 105 ± 11 40 ± 7 - 17 ± 5 757 ± 60 831 ± 73
8 T.W. July 2010 4.5 ± 0.1 - - 223 ± 5 - 106 ± 11
9 R.W. September 2010 36 ± 1 25 ± 1 236 ± 5 131 ± 16 129 ± 3 437 ± 25
10 T.W. September 2010 4.8 ± 0.1 15 ± 1 - 140 ± 7 31 ± 1 111 ± 1
11 R.W. October 2010 22 ± 2 14 ± 1 - 61 ± 5 27 ± 4 209 ± 26
12 T.W. October 2010 5.3 ± 0.1 - - - - 65 ± 1
R.W. raw wastewater T.W. treated wastewater - below detection limits
4. Conclusions
Sorptive extraction using technical grade, disposable silicone sorbents
followed by solvent desorption and GC-MS determination, with large volume
injection, provides suitable figures of merit for the determination of UV filters
in surface and sewage water samples. Optimum extraction conditions are
similar to those previously reported in the literature for Twisters covered with
PDMS; moreover, very close experimental extraction efficiencies were
calculated for both materials. Practical advantages of the methodology
described in this study are (1) the low cost of silicon discs (ca. 0.1 Euro per unit),
Analytical and Bioanalytical Chemistry, submitted
169
(2) the low volume of the sorbent (12 L), which can be easily accommodate at
the bottom of a GC vial and desorbed using just 0.2 mL of organic solvent, and
(3) the possibility to re-analyze the extract, without repeating the whole sample
preparation process, as it occurred in the case of thermal desorption methods.
The slow kinetics of the sorptive extraction for 100 mL samples can be
compensated by processing simultaneously, in an unattended mode, a large
number of specimens, using a multi-position stirrer.
Acknowledgements
This study has been supported by the Spanish Government and E.U.
funds (projects CTQ2009-08377 and DE2009-0020). N.N. thanks a FPU pre-
doctoral contract to the Spanish Ministry of Education.
Metodología desarrollada-Muestras acuosas
170
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1.5. Publicación:
SENSITIVE DETERMINATION OF
SALICYLATE AND BENZOPHENONE
TYPE UV FILTERS IN WATER SAMPLES
USING A SOLID-PHASE MICROEXTRACTION,
DERIVATIZATION AND GAS CHROMATOGRAPHY
TANDEM MASS SPECTROMETRY
N. Negreira, I. Rodríguez, M. Ramil, E. Rubí, R. Cela
Analytica Chimica Acta 638 (2009) 36
(doi:10.1016/j.aca.2009.02.015)
Analytica Chimica Acta 638 (2009) 36
175
Sensitive determination of salicylate and benzophenone type UV filters in
water samples using solid-phase microextraction, derivatization and gas
chromatography tandem mass spectrometry
N. Negreira, I. Rodríguez*, M. Ramil, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto
de Investigación y Análisis Alimentario, Universidad de Santiago de
Compostela, Santiago de Compostela 15782, Spain.
Abstract
A sensitive procedure for the determination of three UV filters:
ethylhexyl salicylate (EHS), 3,3,5-trimethylcyclohexyl salicylate (Homosalate,
HMS), 2-hydroxy-4-methoxybenzophenone (BP-3) and two related
hydroxylated benzophenones (2,4-dihydroxybenzophenone, BP-1 and 2,2´-
dihydroxy-4-methoxybenzophenone, BP-8) in water samples is presented.
Analytes were first concentrated on the coating of a solid-phase microextraction
(SPME) fibre, on-fibre silylated and then determined using gas chromatography
combined with tandem mass spectrometry (GC-MS/MS). Factors affecting the
performance of extraction and derivatization steps are thoroughly evaluated
and their effects on the yield of the sample preparation discussed. Under final
working conditions, a PDMS-DVB coated SPME fibre was exposed directly to
10 mL of water, adjusted at pH 3, for 30 min. After that, the fibre was placed in
the headspace (HS) of a 1.5 mL vial containing 20 L of N-methyl-N-
(trimethylsilyl)-trifluoroacetamide (MSTFA). On-fibre silylation of hydroxyl
groups contained in the structure of target compounds was performed at 45 ºC
for 10 min. The whole sample preparation process was completed in 40 min,
providing limits of quantification from 0.5 to 10 ng L-1 and acceptable precision
(RSDs under 13%) for samples spiked at different concentrations. All
compounds could be accurately determined in river and treated wastewater
(relative recoveries from 89 to 115%) using standards in ultrapure water,
Metodología desarrollada-Muestras acuosas
176
whereas standard addition is recommended to quantify their levels in untreated
wastewater. Analysis of wastewater revealed the systematic presence of BP-3
and BP-1 in raw samples with maximum concentrations close to 500 and 250 ng
L-1, respectively.
Keywords: UV filters; water samples; solid-phase microextraction;
derivatization; gas chromatography.
1. Introduction
Concern about skin damage produced by UV radiation has fostered the
consumption of sunscreen products in developed countries. In the European
Union (EU), 26 organic compounds have been approved to be used as UV filters
in sunscreen products, with maximum individual concentrations up to 10%
[1,2]. Additionally, same compounds, as well as other species with similar
structures, are also included in the formulation of many other personal care
products (PCPs) such as hair shampoos, lipsticks and even in packaging
materials to enhance their light stability [3-6].
Direct release of sunscreens from the skin in bathing areas and indirect
inputs through domestic wastewater are responsible for local and diffuse
discharges of UV filters in the aquatic environment [1,7]. Among them, 2-
hydroxy-4-methoxybenzophenone (BP-3) is one of the most often detected,
reaching concentrations up to 100 ng L-1 in surface water from bathing areas
[8,9]. It is also present in raw [10] and treated domestic wastewater [11], and
even in fatty tissues from river and lake fish [10]. Moreover, when applied on
the skin, BP-3 is partially absorbed by the human body and excreted as more
polar metabolites, such as 2,4-dihydroxybenzophenone (BP-1) and 2,2´-
dihydroxy-4-methoxybenzophenone (BP-8) [12,13]. BP-1 and BP-8 are also used
as UV absorbers to protect goods against UV radiation [3]. In vivo and in vitro
studies have proved that BP-3 and BP-1 present estrogenic activity [14,15];
moreover, they are prone to evolve into halogenated by-products when mixed
Analytica Chimica Acta 638 (2009) 36
177
with chlorinated water [16]. On the other hand, BP-8 presents activity as
bacterial mutagen [1,17].
In addition to BP-3, two salicylates: 2-ethylhexyl salicylate (EHS) and
3,3,5-trimethylcyclohexyl salicylate (Homosalate, HMS) are also often employed
as UV filters in sunscreens [5]. Although they are more lipophilic than the
above benzophenones, the presence of a phenolic group in their structures
might provide them certain mobility in the aquatic environment. In fact, HMS
and EHS have been recently detected in surface water samples [8,18].
Solid-phase extraction (SPE) is the preferred technique for the
concentration of UV filters in water samples [1]. Although it offers considerable
advantages in comparison with liquid-liquid extraction (LLE), SPE still requires
large sample intakes (from 0.3 to 1 L), a considerable volume (10-15 mL) of
organic solvents for analytes desorption and a further clean-up to compensate
for its limited selectivity when applied to wastewater [1,8,9,11]. Theoretically,
microextraction techniques, based on equilibrium processes, and particularly
solid-phase microextraction (SPME), should allow overcoming some of the
above drawbacks. Although SPME, followed by gas chromatography with mass
spectrometry detection (GC-MS), has been previously evaluated for the
concentration of several UV filters in aqueous matrices (including BP-3 and
some related benzophenones) the achieved limits of quantification (LOQs), in
the ng mL-1 range, were unsuitable for waste and surface water analysis [12,19].
The well-known trend of phenolic species to produce tailing GC peaks, added
to a non-exhaustive optimization of extraction conditions, can explain the poor
sensitivity of the resulting methods. The former limitation can be overcome
integrating a derivatization step in the SPME process. In this sense, SPME
combined with on-fibre derivatization has been described as a successful
approach for the determination of emerging pollutants, with carboxylic and
phenolic groups, in water samples at the low ng L-1 level [20,21].
The goal of this work is to evaluate the suitability of SPME, followed by
on-fibre silylation, for the determination of HMS, EHS, BP-3 and two other
hydroxylated benzophenones (BP-1 and BP-8) in waste and surface water
Metodología desarrollada-Muestras acuosas
178
samples, using GC with tandem mass spectrometry (MS/MS) as determination
technique. Although, the use of silylation reactions is quite common to enhance
the detectability of salicylates and hydroxylated benzophenones in GC methods
[8,16,22,23], to the best of our knowledge, the combination of these reactions
with SPME has not yet been proposed for the determination of phenolic UV
absorbers in water samples.
2. Experimental
2.1. Reagents, SPME equipment and samples
HPLC-grade methanol, acetone and ethyl acetate (trace analysis grade),
and glacial acetic acid were obtained from Merck (Darmstadt, Germany).
Standards of BP-3, BP-1, BP-8, EHS and HMS were acquired from Aldrich
(Milwaukee, WI, USA) and Merck. Their chemical structures, pKa and octanol-
water partition coefficients (log Kow) are summarized in Table 1. Derivatization
reagents N-methyl-N-(tert-butyldimethylsilyl)-trifluoroacetamide (MTBSTFA)
and N-methyl-N-(trimethylsilyl)-trifluoroacetamide (MSTFA) were also
purchased from Aldrich. Individual solutions of each compound were prepared
in methanol, further dilutions and mixtures of them, used to fortify water
samples, were made in the same solvent. Silylated compounds in ethyl acetate,
employed during optimisation of GC-MS/MS conditions, were prepared
adding 20 L of MSTFA to a standard solution of target species in the above
solvent. The reaction was accomplished at 60 ºC for 1 hour [8].
Analytica Chimica Acta 638 (2009) 36
179
Table 1. Abbreviated names, structures, log Kow and pKa values of target species.
Abbreviation Name Structure Log
Kowa pKaa
EHS 2-ethylhexyl salicylate
5.97 8.13
HMS
3,3,5-trimethylcyclohexyl
salicylate
(Homosalate)
6.16 8.09
BP-3 2-hydroxy-4-
methoxybenzophenone
3.79 7.56
BP-1 2,4-
dihydroxybenzophenone
3.17 7.53
BP-8 2,2´-dihydroxy-4-
methoxybenzophenone
3.93 6.99
aValues obtained from SciFinder Scholar Database,
http://www.cas.org/products/sfacad/
A manual SPME holder and fibres coated with different polymers:
poly(dimethylsiloxane) (PDMS, 100 m film thickness), polyacrylate (PA, 85
m film thickness), Carboxen-PDMS (CAR-PDMS, 75 m film thickness) and
PDMS-divinylbenzene (PDMS-DVB, 65 m film thickness) were obtained from
Supelco (Bellefonte, PA, USA). Before being used for first time, SPME fibres
were thermally conditioned following conditions recommended by the
supplier.
Ultrapure (Milli-Q), river and wastewater samples, obtained from the
inlet and outlet streams of an urban sewage treatment plant (STP), equipped
with primary and activated sludge units, were employed throughout this study.
Metodología desarrollada-Muestras acuosas
180
River and wastewater samples were filtered using cellulose acetate membranes
(0.45 m pore size) and stored in the dark at 4º C until analysis.
2.2. Sample preparation
Extractions were performed in glass vessels of different nominal
capacities (10, 22 and 110 mL), containing a magnetic stir bar (10 mm x 4 mm)
and sealed with a PTFE layered rubber septa. Under optimised conditions, a
PDMS-DVB fibre was exposed directly to aqueous standards, prepared in
ultrapure water, and real life water samples, previously adjusted at pH 3, for 30
min. Extractions were carried out in the smallest vessels containing 10 mL of
water, at room temperature (20 ± 2 ºC), with magnetic stirring (1200 rpm). After
that, the fibre was retracted into the SPME holder. Water drops attached to the
outlet surface of the metallic needle were removed with a soft paper tissue and
the fibre was exposed the HS of a vessel containing 20 L of MSTFA. On-fibre
silylation of salicylates and benzophenones was performed at 45 ºC for 10 min.
2.3. Determination
Analytes were determined by GC-MS/MS using a Varian (Walnut Creek,
CA, USA) CP 3900 gas chromatograph connected to an ion-trap mass
spectrometer (Varian Saturn 2100). Separations were carried out in a HP-5ms
capillary column (30 m x 0.25 mm I.D., df: 0.25 m) supplied by Agilent
(Wilmington, DE, USA). Helium (99.999 %) was used as carrier gas at a constant
flow of 1 mL min-1. The GC oven was programmed as follows: 70 ºC (held for 3
min), at 12 ºC min-1 to 280 ºC (held for 10 min). The GC-MS interface and the ion
trap temperatures were set at 280 ºC and 220 ºC, respectively. Silylated
standards prepared in ethyl acetate were injected in the splitless mode (splitless
time 1 min), with the injector port at 280 ºC. SPME fibres were desorbed at 270
ºC in case of PDMS-DVB, and 280 ºC for the rest of coatings, for 3 min,
maintaining the injector in the splitless mode during this period.
The mass spectrometer was operated in the electron impact ionisation
mode (70 eV). MS spectra were recorded in the range from 70 to 500 m/z units.
Analytica Chimica Acta 638 (2009) 36
181
The base peak in the spectra of each compound, as trimethylsilyl derivative,
was isolated with a window of 3 m/z units and subjected to collision induced
dissociation.
3. Results and discussion
3.1. Performance of GC-MS/MS determination
Optimisation of GC-MS/MS conditions was carried out using silylated
standards prepared in ethyl acetate. Initially, MTBSTFA was selected as
derivatization reagent, since it leads to derivatives with an excellent stability, it
has been previously employed for the silylation of BP-3 and BP-1 [16], and it is
commonly employed in on-fibre derivatization methodologies [20,21].
Although different experimental conditions (reaction time, temperature and
volume of MTBSTFA) were tested, the two salicylates did not react
quantitatively with MBTSTFA. On the other hand, using MSTFA, all
compounds were converted in the corresponding trimethylsilyl derivatives. As
reported previously, two chromatographic peaks were obtained for HMS [8].
The base peak in the MS spectra of the three benzophenones showed the lost of
a methyl group, appearing at 285, 343 and 373 m/z units for BP-3, BP-1 and BP-
8, respectively [22]. In the case of salicylates, an intense fragment at 195 m/z
units, corresponding to the dimethylsilyl-2-hydroxy benzoic acid moiety, was
observed. The above ions were isolated in the trap and further fragmented
using a resonant waveform. The most intense transitions in the MS/MS spectra
of the two salicylates revealed the loss of H2O and CO2, Fig. 1. BP-3 and BP-1
showed a main transition corresponding to the removal of fragments with 43
and 72 m/z units, respectively. The first corresponds to a SiCH3 moiety,
whereas the second might represent the replacement of a Si(CH3)3 group,
bonded to one of the aromatic hydroxyls in the silylated BP-1, by hydrogen. On
the other hand, BP-8 underwent a much more complex fragmentation leading
to several product ions, Fig. 1. The most intense ones, at 329, 330 and 301 m/z
units (corresponding to the loss of 44, 43 and 72 units of mass, respectively)
were used for the quantification of this compound. It is worth noting that it was
Metodología desarrollada-Muestras acuosas
182
not possible to fragment BP-8 using the non-resonant mode. Under optimised
conditions, see Table 2, the GC-MS/MS system provided linear responses for all
compounds (HMS was quantified as sum of peaks) in the range between 5 and
1000 ng mL-1. Limits of quantification (LOQs), defined as the concentration of
each compound producing a signal ten times higher than the baseline noise in
the corresponding GC-MS/MS chromatograms, ranged from 0.2 to 1.1 ng mL-1,
Table 2. Except for BP-8, these values were 10 times lower than those achieved
using single MS detection.
75 100 125 150 175 m/z
0%
25%
50%
75%
100%
91 135
151
177
O
O
O
Si
O
O
O
Bu-n
Et
SiEHS
HMS
150 200 250 300 m/z
0%
25%
50%
75%
100%
253
271
325343
BP-1
100 150 200 250 m/z
0%
25%
50%
75%
100%
212
242
267
285
O
MeO
Ph
O
Si
BP-3
150 200 250 300 350 m/z
0%
25%
50%
75%
100%
144
181209
239
256 283
301
315
329
343357
373
BP-8
Fig. 1. MS/MS spectra for UV absorbers.
Analytica Chimica Acta 638 (2009) 36
183
Table 2. Optimal MS/MS detection conditions, correlation coefficients and
instrumental LOQs for the trimethylsilyl derivatives of analytes.
Compound Parent
ion
(m/z)
Product ions
(m/z)
Excitation
amplitude
(V)
Storage
level
(m/z)
Correlation
Coefficient
(R2)a
LOQs
(ng mL-
1)b
EHS
HMS
BP-3
BP-1
BP-8
195
195
285
343
373
177,151
177,151
242
271
329,330,301
0.75
0.90
1.10
1.19
1.50
86
86
126
151
164
0.998
0.999
0.995
0.995
0.997
0.2
0.3
0.2
0.3
1.1
a Evaluated with standards at 6 different concentrations between 5 and 1000 ng mL-1.
b Defined as the concentration producing a peak with a signal to noise (S/N) ratio of 10.
3.2. Optimisation of SPME conditions
3.2.1. Preliminary experiments
In order to asses the feasibility of the on-fibre silylation reaction, as well
as to obtain a first evaluation of the extraction capabilities of different SPME
coatings, spiked aliquots of ultrapure water, adjusted at pH 3, were extracted in
the direct mode using four different SPME fibres, for 30 min. The Carbowax-
DVB fibre, whose affinity for benzophenones had been previously
demonstrated [12], was not included in this study since its commercialization
has been recently stopped. After extraction, fibres were exposed to the HS of a
vial containing 20 L of MSTFA at 60 ºC for 15 min. GC-MS was used as
detection technique. Whatever the SPME coating, neither BP-3 nor salicylates
were obsereved in the GC-MS chromatograms monitored selecting the most
intense ions in the mass spectrum of free compounds (151+227 and 120+138,
respectively); thus, it can be assumed that they are converted quantitatively into
the corresponding silyl derivatives. The same behaviour was noticed for BP-1
and BP-8; however, it must be noted that the presence of two hydroxyl groups
in the structure of both species prevents their sensitive detection using GC
based methods. Fig. 2 compares the responses, as peak areas, obtained with the
tested SPME fibres. The lowest extraction efficiency corresponded to the CAR-
PDMS one. Probably, the molecular size of the analytes is too large to allow
Metodología desarrollada-Muestras acuosas
184
their diffusion into the porous structure of the Carboxen sorbent. The PDMS
coating provided high responses for the less polar salicylates but failed to
extract the di-hydroxylated benzophenones. The highest efficiencies were
achieved with PDMS-DVB, except in case of BP-1 (the most polar specie), which
presented a higher affinity for the PA coating. On the basis of these data,
PDMS-DVB and PA fibres were selected for further experiments. The relative
standard deviation values (RSDs) obtained with both fibres for replicate
extractions (n=4) of samples spiked at the 4 ng mL-1 level remained under 12%.
Carry-over effects were evaluated by desorbing each fibre twice. Relative
responses up to 5% were noticed for some compounds, particularly EHS and
HMS, in the second desorption; therefore, to avoid cross contamination
problems, fibres were additional desorbed for 3 min at 270 and 280 ºC (PDMS-
DVB and PA, respectively) using a nitrogen flow of 30 mL min-1 in the split
injector of a non-operative GC instrument.
0,0E+00
5,0E+05
1,0E+06
1,5E+06
2,0E+06
2,5E+06
3,0E+06
3,5E+06
EHS HMS BP-3 BP-1 BP-8
Pe
ak
are
a
PDMS-DVB PA PDMS CAR-PDMS
Fig. 2. Comparison of responses for different SPME fibres. Direct sampling for 30 min.
Analytica Chimica Acta 638 (2009) 36
185
3.2.2. Extraction parameters
The influence of different parameters on the performance of the
microextraction step was evaluated with aliquots of ultrapure water, spiked
with target compounds at 4 ng mL-1. The SPME fibre was immersed in the
water sample or maintained in the headspace (HS) of the vessel, depending on
the experiment, and then exposed to the HS of a 1.5 mL GC autosampler vial
containing MSTFA. Unless otherwise is stated, 22 mL glass vessels were used in
the sampling step. On-fibre derivatization was carried out under conditions
reported in the above paragraph and GC-MS/MS was used as detection
technique.
3.2.2.1. Sample pH, stirring and fibre selection
Optimal values for these factors were simultaneously investigated using
a two-level factorial experimental design, with two central points. Extractions
were carried out in the direct sampling mode, using 20 mL samples, without
salt addition, for 30 min. Table 3 shows extraction conditions corresponding to
the 10 experiments involved in the design, as well as the responses (peak areas)
measured for target species. Fig. 3 depicts the main effects for each factor. The
length of plotted bars is proportional to the change in the response of a given
compound when the associated factor varies from the low to the high level
within the domain of the design. A positive sign indicates an increase in the
observed response, whereas a negative value shows the opposite effect. Dotted
lines correspond to the statistic significance limit, established for a 95%
confidence level. The pH of water samples showed a negative effect on the
efficiency of the extraction step for all species. On the other hand, stirring
played a positive effect and it was statistically significant for the less polar
species, which presumably present the slower diffusion kinetics from the bulk
of the sample to the interface with the SPME fibre. The effect of the SPME
coating was compound dependant. BP-3 and BP-8 showed at higher affinity for
the PDMS-DVB fibre than for the PA one, whereas, the second coating is
preferred for the most polar BP-1. In case of salicylates, the influence of the
Metodología desarrollada-Muestras acuosas
186
factor fibre was negligible. Two-factor interactions played a minor influence on
the SPME process, figure not shown. Stirring of water samples at 1200 rpm,
using a PTFE covered magnetic bar, pH 3 and PDMS-DVB fibres were selected
for further experiments.
Table 3. Experimental conditions employed in the 23 experimental factorial design and
responses (peak areas) attained for each compound.
Exp. Fibre pH Stirring (rpm) Response
EHS HMS BP-3 BP-1 BP-8
1 PA 6 600 1961102 1732152 869855 2291319 885819
2 PDMS-DVB 3 1200 2859116 2517074 2542495 2312986 5986178
3 PA 3 600 2360417 2158650 1223590 3139505 1398965
4 PA 3 1200 3026911 2968394 1150444 3991787 3300669
5 PDMS-DVB 6 1200 3074981 2965423 2855940 2705516 6943826
6 PDMS-DVB 3 600 2537813 2224669 2136804 1911934 2207861
7 PDMS-DVB 6 600 1854872 1662130 1483205 1461656 1566350
8 PDMS-DVB 4.5 900 2552540 2399962 1409212 668191 1744285
9 PA 4.5 900 2566060 2372379 1200280 2771724 2076543
10 PA 6 1200 2669330 2513666 1373519 3112942 2674027
Analytica Chimica Acta 638 (2009) 36
187
-4 -2 0 2 4 6 8
EHS
HMS
BP-3
BP-1
BP-8
Standardized main effect values
pH (3-6) Stirring (600-1200 rpm) Fibre (PA-PDMS/DVB)
Fig. 3. Combined Pareto chart showing the main effects of pH, stirring and type of fibre
on the performance of the extraction step. Low and high values of each factor are given
in the legend of the figure.
3.2.2.2. Sampling mode, temperature and ionic strength
The influence of the sampling mode was assessed with 15 mL aliquots of
a spiked water sample placed in 22 mL vessels. Extractions were carried out for
30 min, at room temperature (20 ± 2 ºC) and 100 ºC, maintaining the fibre in the
HS of the vessel or dipping it into the sample. For the two salicylates, little
differences were observed among responses obtained under different explored
experimental conditions. On the other hand, benzophenones could be hardly
detected using HS extraction, even at 100 ºC. As regards direct sampling,
around 10-fold higher responses were achieved at room temperature versus 100
ºC for these three compounds. Although, increasing the temperature of the
sample speeds up mass transference kinetics from the water sample to the fibre,
it produces a diminution in the affinity of compounds for the PDMS-DVB
Metodología desarrollada-Muestras acuosas
188
coating. Direct sampling, at room temperature was maintained as the extraction
mode.
The effect of the ionic strength on the extraction process was investigated
with samples containing different concentrations of sodium chloride between 0
and 250 mg mL-1. Fig 4 shows the pattern followed by EHS, BP-8 and BP-1.
HMS and BP-3 behaved as EHS and BP-8, respectively. For the most polar
compound, BP-1, the efficiency of the extraction rose slightly with the ionic
strength of the water sample; whereas for the most lipophilic species (EHS and
HMS) it remained constant between 0 and 50 mg mL-1 of NaCl and then started
to decrease. For BP-8 and BP-3, with intermediate log Kow constants, a
maximum was noticed for 50 mg mL-1 of NaCl. Operating in direct immersion,
the addition of salt to the water sample benefices the thermodynamics of the
SPME but slows down its kinetics due to the increase in the viscosity of the
sample. Normally, the result of this balance is an improvement in the efficiency
of the extraction for the most polar species and a diminution for those with
lower water solubility [20], which matches with results depicted in Fig 4.
Despite above results, in order to simplify the setup of the extraction process
and also to avoid the build up of solid residues in the liner of the GC, no salt
was added to the samples in further experiments.
Analytica Chimica Acta 638 (2009) 36
189
0,0E+00
1,0E+06
2,0E+06
3,0E+06
4,0E+06
5,0E+06
6,0E+06
7,0E+06
8,0E+06
0 50 100 150 200 250
NaCl (mg mL-1)
Pea
k a
rea
EHS BP-1 BP-8
Fig. 4. Influence of NaCl concentration on the efficiency of the SPME using the PDMS-
DVB fibre. Values for duplicate extractions.
3.2.2.3. Sample volume and extraction time
According to the SPME theory, in a two-phase system (sample and SPME
coating) the amount of analyte incorporated in the fibre increases initially with
the volume of sample and then, after a given value, it reaches a plateau. In
practise, the size of the sample affects also to kinetics of SPME processes. Fig. 5
compares the responses obtained for aliquots of the same spiked water sample,
poured in vessels with different nominal capacities: 10, 22 and 110 mL. In all
cases, the free HS was limited to the minimum (around 0.5 mL to allow effective
stirring of the samples) and extractions were performed in the direct mode, for
30 min. As observed, the neat effect of the sample volume was negligible, thus
the smaller vessels with capacity to accommodate a stir bar and 10 mL of water
were selected to continue this study.
Metodología desarrollada-Muestras acuosas
190
0,0E+00
2,0E+06
4,0E+06
6,0E+06
8,0E+06
1,0E+07
EHS HMS BP-3 BP-1 BP-8
Pe
ak
are
a10 mL 20 mL 110 mL
Fig. 5. Effect of sample volume on responses obtained after 30 min of sampling, n= 3 replicates.
Fig. 6 depicts the time-course of the SPME process. As observed, the
kinetics of the extraction was rather slow and only EHS and HMS reached
equilibrium conditions after 2 hours of sampling. For practical reasons,
considering that extractions were carried out manually, 30 min was adopted as
working value. The use of longer times, to increase the sensitivity of the
method, is only advisable if the proposed method is automated.
0,0E+00
3,0E+06
6,0E+06
9,0E+06
1,2E+07
0 30 60 90 120 150 180
Time (min)
Pe
ak
are
a
EHS HMS BP-3 BP-1 BP-8/2
Fig. 6. Extraction profile using the PDMS-DVB coated fibre. Direct sampling at room
temperature using magnetic stirring (1200 rpm). Responses for BP-8 have been divided
by a factor of 2.
Analytica Chimica Acta 638 (2009) 36
191
3.3. On-fibre derivatization
Fine optimization of on-fibre derivatization conditions (temperature,
MSTFA volume, and time) was performed using a mixed mode 31 x 22
experimental factorial design. The factor temperature was studied at three
different levels (45, 60 and 75 ºC), whereas the volume of MSTFA and the
derivatization time were varied from 20 to 60 L and from 5 to 15 min,
respectively. Derivatization conditions and experimental responses obtained in
this design are summarized in Table 4. Fig. 7 shows the main effects plot for
each compound. The slope and the length of depicted lines correspond to the
variation in the peak area for each UV filter, when the associated factor changes
from the low to the high level within the domain of the design. Temperature
affected negatively to the performance of the derivatization for all compounds,
being statistically significant (95% confidence level) for EHS, HMS and BP-3.
Thus, it was fixed at 45 ºC. Time showed positive or negative effects depending
on the considered species; therefore, an intermediate value of 10 min was
chosen. Finally, the volume of MSTFA was the factor with a lower influence on
the responses of the analytes, 20 L was maintained as the working value for
this variable. Further attempts to perform the on-fibre derivatization reaction at
room temperature showed a decrease in the peak areas of silylated salicylates.
Under these conditions (room temperature, 10 min and 20 L of MSTFA), the
use of GC-MS detection demonstrated the existence of trace signals at retention
times of non-silylated EHS and HMS.
Metodología desarrollada-Muestras acuosas
192
Table 4. Working conditions and responses (peak areas) obtained during optimization of
the on-fibre derivatization step using a 31 x 22 experimental factorial design.
Exp. Temperature
(ºC)
Time
(min)
MSTFA
vol. (µL)
Response
EHS HMS BP-3 BP-1 BP-8
1 45 15 20 2002034 1719899 2549875 1680316 3670781
2 75 15 60 1722501 1525167 1812293 1667950 3473920
3 75 15 20 2061090 1864535 1950972 1472269 3496397
4 75 5 60 2162177 2022074 2045982 1517034 3932099
5 60 5 20 1773382 1597682 1811395 1344895 3488291
6 60 5 60 3378060 3213493 1763688 1391516 3345540
7 45 15 60 2390730 2139211 2906665 1832605 5201832
8 45 5 20 3207076 3073160 2204482 1634691 4087105
9 45 5 60 3500059 3346986 1978389 1442321 3677789
10 60 15 20 1956833 1838614 2477505 1972256 4802980
11 75 5 20 1597714 1425535 2063501 1678704 2738220
12 60 15 60 1769710 1587229 1969200 1400062 3659315
Analytica Chimica Acta 638 (2009) 36
193
EHS
Temp (ºC)
45 75Time (min)
5 15MSTFA (L)
20 6018
20
22
24
26
28
(x 105)
Peak
are
a
HMS
Temp (ºC)
45 75
Time (min)
5 15
MSTFA (L)
20 6017
19
21
23
25
27
(x 105)
Peak
are
a
BP-1
Temp (ºC)
45 75
Time (min)
5 15MSTFA (L)
20 6014
15
16
(x 105)
Peak
are
a
BP-8
Temp (ºC)45 75
Time (min)5 15
MSTFA (L)
20 6034
36
38
40
42(x 105)
Peak
are
a
BP-3
Temp (ºC)
45 75
Time (min)
5 15
MSTFA (L)
20 6018
20
22
24
26
(x 105)
Peak
are
a
Fig. 7. Main effect plots showing the influence of temperature, volume of MSTFA and
time on the efficiency of the on-fibre silylation reaction.
3.4. Analytical figures of merit
The linear response range of the developed method was evaluated using
standard solutions, prepared in ultrapure water, at seven different
concentration levels between 10 and 4000 ng L-1. Table 5 shows the correlation
coefficients (R2) corresponding to the representation of peak area versus
concentration. Values higher than 0.993 were obtained for all compounds.
Repeatability was assessed with samples spiked at three different
concentrations (40, 400 and 4000 ng L-1) within the calibration range. RSDs for
quadruplicate extractions remained below 13% for all species. The achieved
Metodología desarrollada-Muestras acuosas
194
limits of quantification (LOQs) ranged from 0.5 to 10 ng L-1, Table 5. In case of
benzophenones, they were controlled by the sensitivity of the GC-MS/MS
system and the efficiency of the developed sample preparation process. For the
two salicylates, the lowest quantifiable concentrations (5 ng L-1) were
determined by the signals of these species in procedural blanks. Standard
deviations of peak areas for both compounds in procedural blanks (n=5
replicates) were multiplied by 10 and divided by the slope of their calibration
curves. Achieved LOQs are similar to those reported for stir bar sorptive
extraction (SBSE), followed by thermal desorption of the PDMS coated stir bar
and GC-MS determination [18,24], and nearly three orders of magnitude lower
than LOQs previously obtained for some of these compounds using SPME
without considering the derivatization step [12,19].
Table 5. Linearity, repeatability and limits of quantification (LOQs) of the proposed
method.
Compound Correlation
coefficient (R2)
Repeatability (RSD, %), n=4 replicates LOQ
(ng L-1) Added concentration (ng L-1)
(10-4000 ng L-1) 40 400 4000
EHS
HMS
BP-3
BP-1
BP-8
0.998
0.998
0.998
0.993
0.995
2
7
7
12
12
6
1
6
10
8
7
4
8
13
9
5a
5a
0.5b
10b
2b
a Calculated as 10 times the standard deviation of their responses in procedural blanks (n=5)
divided by the slope of calibration curves.
b Defined as the concentration which produces a signal (peak area) 10 times higher than the
baseline noise.
The effect of the type of matrix on the performance of the sample
preparation method was studied with ultrapure, river, treated and raw
wastewater. After filtration, each sample was divided into two aliquots, one
was processed directly and the other was fortified with the considered
compounds (200 ng L-1). Differences between responses measured for spiked
Analytica Chimica Acta 638 (2009) 36
195
and non-spiked aliquots of each sample were compared with those
corresponding to ultrapure water with the same addition level. Normalized
values are summarized in Table 6. Relative recoveries around 100% were
noticed for all compounds in river and treated wastewater. Benzophenones also
showed relative recoveries over 80% for raw wastewater; however, the yield of
the extraction for salicylates underwent a significant reduction in this matrix.
EHS and HMS are the most lipophilic of the tested compounds; thus,
interaction with dissolved organic compounds, presented in untreated
wastewater, reduces their affinity for the PDMS-DVB coating in a higher
extension than for the more polar benzophenones.
Table 6. Relative recoveries, with their standard deviations (n=4 replicates), for
different water samples spiked at the 200 ng L-1 level.
Compound River water Treated wastewater Raw wastewater
EHS
HMS
BP-3
BP-1
BP-8
110 ± 8
109 ± 8
108 ± 13
99 ± 14
97 ± 16
95 ± 11
89 ± 10
115 ± 6
97 ± 10
108 ± 4
53 ± 5
48 ± 5
93 ± 7
92 ± 5
80 ± 5
3.5. Application to real water samples
Table 7 summarizes the levels of EHS, BP-3 and BP-1 measured in grab
samples of wastewater obtained, in different dates, from the inlet and outlet
streams of the same STP, as well as in a sample of river water. The other two
compounds considered in this study, HMS and BP-8, remained under the limits
of detection of the method in all samples. BP-3 and BP-1 were ubiquitous in the
influent of the STP reaching maximum concentrations of 460 and 245 ng L-1.
According to the information given in a recent review [17], this is one of the first
reports of the existence of BP-1 in the aquatic environment. Although grab
samples have a limited usefulness to evaluate the efficiency of treatment plants,
on the basis of concentrations measured for influents and effluents collected on
the same day, it appears that BP-3 and BP-1 are effectively removed in the
Metodología desarrollada-Muestras acuosas
196
studied STP (primary and activated sludge treatments are applied). In case of
BP-3, this conclusion is in agreement with the results obtained by Balmer et al.
for several municipal STPs in Switzerland [10]; however, in their study, the
level of BP-3 in raw wastewater reached levels as high as 8 ng mL-1. EHS was
found only in one of the processed pairs of wastewater samples at much lower
concentrations than BP-3 and BP-1. Sample 9 corresponds to the river, which
receives the effluent of the STP, as well as some non-controlled discharges of
raw wastewater. Levels of BP-3 and BP-1 in this sample were similar to those
existing in the effluent of the plant; moreover, EHS was also detected, although
not quantified, in the river water. Fig. 8 shows the chromatograms
corresponding to non-spiked and spiked (150 ng L-1) fractions of this river
water sample (code 9, Table 7), as well as a procedural blank for ultrapure
water.
Table 7. Levels of EHS, BP-3, and BP-1 in non-spiked water samples, n=3 replicates.
Code Type Sampling date Conc. (ng L-1) with their standard deviations
EHS BP-3 BP-1
1
2
3
4
5
6
7
8
9
Influent
Efluent
Influent
Efluent
Influent
Efluent
Influent
Efluent
River water
30/11/07
30/11/07
29/2/08
29/2/08
9/6/08
9/6/08
8/9/08
8/9/08
8/9/08
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
28 (1)
7.5 (0.3)
n.q.
224 (3)
n.d.
216 (27)
13 (4)
271 (22)
n.d.
462 (74)
44 (8)
52 (5)
199 (14)
n.d.
161 (11)
n.d.
131 (13)
n.d.
245 (20)
41 (2)
37 (6)
n.d.: under detection limits n.q.: under quantification limits
Analytica Chimica Acta 638 (2009) 36
197
16.50 16.75 17.00 17.25 17.50 minutes
0
50
100
150
200
kCounts
BP-3 m/z 242
15.75 16.00 16.25 16.50minutes
0
25
50
75
100
125
kCounts
EHS
HMS
m/z 177
17.3 17.4 17.5 17.6 17.7 17.8 17.9minutes
0
10
20
30
40
kCounts
BP-1 m/z 271
18.0 18.1 18.2 18.3 18.4 18.5 minutes
0
5
10
15
20
kCounts
BP-8 m/z 301+329+330
ABC
Fig. 8. Overlay of GC-MS/MS chromatograms. A, procedural blank. B, Un-spiked river
water (code 9, Table 7). C, river water spiked at 150 ng L-1.
4. Conclusions
The combination of SPME with GC-MS/MS is an interesting approach
for the sensitive determination of salicylates and hydroxylated benzophenones
in water samples. The proposed method uses only 10 mL of sample, requires a
moderate sample preparation time and provides acceptable precision (RSDs
under 13%) even for samples spiked at the low ng L-1 level; moreover, it
improves considerably the LOQs of previous methods using SPME as
Metodología desarrollada-Muestras acuosas
198
concentration technique. Significant matrix effects were only observed for two
of the considered species in raw wastewater. As far as we could trace, this work
constitutes the first application of GC-MS/MS, using electron impact ionization,
to the determination of silylated salicylates and benzophenones. Results
obtained for a limited number of municipal raw wastewater samples revealed
the ubiquity of BP-3 and BP-1 in this matrix. Thus, urban wastewater
contributes significantly to the discharge of both species in the aquatic
environment. On the other hand, both compounds (BP-3 and BP-1) seemed to
be effectively removed in the activated sludge STP.
Acknowledgments
This study has been supported by Spanish Government, Xunta de
Galicia, and E.U. FEDER funds (projects DGICT CTQ2006-03334 and
PGIDIT06PXIB237039PR). N. N and M. R. thank the Spanish Ministry of Science
and Innovation and Xunta de Galicia for a FPU grant and a Parga Pondal
contract, respectively.
Analytica Chimica Acta 638 (2009) 36
199
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1.6. Publicación:
SOLID-PHASE EXTRACTION
FOLLOWED BY LIQUID CHROMATOGRAPHY
TANDEM MASS SPECTROMETRY
FOR THE DETERMINATION OF
HIDROXILATED BENZOPHENONE UV ABSORBERS
IN ENVIRONMENTAL WATER SAMPLES
N. Negreira, I. Rodríguez, M. Ramil, E. Rubí, R. Cela
Analytica Chimica Acta 654 (2009) 16
(doi:10.1016/j.aca.2009.09.033)
Analytica Chimica Acta 654 (2009) 16
203
Solid-phase extraction followed by liquid chromatography tandem mass
spectrometry for the determination of hydroxylated benzophenone UV
absorbers in environmental water samples
N. Negreira, I. Rodríguez, M. Ramil*, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto de
Investigación y Análisis Alimentario, Universidad de Santiago de Compostela,
Santiago de Compostela 15782, Spain.
Abstract
A procedure for the determination of six derivatives of 2-
hydroxybenzophenone, used as UV absorbers, in water samples is presented.
Compounds were first concentrated using a solid-phase extraction (SPE)
cartridge and then selectively determined by liquid chromatography tandem
mass spectrometry (LC-MS/MS) using electrospray ionization (ESI). The effect
of different parameters on the performance of concentration and determination
steps is discussed. The highly polar and acidic 2-hydroxy-4-
methoxybenzophenone 5-sulfonic acid (BP-4) required the use of ammonium
acetate as modifier during desorption of SPE cartridges and also to improve the
performance of its separation in the LC column. Under optimised conditions,
the proposed method provided limits of quantification from less than 1 to 32 ng
L-1, depending on the compound and the type of water sample. Recoveries from
the SPE step (83-105%) remained unaffected by the nature of the matrix;
however, the efficiency of electospray ionization was compound and sample
dependant. Real sample analysis reflected the presence of three of the six
investigated species (BP-4; 2-hydroxy-4-methoxybenzophenone, BP-3, and 2,4-
dihydroxybenzophenone, BP-1) in the aquatic environment, particularly in raw
wastewater samples. In this latter matrix, BP-4 was the compound presenting
the highest concentrations; moreover, it was poorly removed in sewage
treatment plants and consequently it also appeared in river water.
Metodología desarrollada-Muestras acuosas
204
Keywords: Hydroxylated benzophenones, UV absorbers, liquid
chromatography tandem mass spectrometry, water analysis.
1. Introduction
Derivatives of 2-hydroxybenzophenone are extensively employed as UV
absorbers. In the European Union (EU), 2-hydroxy-4-methoxybenzophenone-5-
sulphonic acid (BP-4) and 2-hydroxy-4-methoxybenzophenone (BP-3) have
been approved to be used as UV filters in sunscreens at maximum individual
concentrations of 5 and 10%, respectively [1]. Other countries, e.g. Japan, also
allow the incorporation of 2,4-dihydroxybenzophenone (BP-1), 2,2´,4,4´-
tetrahydroxybenzophenone (BP-2) and 2,2´-dihydroxy-4,4´-
methoxybenzophenone (BP-6) in sunscreens [2]. In addition, the above
compounds, as well as other related benzophenones such as 2,2´-dihydroxy-4-
methoxybenzophenone (BP-8), are included as photostabilizers in many other
personal care products (e.g. hair dyes and shampoos), varnishes, clothes and
food container plastics [3,4]. BP-1 is also the main metabolite of BP-3, identified
in human urine after topical application of sunscreen products containing the
latter compound [5,6].
The increasing usage of hydroxylated benzophenones, combined with
their moderate to high water solubility, has led to the appearance of some of
these compounds in the aquatic environment. BP-3 has been often detected in
different water samples, from recreational areas (e.g. swimming pool and
bathing waters) to surface water and municipal wastewater [7-10] and even in
fish tissues, which suggests that it is bio-accumulated in the food chain [11,12].
More recently, BP-4 [13,14] and BP-1 [15] have been also found in water
samples and the first species seems to be able to pass through sewage treatment
plants (STPs) without being significantly removed [13,14,16]. On the other
hand, the levels and fate of BP-2, BP-6 and BP-8 in the aquatic environment
remain mostly unknown. As regards toxicological effects, in-vivo and in-vitro
studies have demonstrated that BP-1, BP-2, BP-3 and BP-4 exert estrogenic and
anti-androgenic actions and they might affect the reproduction of fish [17-20].
Analytica Chimica Acta 654 (2009) 16
205
On the other hand, BP-8 is considered as a genotoxic compound [6]. Globally,
the above data raises the concerns about medium and long-term environmental
effects of hydroxylated benzophenones.
Until recently, gas chromatography in combination with mass
spectrometry (GC-MS) was the most common technique for the determination
of benzophenone-type UV absorbers in water and other environmental matrices
[21]. When combined with effective sample preparation techniques, such as
liquid-liquid extraction (LLE) [4,9], solid-phase extraction (SPE) [8,11], solid-
phase microextraction (SPME) [15] or stir-bar sorptive extraction (SBSE) [22,23],
it provides limits of detection in the low ng L-1 range, particularly, if acetic
anhydride [23] or a silylation reagent [4,15,24] are used to derivatize the
phenolic groups contained in the structure of target benzophenones.
Derivatization is particularly important for sensitive detection of those
compounds with two or more phenolic moieties, such as BP-1, BP-2, BP-6 and
BP-8. Unfortunately, BP-4, which contains a sulfonic group attached the
aromatic ring, in addition to the phenolic one, is not amenable to GC analysis.
The above limitations can be overcome using liquid chromatography (LC)-
based methods. BP-3 and BP-4 have been often included in multi-residue LC-
MS/MS methods focused on the determination of personal care products [16]
and multi-class UV filters [13,25] in water samples; however, the performance
of LC-MS/MS for the determination of other hydroxyl benzophenones in water
has received less attention. Although MS in combination with LC has been
recently applied to a broader group of benzophenones [14], this work was
devoted to biological samples and it did not deal either with the concentration
of water samples, or with the use of tandem MS.
Thus, the aim of this research is to evaluate the efficiency of LC-MS/MS,
using electrospray ionization (ESI), for the determination of 6 hydroxylated
benzophenones: BP-1 to BP-4, BP-6 and BP-8 in water samples and to assess
their levels and fate in this matrix, with special attention focused on
wastewater. Samples were concentrated using a reversed-phase SPE sorbent in
Metodología desarrollada-Muestras acuosas
206
order to improve the sensitivity of the method. The influence of different
parameters on the performance of sample preparation and determination steps
is thoroughly evaluated. Data related to the levels of target compounds in
different environmental water samples are also provided.
2. Experimental
2.1. Standards, solvents and material
Standards of BP-1, BP-2, BP-3, BP-4, BP-6 and BP-8 were purchased from
Aldrich (Milwaukee, WI, USA) and Riedel de Haën (Seelze, Germany). The
chemical structures of these compounds and some properties of relevance to
optimize extraction (SPE) and LC separation steps are given in Table 1. Formic
and hydrochloric acid, ammonium acetate and HPLC-grade methanol were
acquired from Merck (Darmstadt, Germany). Ultrapure water was obtained
from a Milli-Q system (Millipore, Billerica, MA, USA). Stock solutions of each
compound were prepared in methanol. Further dilutions and mixtures of them
were made in the same solvent. Standards in methanol were stored in the dark
at 4 ºC for a maximum of two months. Calibration standards in a
methanol:water (1:1), containing a 5 mM concentration of ammonium acetate,
were prepared when needed, at different concentrations in the range from 5 to
1000 ng mL-1.
SPE cartridges, containing 60 mg of the OASIS HLB sorbent, were
provided by Waters (Milford, MA, USA).
Analytica Chimica Acta 654 (2009) 16
207
Table 1. Abbreviated names, CAS number, structures, pKa and octanol-water partition
coefficients of selected compounds.
Abbreviated
name
CAS
number
Molecular
weight (a.m.u.) Structure apKa aLog Kow
BP-1 131-56-6 214.2
7.53 3.17
BP-2 131-55-5 246.2
6.98 3.16
BP-3 131-57-7 228.2
7.56 3.79
BP-4 4065-45-6 308.3
-0.70 0.89
BP-6 131-54-4 274.3
6.80 4.10
BP-8 131-53-3 244.2
6.99 3.93
a Values obtained from SciFinder Scholar Database, http://www.cas.org/products/sfacad/
2.2. Samples and sample preparation
Ultrapure (Milli-Q), river and wastewater samples, obtained from the
inlet and outlet streams of an urban STP, equipped with primary and activated
sludge units, were employed throughout this study. River and wastewater
samples were stored in the dark, at 4º C, for a maximum of 72 hours until being
concentrated by SPE. Under optimised sample preparation conditions, from 200
to 500 mL of acidified water (adjusted to pH 2 using HCl 0.1 M) were filtered,
using glass fibre and cellulose acetate (0.45 m pore size) filters, and passed
through an OASIS HLB cartridge (c.a. 10 mL min-1), previously conditioned
Metodología desarrollada-Muestras acuosas
208
with methanol and ultrapure water (3 mL each) at the same pH as water
samples. After that, the beaker containing the sample and connections with the
SPE cartridge were rinsed with 20 mL of ultrapure water. Finally, cartridges
were dried using a stream of nitrogen and analytes eluted with 3 mL of
methanol, containing a 5 mM concentration of ammonium acetate. This extract
was evaporated to dryness and reconstituted with 1 mL of a methanol:water
(1:1) solution, 5 mM in ammonium acetate.
Performance of the SPE process was assessed by spiking the acidified
samples after the filtration step. Unless otherwise stated, 10 ng mL-1 was used
as the addition level during optimisation of SPE conditions.
2.3. Equipment
Analytes were determined using a Varian (Walnut Creek, CA, USA) LC-
MS/MS system. The LC instrument comprised two isocratic, high-pressure
mixing pumps (Varian ProStar 210), an autosampler and a thermostated
compartment for the column (Varian ProStar 410). The mass spectrometer was
a U-shaped triple quadrupole (Varian MS 1200L) furnished with an
electrospray ionization (ESI) interface. The LC-MS/MS instrument was entirely
controlled by the Varian MS Workstation Version 6.9 software.
Compounds were separated using a Kromasil C18 column (100 mm x 2.1
mm; 5 m) acquired from Sugelabor (Madrid, SPAIN) and connected to a C18
(4 mm x 2 mm) guard cartridge supplied by Phenomenex (Torrance, CA, USA).
Ultrapure water (A) and methanol (B), containing different amounts of formic
acid or ammonium acetate as modifiers, were employed as mobile phases.
Under final conditions, a 5 mM concentration of the latter modifier was added
to both solvents and compounds were separated using the following gradient:
0-2 min, 15% B; 6 min, 40% B; 10 min, 70% B; 18 min, 75% B; 21-24 min, 100% B;
27-36 min, 15% B. The mobile phase flow was set at 0.2 mL min-1 and the
temperature of the column fixed at 35 ºC. Injection volume for standards and
sample extracts was 15 L.
Analytica Chimica Acta 654 (2009) 16
209
Nitrogen (99.999%), used as nebulising (50 PSI) and drying gas (300 ºC,
21 PSI) in the ESI source, was provided by a high purity generator (Domnick
Hunter, Durham, UK). The temperature of the ESI housing was maintained at
50 ºC and the voltage of the ESI needle fixed at 5000 V. Argon (99.999%) was
employed as collision gas (1.5 mTorr) in the mass spectrometer. Benzophenones
were recorded in the multiple reaction monitoring (MRM) mode, using two
transitions per compound. The most intense one was used to quantify the
response of each species in standards and SPE extracts from real water samples.
Under final working conditions, the six benzophenones were grouped in four
segments according to their elution order. Table 2 summarizes retention times,
most intense MS/MS transitions, ionization modes (positive or negative),
capillary voltages and collision energies for target species.
Table 2. Retention times and optimized ESI-MS/MS conditions. CV (capillary voltage,
V); CE (collision energy, eV).
Compound Ret. time
(min)
ESI Segment MRM1
(quantification)
CV/CE MRM2
(confirmation)
CV/CE
BP-4
BP-2
BP-1
BP-8
BP-6
BP-3
10.5
11.8
14.1
15.6
18.2
19.5
-
-
-
-
-
+
1
2
3
3
4
4
307>210
245>135
213>91
243>93
273>123
229>151
80/34
40/14
60/29
30/17
40/20
52/17
307>227
245>109
213>135
243>123
273>108
229>105
80/26
40/22
60/19
30/16
40/31
52/18
Metodología desarrollada-Muestras acuosas
210
3. Results and discussion
3.1. Optimization of LC-MS/MS parameters
3.1.1. MS/MS transitions
MS/MS fragmentation conditions were optimized by infusion of
individual standards (10 g mL-1 in methanol:water, 1:1) at a constant flow of 50
L min-1, operating the electrospray source in the positive (ESI+) and negative
(ESI-) modes. Except for BP-4, which could be ionized only in negative mode,
the rest of compounds yielded the corresponding protonated ([M+H]+)and
deprotonated ([M-H]-) parent ions in ESI+ and ESI-, respectively. Using MS/MS
detection, [M+H]+ ions underwent the cleavage of the bond between the
carbonyl group and one aromatic ring, with the positive charge remaining in
the fragment attached to the carbonyl moiety. In case of BP-3, the most intense
products appeared at 151 and 105 m/z units, Fig. 1A. The same bond was also
broken during fragmentation of [M-H]- precursors. For those compounds
containing just a hydroxyl substituent over each phenolic ring, e.g. BP-3, BP-6
and BP-8, the carbonyl moiety did not remain attached to MS/MS phenolate
product ions, Fig. 1B. Fig. 1C shows the MS/MS fragmentation pattern of those
benzophenones with two hydroxyl moieties in the same aromatic ring (BP-1
and BP-2), using BP-1 as model compound. In addition to the formation of
dihydroxylated phenolic ions, appearing at a m/z of 109 units, the carbonyl
moiety may also remain attached to the negative charged product ion (m/z
135), being further removed together with one atom of oxygen as CO2
(transition 135 > 91 m/z). Finally, the most intense ions in the MS/MS spectrum
of BP-4 maintained the benzophenone structure and corresponded to the loss of
sulfonic (307>227) and sulfonic plus hydroxyl moieties (307>210), Fig. 1D.
Other transitions, e.g. 307>80 and 307>211, previously reported in the literature
[13,16], were also observed; however, their intensities were lower than those
corresponding to the former ones.
Analytica Chimica Acta 654 (2009) 16
211
m/z 229
m/z 105m/z 151
[M+H]+ A
D
C
[M-H]-
m/z 307 m/z 227
m/z 210
O O
MeO
m/z 91
m/z 213
m/z 135
m/z 109
[M-H]-
B
- 44
(CO2)
[M-H]-m/z 243
m/z 93 m/z 123
Fig. 1. Proposed MS/MS fragmentation mechanisms. A, BP-3 (ESI+). B, BP-8 (ESI-).
C, BP-1 (ESI-). D, BP-4 (ESI-).
Metodología desarrollada-Muestras acuosas
212
Flow injection analysis (FIA) was used to investigate the effect of the
ionization mode on the intensity of MS/MS responses for each compound. ESI-
was preferred for all analytes, except for BP-3, which gave an increased
response (three orders of magnitude) in positive mode.
3.1.2. Mobile phase modifiers
Addition of formic acid to the mobile phase (concentrations up to 0.1%
were evaluated) produced a reduction in the responses measured for BP-1, BP-
2, BP-6 and BP-8. These compounds present pKa values from 6.8 to 7.5 units
(Table 1), thus formic acid shifts their acid-base equilibrium towards the neutral
forms, leading to a diminution of their responses in ESI-. On the other hand,
signals recorded for BP-3 in ESI+, and the strong acidic BP-4 remained
unchanged in presence of formic acid. Ammonium acetate, at increased levels
from 0 to 10 mM, produced also a slight reduction in the MRM responses of all
compounds. However, when included in the LC mobile phase, it improved the
chromatographic behavior of BP-4. Without this salt, using just methanol and
water as mobile phases, the retention of this highly polar compound (log Kow
0.89) was low and two peaks with same MS/MS spectra were observed. Adding
a 5 mM of ammonium acetate to the mobile phase increased the retention time
of BP-4 around 5 min, leading to a single and symmetrical peak for this
compound.
3.1.3. Instrumental performance
Using the gradient given in the experimental section, selected
compounds were separated in about 20 min, Fig. 2. In order to improve the
achieved limits of quantification (LOQs), they were grouped into four different
segments according to their retention times. In the latter one, negative and
positive ionization modes were combined to record BP-6 and BP-3 under
optimal conditions. The total dwell time per segment was maintained at 1.2 s.
Analytica Chimica Acta 654 (2009) 16
213
Other MS/MS detection conditions, as well as retention times of selected
benzophenones are compiled in Table 2.
10.0 12.5 15.0 17.5 20.0Min.
0
20
40
kCounts
0.0
0.5
1.0
1.5
MCounts
0.50
1.00
1.50MCounts
0
100
200
kCounts
0
100
200
300
kCounts
0
20
40
kCounts
BP-4307 > 210
BP-2245 > 135
BP-1213 > 91
BP-8243 > 93
BP-6273 > 123
BP-3229 > 151
0.0
Fig. 2. LC-MS/MS chromatogram for a standard of benzophenones (25 ng mL-1 per
compound).
Relevant data related to the performance of the optimized LC-MS/MS
method are summarized in Table 3. The dependence between peak areas and
analytes concentration was investigated with standards (injection volume 15
Metodología desarrollada-Muestras acuosas
214
L) at 7 different concentrations in the range from 5 to 1000 ng mL-1. BP-3, BP-4,
BP-6 and BP-8 gave a linear response in the above range, with correlation
coefficients (R2) between 0.998 and 0.999; whereas, BP-1 and BP-2 fitted better a
quadratic plot with R2 values of 0.999 and 0.997, respectively. Instrumental
limits of quantification, defined for a signal to noise ratio of 10 (S/N=10) varied
from 0.2 for BP-4 up to 4 ng mL-1 in case of BP-3 and BP-8, Table 3. The
repeatability in the responses of the system was evaluated with standards at
two different concentrations: 25 and 125 ng mL-1. Relative standard deviations
(RSDs, %) for 5 injections made in the same day ranged from 1 to 6%, Table 3.
For injections performed on consecutive days RSDs between 3 and 14% were
obtained.
Table 3. Linearity, instrumental limits of quantification (LOQs), repeatability and
reproducibility of the LC-MS/MS system.
Compound Correlation
coefficient (R2)
LOQs
(ng mL-1)
aRepeatability (RSD, %) bReproducibility (RSD, %)
125 ng mL-1 25 ng mL-1 125 ng mL-1
BP-4
BP-2
BP-1
BP-8
BP-6
BP-3
0.998
c0.997
c0.999
0.999
0.999
0.998
0.2
0.4
3
4
3
4
6.1
4.0
3.2
4.0
5.0
3.5
3.7
3.4
2.4
1.1
3.5
2.0
3.4
3.1
4.5
8.3
13.5
2.6
aN=5 injections in the same day
bN=12 injections in 3 consecutive days
cQuadratic model
3.2. Solid-phase extraction conditions
The optimisation of SPE conditions was performed with aliquots of
ultrapure water spiked at 10 ng mL-1 and passed through 2 cartridges
connected in series. After that, sample containers and connections with the
sorbents were rinsed with ultrapure water, ca. 20 mL. Cartridges were then
Analytica Chimica Acta 654 (2009) 16
215
disassembled, dried and eluted separately with different volumes (from 1 to 10
mL) of methanol or methanol containing a 5 mM concentration of ammonium
acetate. Extracts in methanol were diluted (1:1) with ultrapure water 5 mM in
ammonium acetate before injection in the LC-MS/MS system. The pH of the
sample showed a strong effect on the breakthrough volume of BP-4. When
samples were acidified at pH 2, up to 1000 mL of water could be concentrated
without noticing the presence of any compound in the extract from the second
cartridge. At neutral pH values (between 6 and 8 units), BP-4 showed a
breakthrough volume around 200 mL, whereas the rest of species were still
quantitatively retained in the first SPE cartridge. In the elution step,
hydroxylated benzophenones could be recovered with 3 mL of methanol;
however, BP-4 presented a lower affinity for this solvent, requiring much larger
volumes of eluent. This problem was overcome using methanol containing
ammonium acetate (5 mM). All compounds were recovered with 3 mL of this
mixture, a volume significantly lower than those reported in previous works
dealing with SPE of BP-4 and BP-3 [13, 25]. It is probable that the ammonium
ions neutralize the negative charge of the sulfonic group, increasing the
solubility of BP-4 in methanol.
In a further series of assays, performed with samples spiked at a lower
concentration level (ca. 1 ng mL-1), extracts from SPE cartridges were
concentrated to dryness and reconstituted with 1 mL of the same solution used
to prepare calibration standards. Losses of the analytes were not detected when
evaporation was carried out at room temperature with a gentle stream of
nitrogen. Table 4 shows the absolute recoveries, estimated against external
calibration, for 500 and 1000 mL aliquots of spiked ultrapure water (1 ng mL-1)
using above conditions. Average recoveries from 87 to 103%, with acceptable
precision, were attained for all compounds. These values are similar to the
recoveries obtained for BP-3 and other UV filters in previous works [7,11] using
different SPE sorbents.
Metodología desarrollada-Muestras acuosas
216
Table 4. Extraction yields of the solid-phase extraction process for ultrapure water
samples. Added concentration 1 ng mL-1, n=3 replicates.
Compound Extraction yield (%) ± SD
a500 mL a1000 mL
BP-4
BP-2
BP-1
BP-8
BP-6
BP-3
87 ± 1
96 ± 1
96 ± 6
90 ± 8
87 ± 9
103 ± 6
92 ± 1
87 ± 6
99 ± 3
91 ± 4
89 ± 3
97 ± 7
aVolume of sample
3.3. Matrix effects and figures of merit
The efficiency of the ESI interface is prone to changes depending on the
complexity of the sample. Matrix effects for the proposed method were assessed
comparing the differences between responses obtained for spiked and non-
spiked extracts from different water samples (extracts were fortified after SPE at
the 500 ng mL-1 level) with those measured for a standard of the same
concentration, prepared in methanol: water (1:1), both 5 mM in ammonium
acetate. Fig. 3 shows the results obtained for river (500 mL), treated (300 mL)
and raw (200 mL) wastewater samples. In case of river and treated wastewater,
BP-2 was the only compound undergoing strong signal suppression (up to 80%
in the latter matrix); whereas, responses measured for the rest of hydroxylated
benzophenones represented between 75 and 113% of those corresponding to the
reference standard. The response of BP-4 experienced a moderate enhancement
(around 30%) for the above matrices, similar to previous data published by
Rodil and co-workers [13]. Probably, the elution of polar compounds at short
retention times impaired the ionization of BP-2 in the negative mode. Although
BP-4 elutes even earlier, the high acidity of the sulfonic group, led to a more
robust ionization ESI-. As expected, for raw wastewater the signal suppression
for BP-2 was even more acute. The reduction in the efficiency of the ionization
for the rest of species ranged from 25 to 50% in this matrix; however, no
Analytica Chimica Acta 654 (2009) 16
217
problems were observed for BP-4, Fig. 3. Signal suppression effects for BP-3 and
BP-4 in raw wastewater samples were less significant than those reported for
the same sample intake in a recent work [25]. Differences between (1) employed
SPE conditions, (2) selected MS/MS transitions for BP-4 and (3) configurations
of the ESI interface in LC-MS/MS systems from different suppliers might
contribute to these findings. On the basis of data depicted in Fig. 3, and
considering the lack of isotopic labeled analogous of target compounds, the
standard addition method is recommended to assess the levels of hydroxylated
benzophenones in raw wastewater. For the rest of matrices, this quantification
approach is only mandatory for the accurate determination of BP-2. Anyhow,
matrix effects have to be systematically checked since they might change among
different samples, even of the same type (e.g. surface of treated wastewater),
depending on the levels of dissolved salts and organic compounds.
0%
20%
40%
60%
80%
100%
120%
140%
BP-4 BP-2 BP-1 BP-8 BP-6 BP-3
No
rma
lize
d r
esp
on
se
River Treated wastewater Raw wastewater
Fig. 3. Assessment of matrix effects for river (500 mL), treated (300 mL) and raw
wastewater samples (200 mL). Normalized responses to a standard of same
concentration, n=4 replicates.
Table 5 shows the recoveries of the method for river, treated and raw
wastewater after correcting the responses measured for the final extract from
the SPE cartridge with above referred matrix effects. Each sample was divided
into several fractions, some of them were processed directly and the others
Metodología desarrollada-Muestras acuosas
218
fortified with different concentrations of target compounds in the range from
0.15 to 2.0 ng mL-1, depending on the matrix being investigated. Corrected
recoveries ranged from 83 to 105%, indicating that the efficiency of the SPE
process was not comprised by the type of water sample. Despite some
researchers have reported cross contamination problems for UV filters [26],
procedural blanks, corresponding to the concentration of ultrapure water
samples, did not reveal noticeable responses for target species. The lower log
Kow values of hydroxylated benzophenones (from 0.9 to 4.1), in comparison
with other UV filters such as camphors and cinnamates (e.g. octocrylene log
Kow 7.5), may explain the absence of carryover problems due to analytes
adsorption on glass material and connectors of the SPE sample preparation
station. Anyhow, the use of gloves during manipulation of water samples and
SPE extracts is mandatory in order to prevent contamination from personal care
products used by operators. Taking these results into account, the limits of
quantification (LOQs) of the proposed method, defined for a signal to noise
(S/N) of 10, are controlled by the sensitivity of the LC-MS/MS system, the
concentration factor provided by the sample preparation method: 500, 300 and
200-fold for river, treated and raw wastewater, respectively and signal
suppression effects depicted in Fig. 3. Estimated values ranged from less than 1
ng L-1 up to 32 ng L-1, depending on the compound and the type of water
sample. In case of BP-4 and BP-3, LOQs achieved in this work are significantly
lower than those reported recently by Kasprzyk-Hordern and co-workers in
wastewater samples: 10 and 80 ng L-1 for BP-4 and BP-3, respectively [16]. The
detection of BP-3 in the negative ionization mode probably contributed to the
limited sensitivity of their method for this compound [16]. LOQs reported on
Table 5 are also similar to those achieved combining SPME with GC-MS/MS
detection after on-fiber derivatization of target species [15]; however, this latter
procedure cannot be applied to BP-4.
Analytica Chimica Acta 654 (2009) 16
219
T
able
5. C
orre
cted
rec
over
ies
(n=
4 r
eplic
ates
) and
est
imat
ed L
OQ
s fo
r th
e pr
opos
ed m
etho
d.
R
iver
(500
mL
) T
reat
ed w
aste
wat
er (3
00 m
L)
Raw
was
tew
ater
(200
mL
)
Com
poun
d
Rec
over
y (%
) S
D
LO
Q (n
g L
-1)
Rec
over
y (%
) S
D
LO
Q (n
g L
-1)
Rec
over
y (%
) S
D
LO
Q (n
g L
-1)
a 0.
15 n
g m
L-1
a 0
.4 n
g m
L-1
a 2.0
ng
mL
-1
a 2
.0 n
g m
L-1
BP
-4
BP
-2
BP
-1
BP
-8
BP
-6
BP
-3
100
± 6
101
± 5
105
± 3
104
± 5
103
± 3
98 ±
6
90 ±
5
91 ±
5
96 ±
3
95 ±
3
93 ±
4
84 ±
4
0.4
b 2 6 8 6 8
104
± 2
97 ±
4
93 ±
2
91 ±
2
91 ±
2
96 ±
2
0.7
b 14
10
13
10
13
83 ±
6
96 ±
4
101
± 2
98 ±
3
90 ±
4
87 ±
5
1
b 20
b 30
b 32
b 25
b 32
a Ad
ded
con
cent
rati
on
b Rep
orte
d v
alu
es h
ave
been
cor
rect
ed w
ith
mat
rix
effe
cts
dep
icte
d in
Fig
. 3
Metodología desarrollada-Muestras acuosas
220
3.4. Real samples analysis
The developed method was applied to the analysis of several river and
wastewater samples taken during the period from June to September of 2008.
This season corresponds to the highest consumption of sunscreen products,
urban wastewater is not diluted by rain and rivers show their lowest flow rates.
Thus, these samples are representative of a worst-case scenario for the
investigated geographical area. The resultant concentrations for BP-1, BP-3 and
BP-4 are shown in Table 6, the rest of species remained below the detection
limits (LODs) of the method. External calibration and standard addition over
sample extracts, before dryness evaporation and re-constitution to 1 mL, were
used to quantify their levels in river and wastewater, respectively. All
wastewater samples (codes 1-8, Table 6) corresponded to the same STP,
equipped with primary and secondary treatments, which receives municipal
wastewater from a 125000 inhabitants, non-coastal city. Grab samples, without
considering the residence time of the STP, were obtained from the inlet and
outlet of the plant. BP-1, BP-3 and BP-4 were found in the four influents with
concentrations increasing in the above order from 30 ng L-1 for BP-1 up to 1600
ng L-1 for BP-4. This trend agrees with levels found in STPs from Wales [16].
Fig. 4 shows the LC-MS/MS chromatograms for spiked and non-spiked
aliquots of a raw wastewater (code 1, Table 6), as well as a procedural blank of
ultrapure water. Dissolved concentrations of BP-1 and BP-3 were considerably
reduced in the effluent of the plant; however, its efficiency was poor for BP-4,
Table 6. Although this conclusion needs to be confirmed with integrated
samples, the resistance of BP-4 to degradation agrees with (1) its high polarity,
(2) recent results published during last year for other STPs [13, 16], and (3) the
reports of this species in surface water obtained with passive sampling [14]. BP-
4 was also present in the four considered river samples (codes 9-12, Table 6).
Samples 9 and 10 were collected in the same river, five Km downstream from
the STP, in two different months. In addition to BP-4, they showed levels of BP-
1 (only sample 9) and BP-3 similar to those measured in the effluent of the STP
(codes 6 and 8). Codes 11 and 12 corresponded to small rivers flowing through
Analytica Chimica Acta 654 (2009) 162
221
urban areas, which do not receive direct discharges of STPs and without
recreational bathing areas. In spite of this, BP-4 was over the quantification
limits of the method in both samples.
Table 6. Concentrations (ng L-1) of BP-4, BP-1 and BP-3 in environmental water
samples, n= 3 replicate extractions.
Code Type Sampling data Concentration, ng L-1,
(standard deviation)
BP-4 BP-1 BP-3
1
2
3
4
5
6
7
8
9
10
11
12
Raw wastewater
Treated wastewater
Raw wastewater
Treated wastewater
Raw wastewater
Treated wastewater
Raw wastewater
Treated wastewater
River
River
River
River
June 2008
June 2008
July 2008
July 2008
August 2008
August 2008
September 2008
September 2008
August 2008
September 2008
July 2008
September 2008
1293 (74)
821 (84)
1237 (10)
1028 (61)
1354 (147)
890 (97)
1596 (36)
765 (22)
416 (45)
283 (16)
20 (5)
147 (3)
40 (3)
< LOD
148 (7)
13 (2)
120 (4)
11 (1)
31 (2)
< LOD
24 (1)
< LOD
< LOD
< LOD
184 (8)
< LOD
317 (24)
83 (12)
429 (23)
77 (4)
369 (12)
84 (3)
54 (3)
87 (8)
< LOD
< LOD
Metodología desarrollada-Muestras acuosas
222
9.5 10.0 10.5 11.0 min.
0
100
200
300
400
500
600
(x103 Counts)
BP-4
16 17 18 19 20.
0.0
0.5
1.0
1.5
2.0
(x106 Counts)
BP-6
13.0 14.0 15.0 min,
0.0
0.5
1.0
1.5
2.0
2.5
(x106 Counts)
BP-1
11.0 12.0 Min.0
100
200
300
400
500
600
700
(x103 Counts)
BP-2
16 17 18 19 20 min.
0
100
200
300
400
500
600
700
(x103 Counts)
BP-3
14.0 15.0 16.0 min.0.00
0.25
0.50
0.75
1.00
1.25
(x106 Counts)
BP-8
CBA
Fig. 4. Overlay of LC-MS/MS chromatograms. A, procedural blank for 500 mL of
ultrapure water. B, raw wastewater (code 1, Table 6). C, same sample spiked at 2 ng
mL-1.
Analytica Chimica Acta 654 (2009) 162
223
4. Conclusions
LC-MS/MS combined with an SPE pre-concentration step allows the
sensitive determination of six hydroxylated benzophenones in water samples
avoiding the need of analytes derivatization, as required in GC-based methods.
The differences in polarity and acidity of BP-4 when compared with the rest of
analytes required a careful tuning of SPE and LC conditions, using an ion pair
reagent to reduce the elution volume for this compound in the SPE step and to
improve its chromatographic behavior. Conversely to the rest of species, BP-3 is
more efficiently ionized in the positive than in the negative ESI mode, thus both
ionization modes must be combined in the same chromatographic run. Under
optimized conditions, the proposed method provided enough sensitivity for the
determination of the six considered compounds in environmental water
samples. With the exception of BP-2, the rest of compounds were affected by
strong matrix suppression effects just in raw wastewater, but not in river and
treated sewage samples. Analysis of real samples confirmed the ubiquitous
presence of BP-4 in the aquatic environment, as well as its recalcitrant behavior
in STPs. This behavior added to its high polarity turns BP-4 in a mobile
compound in the aquatic media, with the potential risk to reach potable water
sources. BP-3 and BP-1 were also detected in waste and river samples;
however, they seemed to be removed to a considerable extent during
conventional wastewater treatments.
Acknowledgments
Financial support from the Spanish Government, Xunta de Galicia and
FEDER funds (projects DGICT CTQ2006-03334 and PGIDIT06PXIB237039PR) is
acknowledged. N. N. and M. R. thank the Spanish Ministry of Sciene and the
Xunta de Galicia for a FPU grant and an I. Parga Pondal contract, respectively.
Metodología desarrollada-Muestras acuosas
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[2] N.A. Shaath, The Encyclopedia of Ultraviolet Filters, Allured, Illinois, 2007.
[3] S.D. Richardson, Anal. Chem. 80 (2008) 4373.
[4] H.K. Jeon, Y. Chung, J. C. Ryu, J. Chromatogr. A 1131 (2006).
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[6] M. S. Díaz-Cruz, M. Llorca, D. Barceló, Trends Anal. Chem. 27 (2008) 873.
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[8] W. Li, Y. Ma, C. Guo, Y. Hu, K. Liu, Y. Wang, T. Zhu, Water Res. 41 (2007) 3506.
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[10] D. L. Giokas, A. Salvador, A. Chisvert, Trends Anal. Chem. 26 (2005) 360.
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[12] M.A. Mottaleb, S. Usenko, J.G. O´Donnell, A.J. Ramirez, B.W. Brooks, C.K. Chambliss, J.
Chromatogr. A 1216 (2009) 815.
[13] R. Rodil, J.B. Quintana, P. López-Mahía, S. Muniategui-Lorenzo, D. Prada-Rodríguez, Anal.
Chem. 80 (2008) 1307.
[14] A. Zenker, H. Schmutz, K. Fent, J. Chromatogr. A 1202 (2008) 64.
[15] N. Negreira, I. Rodríguez, M. Ramil, E. Rubi, R. Cela, Anal. Chim. Acta 638 (2009) 36.
[16] B. Kasprzyk-Hordern, R.M. Dinsdale, A.J. Guwy, Anal. Bional. Chem. 391 (2008) 1293.
[17] P.Y. Kunz, H.F. Galicia, K. Fent, Toxicol. Sci. 90 (2006) 349.
[18] K. Fent, P.Y. Kunz, E. Gomez, Chimia 62 (2008) 368.
[19] C.J. Weisbrod, P.Y. Kunz, A.K. Zenker, K. Fent, Toxicol. Appl. Pharmacol. 225 (2007) 255.
[20] M. Heneweer, M. Muusse, M. van den Berg, J.T. Sanderson, Toxicol. Appl. Pharmacol. 208
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[21] T. Felix, B.J. Hall, J.S. Brodbelt, Anal. Chim. Acta 371 (1998) 195.
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Chromatogr. A 1200 (2008) 260.
[24] K.W. Ro, J.B. Choi, M.H. Lee, J.W. Kim, J. Chromatogr. A 688 (1994) 375.
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1.7. Publicación:
STUDY OF SOME UV FILTERS STABILITY
IN CHLORINATED WATER AND IDENTIFICATION
OF HALOGENATED BY-PRODUCTS BY
GAS CHROMATOGRAPHY-MASS SPECTROMETRY
N. Negreira, P. Canosa, I. Rodríguez, M. Ramil, E. Rubí, R. Cela
Journal of Chromatography A 1178 (2008) 206
(doi:10.1016/j.chroma.2007.11.057)
Journal of Chromatography A 1178 (2008) 206
227
Study of some UV filters stability in chlorinated water and identification of
halogenated by-products by gas chromatography- mass spectrometry
N. Negreira, P. Canosa, I. Rodríguez*, M. Ramil, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto de
Investigación y Análisis Alimentario, Universidad de Santiago de Compostela,
Santiago de Compostela 15782, Spain.
Abstract
This work studies the stability of three UV filters: 2-ethylhexyl salicylate
(ES), 2-ethylhexyl 4-(dimethylamino) benzoate (EHPABA) and 2-hydroxy-4-
methoxybenzophenone (BP-3), in water samples containing low concentrations
of free chlorine. Moreover, 2,4-dihydroxybenzophenone (2,4-DHBP), a
metabolite of BP-3, was also included in some of the performed assays.
Experiments were carried out considering free chlorine and analytes
concentrations at the g mL-1 and ng mL-1 level, respectively. Gas
chromatography with mass spectrometry was used to follow the time course of
target compounds and to identify their halogenated by-products. Concentration
of water samples with solid-phase extraction cartridges and derivatization
(silylation) of some species were also employed to improve their detectability.
Under the experimental conditions explored in this work, ES showed an
acceptable stability whereas the rest of species reacted with free chlorine at
significant rates following pseudo-first-order kinetics. Their half-lives ranged
from 0.4 to 25 min depending on the UV filter, chlorine concentration, water pH
and presence of bromide traces. For EHPABA a relatively simple degradation
pathway was established. It consisted of aromatic substitution of one atom of
hydrogen per chlorine or bromide. The same reaction pattern was observed for
BP-3 leading, in this case, to mono- and di-halogenated by-products. In
addition, several halogenated forms of 3-methoxyphenol were identified as BP-
3 cleavage by-products.
Metodología desarrollada-Muestras acuosas
228
Keywords: UV filters, halogenated by-products, chlorine, bromide, water
samples, gas chromatography-mass spectrometry
1. Introduction
Awareness regarding harmful effects of solar radiation on the skin has
increased the production and consumption of the so-called UV filters. These
compounds are classified as either inorganic or organic species. Most of the
organic UV filters are relatively lipophilic compounds, which contain aromatic
rings and conjugated carbon-carbon double bonds in their structures. They
absorb radiation at certain wavelengths in the range of 280-400 nm. UV filters
are considered as cosmetics which are incorporated in the formulation of
several personal care products, such as lipsticks, hair dyes, beauty creams,
shampoos and in particular sunscreen lotions [1-3]. In the latter application,
several organic filters are normally combined in the same product to provide
UV protection in a wide range of wavelengths. Their individual concentrations
may represent up to 10% of the product weight, according to the EU legislation
[4-6].
The presence of organic UV filters in the aquatic media, added to the
controversy about the potential endocrine disrupting activity of some of these
compounds [7-9], has increased concern about their possible environmental
effects. In the case of water samples, measured concentrations ranged from the
low ng L-1 in surface water [10-14] up to the ng mL-1 level in swimming pool,
bathing and wastewater [5,14-16]. Direct release from skin in swimming pools
and sunbathing areas, industrial wastewater and indirect discharges (e.g.
during showering, clothes washing and urinary excretion) through domestic
wastewater, represent the main input routes of these compounds in water
bodies [2,5-6,17]. UV filters have been detected in sludge from sewage
treatment plants with average concentrations over the 1 g g-1 level for the most
lipophilic ones, such as 3-(4-Methylbenzylidene) camphor (4-MBC), octocrylene
Journal of Chromatography A 1178 (2008) 206
229
(OC) and octyl triazone (OT) [18]. 2-Ethylhexyl methoxycinnamate (EHMC), 4-
MBC, OC and 2-hydroxy-4-methoxybenzophenone (BP-3) have been also found
in fish tissues from Swiss rivers and lakes receiving discharges of waste and
bathing water [14, 19]. In summary, the available information suggests that
some UV filters behave as persistent and bioaccumulative pollutants in the
aquatic environment.
Additionally to monitoring studies, a realistic estimation of potential
environmental risks associated with the presence of UV filters in the aquatic
media requires also to evaluate their degradation rates and the formation of by-
products. Up to now, very few articles have dealt with this topic. Most of them
have been focussed on the study of aqueous photolysis reactions [20-22]. These
works estimated the half-lives and the rate constants of several UV filters under
different experimental conditions [20-21]. In addition, Sakkas et al. [20] have
reported the formation of several chlorinated forms of 2-ethylhexyl 4-
(dimethylamino) benzoate (EHPABA) after a prolonged (60 hours) exposure of
chlorinated swimming-pool water samples, previously spiked with this
compound, to solar irradiation. This work constitutes the first evidence of the
reactivity between EHPABA and chlorine in water samples; however, it does
not provide data neither about the rate of EHPABA halogenation reactions nor
about the stability of the generated by-products.
Formation of halogenated disinfection by-products in chlorinated water
samples is a relevant process for many organic substances. This kind of
reactions is particularly favourable for species with phenolic and/or amino
groups in their structures [23-25]. In this sense, the reactivity of several
pharmaceuticals and personal care products in chlorinated water samples has
been studied recently [23,26-28]. For some compounds, kinetically favourable
reactions leading to the formation of relatively stable and toxic by-products
have been reported [23,29]. As far as we know, such information is not available
for organic UV filters.
Metodología desarrollada-Muestras acuosas
230
In this work, the reactivity of three commonly used UV filters containing
phenolic or amino groups, BP-3, EHPABA and 2-ethylhexyl salicylate (ES), in
chlorinated water samples was assessed. Aims of the study were: (1) to evaluate
the stability of parent species in presence of free chlorine (sodium hypochlorite
plus hypochlorous acid) at neutral pH values, (2) to obtain their half-lives under
different experimental conditions and (3) to identify the corresponding
halogenated by-products, as well as to determine their stability under oxidative
conditions. Experiments were carried out using aliquots of ultrapure water
spiked with free chlorine concentrations at the low g mL-1 level, similar to
those found in tap and swimming pool water, and standard solutions of each
UV filter in the ng mL-1 range. In further assays, commercial personal care
products, containing UV filters, were mixed directly with chlorinated tap water
to confirm the formation of those by-products observed in model experiments.
Solid-phase extraction and gas chromatography-mass spectrometry (GC-MS)
were employed for the concentration of water samples, the determination of
remaining UV filters levels and the identification of generated by-products.
2. Experimental
2.1. Reagents, standards and material
Methanol, dichloromethane and ethyl acetate, trace analysis grade, were
acquired from Scharlab (Barcelona, Spain). Standards of BP-3, ES, EHPABA and
2,4-dihydroxybenzophenone (2,4-DHBP), Fig. 1, as well as the silylation reagent
N-methyl-N-(tert-butyldimethylsilyl)trifluoroacetamide (MTBSTFA) were
purchased from Aldrich (Milwaukee, WI, USA). Sodium thiosulphate,
potassium bromide and potassium dihydrogen phosphate were obtained from
Merck (Darmstadt, Germany). Individual solutions of each compound were
prepared in methanol. Two series of calibration standards were made in ethyl
acetate (BP-3 and 2,4-DHBP) and dichloromethane (ES and EHPABA). BP-3, 2,4-
DHBP, and their by-products, were converted into the corresponding tert-
butyldimethylsilyl derivatives to improve their detectability by GC-MS.
Journal of Chromatography A 1178 (2008) 206
231
Derivatization was carried out adding 20 L of MTBSTFA to 500 L aliquots of
calibration standards or sample extracts in ethyl acetate [27,29]. The mixture
was shaken manually for 5 minutes and then injected in the GC-MS system.
MeO
C
OOH
BP-3
Bu-n
Et
Me2N
O CH2 CHC
O
EHPABA
C
O
Bu-n
Et
O CH2 CH
OH ES
HO
C
OOH
2,4-DHBP
Fig. 1. Structures of the considered UV filters and 2,4-DHBP.
Sodium hypochlorite with a nominal free chlorine content of 4% (w/v)
was purchased from Aldrich. This solution was stored at 4 ºC and its exact
concentration determined weekly by iodometric titration as reported elsewhere
[29]. Oasis HLB (60 mg) SPE cartridges were acquired from Waters (Milford,
MA, USA).
2.2. Chlorination experiments
Stability of selected UV filters and formation of halogenated by-products
was evaluated at room temperature (18 2 ºC) considering initial
concentrations of parent species and free chlorine in the ranges of 10-50 ng mL-1
and 0.1-3 g mL-1, respectively. Reactions were carried out in amber glass
vessels containing between 100 and 250 mL of ultrapure or tap water. Ultrapure
water samples were spiked with a titrated sodium hypochlorite solution to get
the required initial concentration of free chlorine. In the case of tap water
samples, their initial free chlorine content was determined using the N,N-
diethyl-p-phenylenediamine method with photometric detection [30]. After
that, samples were adjusted at different pHs, between 6 and 8 units (with the
intent to cover the range of values expected in real water samples, e.g. tap,
Metodología desarrollada-Muestras acuosas
232
swimming-pool and wastewater), and spiked with a solution of the considered
UV filter in methanol. In some experiments, personal care products, containing
UV filters, replaced standard solutions. Then, vessels were closed, shaken
manually for 2 minutes and allowed to stand. After an established reaction
time, the excess of chlorine was quenched with 10 mg of sodium thiosulphate
and samples processed as described below.
2.3. Analytical procedure
Water samples were concentrated using Oasis HLB 60 mg cartridges,
previously conditioned with methanol and ultrapure water (3 mL each). In
order to maximise the yield of the concentration step, slightly different
conditions were used depending on the considered compound. Samples spiked
with BP-3 or 2,4-DHBP were adjusted at pH 3, passed through the SPE cartridge
and analytes eluted with 2 mL of ethyl acetate. An aliquot of this extract was
derivatized as indicated in previous sections. For ES and EHPABA, samples
were adjusted to pH 4.5 and methanol (20% of water volume) was added to
prevent sorption losses on glass vessels and/or connections with the SPE
cartridge. After the enrichment step, analytes were recovered with 2 mL of
dichloromethane.
Parent UV filters and their by-products were determined by GC-MS
using a Varian CP 3900 gas chromatograph (Walnut Creek, CA, USA)
connected to an ion-trap mass spectrometer (Varian Saturn 2100T). Normally,
separations were carried out in a HP-5ms type capillary column (30 m x 0.25
mm I.D., df: 0.25 m) purchased from Agilent. Moreover, a HP-35 column, from
the same supplier and with same dimensions as the HP-5 one, was also used.
For both columns, the temperature of the GC oven was programmed as follows:
70 ºC (1 min), rate at 12 ºC min-1 to 280 ºC (10 min). Helium (99.999 %) was used
as carrier gas at a constant flow of 1 mL min-1. The injector was maintained at
280 ºC and injections (1-2 l) were made in the splitless mode with a purge time
of 1 min. Transfer line and ion trap temperatures were set at 280 and 220 ºC,
Journal of Chromatography A 1178 (2008) 206
233
respectively. Electron Impact (EI) mass spectra (MS) were recorded in the range
of 100-650 m/z units.
3. Results and discussion
3.1. Performance of the analytical procedure
Table 1 summarizes the most relevant data related to the performance of
the analytical procedure. Reported retention times were obtained with the HP-
5ms column, and, in the case of BP-3 and 2,4-DHBP correspond to their
silylated derivatives. The linearity in the response of the GC-MS system was
evaluated with standards at seven concentration levels between 2 and 2000 ng
mL-1. Correlation coefficients of the resulting graphs varied from 0.997 to 0.999,
and the instrumental quantification limits of the GC-MS system, defined for a
signal to noise ratio of 10 (S/N = 10), remained under 5 ng mL-1 (data not
shown). If the silylation step was not considered, no signal was observed for
2,4-DHPB, whereas a wide, tailing peak was obtained for BP-3. Yields of the
SPE process (n= 4 replicates) were calculated by external calibration. Under
conditions reported in the previous section, recoveries over 80 % and limits of
quantification (LOQs) between 8 and 25 ng L-1 were achieved for 500 mL
volume samples. Although, the performance of the analytical procedure for the
potential by-products of UV filters was not evaluated, it was assumed that they
behave in a similar way to parent species (particularly those presenting very
close structures), as regards the silylation reaction and affinity for the SPE
sorbent.
Table 1. Figures of merit of the analytical procedure for parent UV filters.
Analyte Retention time
(min)
Quantification
Ions (m/z)
Linearity, R2
(range, ng mL-1)
Recovery (%)
SD
LOQs
(ng L-1)
ES
EHPABA
BP-3a
2,4-DHBPa
13.09
16.91
17.34
21.65
138+120
277+165
285
385
0.999 (5-2000)
0.997 (5-2000)
0.999 (2-2000)
0.999 (2-2000)
84 7
82 6
98 8
89 8
24
25
8
8
a Data for their tert-butyldimethylsilyl derivatives
Metodología desarrollada-Muestras acuosas
234
3.2. Stability of UV filters at different pHs and free chlorine
concentrations
Previously to investigate the formation of halogenated by-products from
selected UV filters, the stability of BP-3, EHPABA and ES in presence of
increasing chlorine levels (up to 3 g mL-1) was assessed. Experiments were
carried with ultrapure water (100 mL aliquots) buffered at three different pHs,
covering the range of values expected in different water samples, and spiked at
50 ng mL-1 with only one of the considered compounds. After a fixed time (from
2 to 30 min), the excess of chlorine was removed and samples were
concentrated to determine the remaining amount of parent compound.
Obtained results are depicted in Fig. 2. Normalised values in the Y-axis
correspond to the ratios between the responses for each analyte in the SPE
extracts from chlorinated and non-chlorinated aliquots of ultrapure water,
multiplied by 100. Each point represents the average value for duplicate
experiments. As observed, the stability of the UV filters increased in the
following order: BP-3 < EHPABA < ES. For the latter, at the 3 investigated pHs,
more than 60% of the added amount was found in the sample even when a
relatively high chlorine concentration (3 g mL-1) and a long reaction time (30
min) were considered, Fig. 2. In a real life situation, with several organic species
competing for the available chlorine, the extension of ES halogenation reactions
was estimated as negligible; consequently, it was not longer considered in this
study. On the other hand, EHPABA and BP-3 showed a lower and pH
dependant stability. EHPABA was more stable at pH 8.2 than at pHs 6.2 and
7.2, Fig. 2. This behaviour suggests that the kinetics of the reaction between the
UV filter and sodium hypochlorite (pKa 7.5) is less favourable than with
hypochlorous acid. For BP-3 (pKa 7.6), at the two considered reaction times: 10
and 2 min, its stability decreased with the increase in the pH of the water
samples, Fig. 2. Therefore, within the explored range of pH, the ratio between
non-protonated and protonated forms of the UV filter seems to be the key factor
controlling its stability.
Journal of Chromatography A 1178 (2008) 206
235
BP-3
0%
20%
40%
60%
80%
100%
120%
0 0.5 1 1.5 2 2.5 3 3.5
Free chlorine (g mL-1)
Nor
mal
ised
resp
onse
pH 7.2
pH 6.2
pH 8.2
2 min
0%
20%40%
60%
80%100%
120%
0.0 0.2 0.4 0.6Free chlorine (g mL-1)
No
rmal
ised
resp
onse10 min
EHPABA, 10 min
0%
20%
40%
60%
80%
100%
120%
0 0.5 1 1.5 2 2.5 3 3.5
Free chlorine (g mL-1)
No
rmal
ised
resp
ons
e
pH 7.2
pH 6.2
pH 8.2
ES, 30 min
0%
20%
40%
60%
80%
100%
120%
0 0.5 1 1.5 2 2.5 3 3.5
Free chlorine (g mL-1)
Nor
mal
ised
resp
onse
pH 7.2
pH 6.2
pH 8.2
Fig. 2. Effect of free chlorine concentration and water pH on the stability of BP-3,
EHPABA and ES.
Metodología desarrollada-Muestras acuosas
236
3.3. Reaction kinetics and half-lives
Time course of BP-3 and EHPABA concentrations were followed using
250 mL aliquots of ultrapure and chlorinated tap water. Samples were buffered
at pH 7.2 (unless otherwise stated) and spiked with one of the above
compounds at the 10 ng mL-1 level (from 36 to 44 nM). The percentage of
methanol in the samples was maintained under 0.05% to avoid any possible
effect on the reaction rates. In addition to above compounds, a series of
experiments was carried out with 2,4-DHBP, which is the major metabolite of
BP-3. Initial concentrations of free chlorine added to ultrapure water varied
from 0.3 to 1 g mL-1 (4.4 to 14.5 M); moreover, in some assays, bromide (as
potassium salt) was also added to the samples to investigate its effect on the
stability of parent species. Tap water was collected when needed and the
content of free chlorine measured as indicated in the experimental section. After
a given time, the excess of chlorine was removed and the remaining amount of
each compound determined. At least 5 data were obtained between zero (the
quenching reagent was added to chlorinated water samples previously to the
UV filter) and 2-3 times the half-life of the considered compound. In excess of
chlorine, the removal of BP-3, EHPABA and 2,4-DHBP followed pseudo-first-
order kinetics, Fig. 3. For the latter compound, the reaction was so fast that it
could be studied only in samples with a relatively low concentration of
chlorine: 0.1 g mL-1, Fig. 3. Under these conditions, a half-life of 0.4 min was
measured. Taking into account that 2,4-DHBP only differs from BP-3 in the
substitution of the methoxy moiety by a hydroxyl one (Fig. 1), the effect of the
latter group on the reactivity of the compound is evident. This finding is in
agreement with differences between the stabilities of phenol and 1,3-
dihydroxibenzene in chlorinated water [25].
Journal of Chromatography A 1178 (2008) 206
237
BP-3Initial chlorine 0.3 g mL-1
y = -0.302x + 11.504
R2 = 0.997
8
9
10
11
12
0 1 2 3 4 5 6 7
Reaction time (min)
Ln (P
eak
area
)
EHPABAInitial chlorine 0.6 g mL-1
y = -0.026x + 12.474
R2 = 0.995
10
10.5
11
11.5
12
12.5
13
0 5 10 15 20 25 30 35 40 45
Reaction time (min)
Ln (P
eak
area
)
2,4-DHBPInitial chlorine 0.1 g mL-1
y = -1.742x + 13.251
R2 = 0.991
23456789
1011121314
0 0.5 1 1.5 2 2.5 3 3.5
Ln (P
eak
area
)
Fig. 3. Natural logarithm plots of analytes responses versus reaction time. Data for
ultrapure water buffered at pH 7.2.
Metodología desarrollada-Muestras acuosas
238
Half-lives measured for BP-3 and EHPABA, under different
experimental conditions, are given in Table 2. In agreement with the
information depicted in Fig. 2, BP-3 was less stable at pH 8.2 than at pH 7.2.
Addition of bromide to ultrapure water, even at the low ng mL-1 level, reduced
the stability of both UV filters, with the most significant influence on EHPABA.
This effect can be explained due to the formation of bromine, which shows a
strong tendency to react with aromatic compounds. It is also remarkable that
the half-life of BP-3 in chlorinated tap water was similar to that measured in
ultrapure water, spiked with a similar level of free chlorine. On the other hand,
significant differences were noticed for EHPABA.
Table 2. Half-lives (t1/2) measured for BP-3 and EHPABA under different
experimental conditions.
Analyte Matrix pH Free chlorine
(g mL-1)
Bromide
(ng mL-1)
t1/2 (min)
BP-3
Ultrapure
water
7.2 0.30a 0 2.7
7.2 0.60a 0 1.2
8.2 0.30a 0 1.8
8.2 0.60a 0 0.8
7.2 0.30a 1 2.8
7.2 0.30a 10 0.8
Tap water 7.2 0.76b Not determined 1.0
EHPABA
Ultrapure
water
7.2 0.6a 0 26.7
7.2 1.0a 0 18.7
7.2 0.60a 1 21.3
7.2 0.60a 10 4.6
Tap water 7.2 0.63b Not determined 12.1
a Added concentration
b Measured concentration
Journal of Chromatography A 1178 (2008) 206
239
3.4. Halogenated by-products
Retention times and most intense ions in the MS spectra of major BP-3
and EHPABA by-products are shown in Table 3. They were the result of
hydrogen replacement per chlorine and/or bromine in the aromatic rings of
both UV filters. Positions where those replacements occurred were not
confirmed experimentally; however, considering the structures of parent
species (Fig. 1) and the activation effects of hydroxyl and amino groups towards
electrophilic substitution reactions, the most probable ones are the carbons in
ortho- to the amino moiety (EHPABA), and those in ortho- and para- to the
hydroxyl group (BP-3). For EHPABA only mono-halogenated species were
detected, whilst for BP-3 mono and di-substituted by-products were identified.
Although EHPABA and its chlorinated by-product could not be separated with
the HP-5ms column, using a HP-35 one, it was confirmed that signals in their
mass spectra appeared at different m/z ratios (Fig. 4); therefore, their relative
amounts can be determined even when they co-eluted in the same peak.
Table 3. Retention times (HP-5ms column), identities and most intense ions, with their
relative abundances, for the halogenated substitution by-products of BP-3 and
EHPABA.
UV filter By-product Retention time (min) m/z ions
(relative abundances)
BP-3
Cl-BP-3 (1) 18.55 319 (100), 321 (33), 276 (12)
Cl-BP-3 (2) 18.71 319 (100), 321 (33), 276 (12)
DCl-BP-3 19.03 353 (100), 355 (66), 310 (17)
Br-BP-3 (1) 19.25 363 (100), 365 (100), 320 (11), 323 (4)
Br-BP-3 (2) 19.41 363 (100), 365 (100), 320 (11), 323 (4)
DBr-BP-3 20.70 443 (100), 441 (50), 445 (50), 428 (13)
Br-Cl-BP-3 (1) 19.75 399 (100), 397 (82), 401 (53), 384 (20)
Br-Cl-BP-3 (2) 19.80 399 (100), 397 (82), 401 (53), 384 (20)
EHPABA Cl-EHPABA 16.91 198 (100), 311 (75), 313 (25)
Br-EHPABA 17.50 355 (100), 357 (100), 243 (94)
(1) and (2) refer to isomeric by-products
Metodología desarrollada-Muestras acuosas
240
Ultrapure water
Ultrapure water plus 0.6 g mL-1
of chlorine, 10 min
198
150 200 250 300 350m/z
0%
25%
50%
75%
100%
118
154
182
311
Cl-EHPABA
150 200 250 300 m/z
0%
25%
50%
75%
100%
120
148
165
277
EHPABA
21.25 21.50 21.75 22.00 22.25 22.50 22.75 minutes
0
100
200
300
400
(x103 Counts)
Fig. 4. Total ionic current (TIC) chromatograms (HP-35 column) and MS spectra for
EHPABA and its chlorinated by-product.
Table 4 summarizes which substitution by-products were observed
under different experimental conditions. Formation of brominated species
when BP-3 and EHPABA standards were mixed with chlorinated tap water
confirmed the presence of bromide in this matrix and explained the differences
between the half-lives of EHPABA in ultrapure (26.7 min) and tap water (12.1
min) containing a similar concentration of free chlorine, see Table 2. Last
column on Table 4 presents the most concerning results: with the only
exception of DBr-BP-3, the rest of EHPABA and BP-3 substitution by-products,
identified in model experiments, were also noticed after mixing tap water with
two personal care products including the parent UV filters in their formulation.
Journal of Chromatography A 1178 (2008) 206
241
Table 4. Summary of BP-3 and EHPABA substitution by-products detected under
different experimental conditions.
Ultrapure
water plus
UV filter
standard
Ultrapure
water plus
UV filter
standard
Ultrapure
water plus
UV filter
standard
Tap water
plus
UV filter
standard
Tap water
plus
suncare
product
BP-3
Free
chlorine
(g mL-1)
0.3a 0.3a 0.3a 0.76b 0.55b
Bromide
(ng mL-1) 0 1a 10a
not
determined
not
determined
Cl-BP-3 (1) X X X X X
Cl-BP-3 (2) X X X X X
DCl-BP-3 X X X X X
Br-BP-3 (1) -- X X X X
Br-BP-3 (2) -- X X X X
DBr-BP-3 -- X X -- --
Br-Cl-BP-3
(1) -- X X X X
Br-Cl-BP-3
(2) -- X X X X
EHPABA
Free
chlorine
(g mL-1) 0.6a 0.6a 0.6a 0.63b 0.89b
Bromide
(ng mL-1) 0 1a 10a
not
determined
not
determined
Cl-
EHPABA X X X X X
Br-
EHPABA -- X X X X
aadded concentration
bmeasured concentration
Additionally to the above mentioned species, in the case of BP-3 another
group of by-products was detected. They were tentatively identified as
Metodología desarrollada-Muestras acuosas
242
halogenated forms of 3-methoxyphenol, which could be generated from
cleavage of the carbonyl bond between the two aromatic rings in the molecule
of BP-3 (Fig. 1), followed by halogenation of the methoxyphenol fragment.
Moreover, mono- and di-halogenated substitution by-products of BP-3 might
also break down rendering different halogenated methoxyphenols. Fig. 5 shows
the MS spectra, as tert-butyldimethylsilyl derivatives, for two of these cleavage
by-products. Retention times, most intense ions, and the number of chlorine
and/or bromine atoms for all the BP-3 cleavage by-products are shown in Table
5. As observed, only di- and tri-halogenated species were detected.
Chromatographic responses of the first represented less than 2% of the BP-3
peak area in the reference experiment, performed in absence of chlorine;
whereas, for tri-halogenated phenols, they reached up to 10%. Moreover, the
trichloro-methoxyphenol was the only cleavage by-product detected when
mixing tap water with a BP-3 containing suncare lotion.
150 200 250 300 350 400m/z
0%
25%
50%
75%
100%
213
294
373
A B
150 200 250 300 350m/z
0%
25%
50%
75%
100%
248
268
285
m/z 283+285
MeO
O
Cl
Cl
Cl
Si
m/z 248+250
MeO
OCl
Si
m/z 373+375
Br
Br
m/z 292+294
Fig. 5. MS spectra and tentative structures for two cleavage halogenated by-products of
BP-3.
Journal of Chromatography A 1178 (2008) 206
243
Table 5. Halogenated cleavage by-products of BP-3.
By-product structure Atoms of
chlorine
Atoms of
bromine
Retention time
(min)
Most intense ions
(m/z)
Methoxy-phenol
2 0 13.63 249, 251
3 0 14.47 283, 285
0 2 15.20 337, 339, 341
0 3 16.79 417, 419
1 2 16.04 373, 375
3.5. Stability of the halogenated by-products
Stability of major EHPABA and BP-3 by-products was followed
considering reaction times up to 180 and 120 min, respectively and chlorine
concentrations up to 1 g mL-1. Some of the obtained data are plotted in Fig. 6.
Normalised responses represented in the Y-axis correspond to the ratio between
the peak area of each by-product, at a given reaction time, and that measured
for the parent UV filter in the reference experiment at zero time; therefore, they
serve as a rough estimation for the yield of depicted transformations. EHPABA
by-products showed an excellent stability under explored conditions, Fig. 6A.
The presence of just 10 ng mL-1 of bromide, led to a significant diminution in
the production of Cl-EHPABA and shifted the reaction towards the brominated
specie. Di-halogenated forms of EHPABA were never detected, thus, insertion
of one atom of chlorine or bromide in the structure of the parent specie seems to
deactivate it to further electrophilic substitution reactions. This finding differs
from the data published by Sakkas et al. for chlorinated swimming-pool water
[20]. The use of longer reaction times added to the effect of solar radiation
(exposure time 60 hours) might be responsible for the formation of di-
halogenated EHPABA by-products reported in the above work [20].
Metodología desarrollada-Muestras acuosas
244
0
10
20
30
40
50
60
0 50 100 150 200
Reaction time (min)
No
rmal
ised
res
po
nse
(%
)Cl-EHPABA, 0 ng mL-1 Br-
Cl-EHPABA, 10 ng mL-1 Br-
Br-EHPABA, 10 ng mL-1 Br-
0
10
20
30
40
50
0 20 40 60 80 100 120
Reaction time (min)
No
rmal
ized
res
po
nse
(%
)
DCl-BP-3Cl-BP-3 Br-Cl-BP-3 DBr-BP-3 x 5Br-BP-3
A
B
C
0
1
2
3
4
5
6
7
8
9
0 20 40 60 80 100 120 140
Reaction time (min)
No
rmal
ised
res
po
nse
(%
)
Ultrapure water (1 g mL-1 chlorine) plus BP-3 standard
Tap water (0.55 g mL-1 chlorine) plus sun care product
Fig. 6. Time course of some UV filters halogenated by-products. A, EHPABA by-
products for 0.6 g mL-1 of free chlorine and different levels of bromide. B, Mono- and
di-halogenated BP-3 species for 0.3 g mL-1 of chlorine and 1 ng mL-1 of bromide. C,
Trichloro-3-methoxyphenol in chlorinated ultrapure and tap water.
Journal of Chromatography A 1178 (2008) 206
245
Fig. 6B presents the time course of BP-3 substitution by-products for 0.3
g mL-1 of chlorine and 1 ng mL-1 of bromide. The maximum responses for
mono-halogenated species were observed for short reaction times, whereas the
di-halogenated ones reached a plateau after 20 min and then remained stable.
Additional experiments, using higher concentrations of free chlorine, showed
that the latter are also slowly degraded. Particularly, for 1 g mL-1 of free
chlorine in absence of bromide, the normalised response of DCl-BP-3 reached a
maximum (around 50%) after 20 min and then decreased to 20% in 2 hours,
figure not shown.
Trichloro-methoxyphenol, the most abundant of the BP-3 cleavage by-
products, showed a considerable stability in ultrapure water containing 1 g
mL-1 of free chlorine and also when a suncare lotion was mixed with tap water
(0.55 g mL-1 of free chlorine). In both cases, its normalised response rose until
40-60 min and then decreased very slowly, Fig. 6C.
On the basis of the identified by-products and the temporal stabilities of
the most abundant ones, a tentative reaction pathway for BP-3 has been
proposed, Fig. 7. On one hand, the parent species is converted into mono and
di-halogenated substitution by-products. Moreover, BP-3 and/or its
substitution by-products may undergo a cleavage process rendering
methoxyphenols, which can be further halogenated by the excess chlorine
and/or bromine.
Metodología desarrollada-Muestras acuosas
246
MeO
C
OH
X
MeO
C
OOH
X
X
BP-3
O
MeO
C
OH O
MeO
C
OH
X
O
Substitution by-products
X= Cl and/or Br
MeO
OH
X
X
X
Cleavage by-products
MeO
OH
X
X
Fig. 7. Proposed degradation pathway for BP-3.
4. Conclusions
Conversely to ES, EHPABA and BP-3 exhibited a limited stability at
neutral pHs in water samples containing low levels of free chlorine. Differences
among reactivities of ES, BP-3 and 2,4-DHBP highlight the effect of different
organic groups on the activation or deactivation of the phenolic ring towards
electrophilic substitution reactions. Whilst EHPABA rendered only mono-
halogenated by-products, with the same structure as the parent UV filter, BP-3
followed a more complex reaction pathway leading to mono- and di-
halogenated substitution compounds and cleavage halogenated
methoxyphenols. Traces of bromide, at similar levels to those existing in
aquifers and sources of tap water from coastal areas, shortened considerably the
half-life of EHPABA; moreover, they shifted the degradation of BP-3 and
EHPABA towards the formation of brominated by-products, which, in general,
are considered more concerning than the chlorinated ones.
Journal of Chromatography A 1178 (2008) 206
247
By-products of EHPABA, di-halogenated BP-3 forms and the tri-
halogenated methoxyphenols were produced in a significant extension, showed
a considerable stability to further oxidation reactions and their formation was
even observed under quasi real-life conditions; therefore, it is expected that
they could be generated in chlorinated bath waters (e.g. swimming-pools) and
during showering, after dermal application of suncare products. Further studies
should investigate the presence of these by-products in waste and swimming-
pool water, address their potential toxicity due to dermal exposition and
investigate possible environmental effects.
Acknowledgements
This study has been supported by Spanish Government, Xunta de
Galicia, and E.U. FEDER funds (projects DGICT CTQ2006-03334 and
PGIDIT06PXIB237039PR). N. N and P. C. thank the Spanish Ministry of
Education and Science for their FPU grants.
Metodología desarrollada-Muestras acuosas
248
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18. C. Plagellat, T. Kupper, R. Furrer, L. F. de Alencastro, D. Grandjean, J. Tarradellas,
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19. H.R. Buser, M.E. Balmer, P. Schmid, M. Kohler, Environ. Sci. Technol. 40 (2006) 1427.
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21. D.L. Giokas, A.G. Vlessidis, Talanta 71 (2007) 288.
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Metodología desarrollada-Muestras sólidas
253
2. MUESTRAS SÓLIDAS
2.1. Introducción
Algunos filtros solares presentan carácter fuertemente lipofílico y
elevada estabilidad en condiciones ambientales. Por tanto, son propensos a
acumularse en matrices sólidas ambientales, tales como lodos y sedimentos.
Desde un punto de vista analítico, la preparación de muestra para la
determinación de contaminantes emergentes en lodos es un proceso complejo,
que requiere múltiples etapas de clean-up y un elevado consumo de disolventes
orgánicos. Además, a diferencia de lo que ocurre con las matrices acuosas, el
número de aplicaciones desarrolladas es limitado, así como el número de
analitos incluidos en las mismas. En esta Tesis, se propone un protocolo semi-
automatizado combinando PLE con una primera etapa de limpieza on-line
usando carbón grafitizado para la eliminación de pigmentos, seguido de una
etapa adicional de purificación en modo off-line, utilizando PSA, para la
determinación de 8 filtros solares en lodos de depuradoras mediante GC-MS. El
método desarrollado ha sido validado y aplicado al análisis de lodos de
diferentes orígenes, confirmando la acumulación de determinados analitos en
esta matriz [Negreira, in press].
Además de su utilización en protectores solares, los filtros UV se
emplean también en otros productos de cuidado personal y materiales
ampliamente usados en el hogar, tales como pinturas y barnices, plásticos,
tapicerías y alfombras. Estos usos permiten inferir su presencia en atmósferas
interiores, lo que provocaría una exposición continua e inadvertida a estos
compuestos a través de la inhalación e incluso, la ingestión. Dada su limitada
volatilidad, parece lógico que los filtros UV se encuentren asociados a polvo y
material particulado, aunque su existencia en estas matrices no había sido
descrita previamente en la bibliografía. En esta Tesis, se desarrolló un método
Metodología desarrollada-Muestras sólidas
254
rápido, sencillo y de bajo coste para investigar la presencia de filtros solares en
muestras de polvo procedentes de viviendas particulares, edificios públicos y
vehículos. Las etapas de extracción y purificación se integraron en un solo paso
empleando MSPD como técnica de preparación de muestra y GC-MS/MS en la
etapa de determinación [Negreira, 2009-C]. La aplicación del método
desarrollado puso de manifiesto la existencia de niveles significativos de
algunos filtros solares en atmósferas interiores, asociados a partículas de polvo.
Metodología desarrollada-Muestras sólidas
255
2.2. Esquemas de los métodos desarrollados para muestras
sólidas
Extracción
Mezclar con 2 g tierra de diatormeas
Muestra lodo 0,5 g
Relleno de la celda de extracción:
PLE:Hexano: Diclorometano (80:20), 1 ciclo, 5 min, 75 ºC, 1500 psi,
100% flush, 2 min purga
Concentración a 1 mL
Limpieza
Clean-up en cartucho de PSA-sílica (0,5 g)
Fracción 2ª :5 mL de hexano: éter (1:1)
Fracción 1ª : 1 mL hexano
Añadir 2 mL isooctano
Inyección 2 µLGC-MS (modo SIM)
Concentrar a 1 mL
1ºFiltros (1 de celulosa + 1 de fibra de vidrio)2ºTierra de diatomeas (1,5 g)3ºCarbón (0,5 g)4ºTierra de diatomeas (0,5 g)5ºMezcla de lodo + tierra de diatomeas6ºFiltro de celulosa
Figura 21: Esquema seguido para la determinación de filtros solares en lodo mediante
PLE y GC-MS.
Metodología desarrollada-Muestras sólidas
256
Frita de polietileno
Dispersión con1,25 g C18
Evaporación a 1 mL
Polvo0,5 g
Secar con 0,5 g sulfato sódico anhidro
2 g de Sílica (Co-adsorbente)
Transferir la mezcla al cartucho de MSPD
4 mL acetonitrilo
..Frita de polietileno
Elución
Inyección 2 µL GC-MS/MS
Figura 22: Esquema empleado para la determinación de filtros solares en polvo mediante
MSPD y GC-MS/MS.
2.3. Publicación:
OPTIMIZATION OF
PRESSURIZED LIQUID EXTRACTION
AND PURIFICATION CONDITIONS FOR
GAS CHROMATOGRAPHY-MASS SPECTROMETRY
DETERMINATION OF UV FILTERS IN SLUDGE
N. Negreira, I. Rodríguez, E. Rubí, R. Cela
Journal of Chromatography A, in press
(doi:10.1016/j.chroma.2010.11.028)
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
259
Optimization of pressurized liquid extraction and purification conditions for
gas chromatography-mass spectrometry determination of UV filters in sludge
N. Negreira, I. Rodríguez*, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto de
Investigación y Análisis Alimentario (IIAA), Universidad de Santiago de
Compostela, Santiago de Compostela 15782, Spain.
Abstract
This work presents an effective sample preparation method for the
determination of eight UV filter compounds, belonging to different chemical
classes, in freeze-dried sludge samples. Pressurized liquid extraction (PLE) and
gas chromatography-mass spectrometry (GC-MS) were selected as extraction
and determination techniques, respectively. Normal-phase, reversed-phase and
anionic exchange materials were tested as clean-up sorbents to reduce the
complexity of raw PLE extracts. Under final working conditions, graphitized
carbon (0.5 g) was used as in-cell purification sorbent for the retention of co-
extracted pigments. Thereafter, a solid-phase extraction cartridge, containing
0.5 g of primary secondary amine (PSA) bonded silica, was employed for off-
line removal of interferences overlapping the chromatographic peaks of some
UV filters. Extractions were performed with a n-hexane:dichloromethane (80:20,
v:v) solution at 75 ºC, using a single extraction cycle of 5 min at 1500 psi. Flush
volume and purge time were set at 100% and 2 min, respectively. Considering
0.5 g of sample and 1 mL as the final volume of the purified extract, the
developed method provided recoveries between 73% and 112%, with limits of
quantification (LOQs) from 17 to 61 ng g-1. Total solvent consumption remained
around 30 mL per sample. The analysis of non-spiked samples confirmed the
sorption of significant amounts of several UV filters in sludge with average
concentrations above 0.6 g g-1 for 3-(4-methylbenzylidene) camphor (4-MBC),
2-ethylhexyl-p-methoxycinnamate (EHMC) and octocrylene (OC).
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Keywords: UV filters; sludge; purification; pressurized liquid extraction;
GC-MS.
1. Introduction
Organic UV filters are compounds designed to absorb the ultraviolet
wavelengths of solar radiation preventing photo-aging and other harmful
effects in human health. The concentration of UV filters in sunscreen lotions
may represent up to 10% of the product weight; moreover, they are included, at
lower levels, in the formulation of many other personal care products [1-2]. The
above uses contribute to the direct input of UV filters in bathing waters and
their indirect release in the aquatic environment through domestic sewage
water [3-7]. The activity of some UV filters as endocrine disrupters [8-10],
added to their ubiquity in sewage and surface water, has awaken the concern
about their potential medium-term environmental effects.
Gas and liquid chromatography-mass spectrometry techniques,
combined with effective sample concentration approaches [5,11-13], have been
applied to obtain an overview of UV filters occurrence in different water
samples, including wastewater from sewage treatment plants (STPs). However,
understanding the behaviour of UV filters in STPs requires not only measuring
their concentrations in the water phase, but also determining the fraction which
remains attached to sludge particles [7]. This latter information is necessary to
distinguish between biodegradation and sorption processes, and to assess the
risk of introducing the UV filters in the terrestrial environment through the
application of sludge as fertilizer in agriculture.
From the analytical point of view, sludge is an extremely complex matrix
which requires well-tuned sample preparation approaches providing a balance
among efficiency, selectivity, extraction time and cost. These constraints explain
the limited number of studies dealing with the analysis of UV filters in sludge
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
261
versus the plethora of publications focussed on water samples. The first method
for sludge was proposed by Plagellat and co-workers [14]. It involved three
consecutive liquid-liquid extractions of fresh sludge samples (60 g), followed by
dryness evaporation of the combined extract and column purification with
activated silica. The optimized protocol provided an excellent performance for
selected analytes; however, it was time and solvent consuming (more than 200
mL of organic solvent per sample), as well as difficult to automate.
Pressurized liquid extraction (PLE) is a popular sample preparation
technique for solid matrices showing limited solvent consumption, excellent
extraction yields and possibility to integrate extraction and purification steps.
The applications of PLE to the extraction of personal care compounds from
sludge have been compiled in a recent review [15]. PLE, combined with in-cell
clean-up using activated silica, has been reported as a straight forward
alternative for gas chromatography-mass spectrometry (GC-MS) determination
of UV filters in low carbon content sediment samples [16]; however, the above
strategy provided too complex extracts in the case of sludge [17]. Although, the
selectivity of the extraction could be improved by enclosing the sludge sample
in a non-porous polyethylene membrane bag, within the cell, the efficiency of
the extraction underwent a dramatic reduction, with recoveries around or
below 50% for most UV filters [17]. In addition to the above procedures,
specifically designed for UV filters, Nieto and co-workers [18] have developed a
PLE method for the extraction of several personal care products, including three
UV filters (benzophenone-3, BP-3; octocrylene, OC; and 2-ethylhexyl-p-
dimethylaminobenzoate, EHPABA), from sludge samples. Analytes were
recovered with methanol followed by methanol:water mixtures and
simultaneously purified in a layer of alumina packed inside the extraction cell.
Considering a sample intake of 1 g and 25 mL as the volume of the final extract,
recoveries over 79% and low signal suppression effects (below 15%) were
observed in the further LC-(ESI)-MS/MS determination. Unfortunately, the
performance of this method has not been assessed for other UV filters of
Metodología desarrollada-Muestras sólidas
262
environmental relevance, such as salicylates, methoxycinnamates and
camphors; moreover, the employed aqueous extraction mixture is not
compatible with GC-based determinations.
In this study, we optimize an alternative sample preparation method for
the determination of eight UV filters, belonging to different chemical classes, in
freeze-dried sludge samples. PLE was selected as extraction technique due to its
high automation capabilities. Purification conditions were optimized in order
(1) to reduce the content of interferences in the final extract and (2) to maintain
the consumption of organic solvents and the complexity of the method at
acceptable levels. GC-MS was considered as determination technique on the
basis of (1) its worldwide availability in environmental monitoring laboratories
and (2) the poor detection limits reported for salicylate type UV filters using
LC-(ESI)-MS systems [19]. Finally, the applicability of the method was
demonstrated with sludge samples from urban STPs.
2. Experimental
2.1. Solvents, standards and sorbents
N-hexane, isooctane, acetone, dichloromethane and ethyl ether (trace
analysis grade) and HPLC-grade methanol were supplied by Merck
(Darmstadt, Germany). The list of UV filters included in this study is compiled
on Table 1. Standards of target analytes were acquired from Aldrich
(Milwaukee, WI, USA) and Merck, except isoamyl-p-methoxycinnamate
(IAMC), which was kindly provided by Dr. R. Rodil (University of Santiago de
Compostela, Spain). Individual solutions of each species (ca. 1000 g mL-1) were
prepared in methanol. Further dilutions and mixtures of them were dissolved
in acetone (when used to prepare the spiked sludge samples employed during
optimization and validation of sample preparation conditions) and in isooctane
(case of calibration standards).
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
263
Table 1. Abbreviations, retention times, selected ions and instrumental limits of
quantification (LOQs) of the GC-MS system of target analytes.
Analyte Abbreviation Retention
time (min)
Quantification
ion (m/z)
Qualification
ion (m/z)
LOQs
(ng mL-1)
(S/N 10)
2-Ethylhexyl salicylate EHS 9.65 120 138 2
Homosalate HMS 10.26,10.40 120 138 4
Isoamyl-p-
methoxycinnamate IAMC 11.63a 178 161 2
2-Hydroxy-4-
methoxybenzophenone BP-3 11.63 227 151 6
3-(4-
Methylbenzylidene)
camphor
4-MBC 11.87a 254 239 3
2-Ethylhexyl-p-
dimethylaminobenzoate EHPABA 13.41 277 165 1
2-Ethylhexyl-p-
methoxycinnamate EHMC 13.73a 178 161 1
Octocrylene OC 16.15 360 232, 249 2 aRetention time values for the E isomers
Alumina, Florisil and silica solid-phase extraction (SPE) cartridges (0.5 g)
were acquired from Waters (Milford, MA, USA). Cartridges containing 0.5 g of
silica bonded to ethylenediamine-N-propyl groups (PSA sorbent) and 0.25 g of
graphitized carbon were purchased from Supelco (Bellefonte, PA, USA). Both
sorbents, in the bulk format, were also obtained from Supelco. Diatomaceous
earth was provided by Aldrich.
2.2. Samples
Optimization of sample preparation (extraction and purification)
conditions was performed with a freeze-dried pooled matrix of primary and
biological sludge, fortified with 5 g g-1 of each UV filter. The total carbon (TC)
content of the pooled matrix was 33%. The spiking procedure consisted of the
addition of a measured volume of a standard in acetone to an accurately
Metodología desarrollada-Muestras sólidas
264
weighed fraction of sludge. Approximately, 1 mL of standard was used per g of
freeze-dried sludge. The resulting slurry was protected from light,
homogenized periodically and kept in a hood until complete elimination of the
acetone. The recoveries of the method were evaluated with individual samples
of primary and biological sludge fortified at different concentrations. All spiked
samples were aged for a minimum of two weeks before extraction.
2.3. Sample preparation
Extractions were performed with a pressurized liquid extractor, ASE 200
Dionex (Sunnyvale, CA, USA), furnished with 11 mL stainless-steel cells. A
cellulose filter, followed by a glass fibre one, was placed on the bottom of each
cell. Under final working conditions, cells were filled (bottom to top) with 1 g of
diatomaceous earth, 0.5 g of graphitized carbon, 0.5 g of diatomaceous earth
and 0.5 g of sludge, previously homogenized with 2 g of diatomaceous earth.
Analytes were extracted with n-hexane:dichloromethane (80:20), at 75 ºC,
considering a single static extraction cycle of 5 min with the cell pressurized at
1500 psi. The flush volume was 100% and the purge time 2 min.
PLE extracts were evaporated, ca. 1 mL, and additionally purified with a
PSA cartridge (0.5 g) previously conditioned with n-hexane:ether (1:1) and n-
hexane (5 mL each). After loading the concentrated extract, the sorbent was
rinsed with n-hexane (1 mL). Analytes were further recovered with 5 mL of n-
hexane:ether (1:1). Thereafter, 1 mL of isooctane was added as a keeper to the
purified extract, which was evaporated and adjusted to a final volume of 1 mL
with the same solvent.
2.3. GC-MS equipment
UV filters were determined with a GC-MS system consisting of an
Agilent (Wilmington, DE, USA) 7890A gas chromatograph connected to a
quadrupole type mass spectrometer (Agilent MS 5975C), furnished with an
electron-impact (EI) ionization source. Separations were carried out in a HP-
5ms type capillary column (30 m x 0.25 mm i.d., df: 0.25 m) supplied by
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
265
Agilent. Helium (99.999 %) was used as carrier gas at a constant flow of 1.2 mL
min-1. The GC oven was programmed as follows: 110 ºC (held for 1 min),
increased at 12 ºC min-1 to 280 ºC (held for 10 min). Ionization source, mass
analyzer and transfer line temperatures were set at 230, 150 and 290 ºC,
respectively. Standards and sample extracts were injected in the splitless mode,
maintaining the injection port at 280 ºC. The splitless time and the split flow
were set at 1 min and 20 mL min-1, respectively. The mass spectrometer was
operated in the SCAN mode (m/z range from 45 to 400) to assess the efficiency
of the purification process, and in the SIM mode for quantification purposes.
Retention times and ions monitored for each compound are summarized in
Table 1. Analytes were grouped in three chromatographic segments. The dwell
time per ion was 100 ms in the first and third segment and 50 ms for the second
one.
2.4. Recoveries and procedural blanks
Signals (peak areas) measured for sample extracts were compared with
those obtained for calibration standards in isooctane, covering the range of
concentrations between 5 and 5000 ng mL-1. Within this range, the GC-MS
system provided linear response plots with determination coefficients (R2)
higher than 0.996 for all compounds. The instrumental limits of quantification
(LOQs), defined as the concentration of each compound producing a response
ten times higher than the baseline noise in the SIM acquisition mode, ranged
from 1 to 6 ng mL-1, Table 1. Recoveries were calculated as the difference
between concentrations obtained for spiked and non-spiked fractions of the
same sludge sample divided by the added amount and multiplied by 100.
Procedural blanks represent the whole sample preparation process (extraction
plus purification) performed without sludge.
Metodología desarrollada-Muestras sólidas
266
3. Results and discussion
3.1. Preliminary experiments
Previous applications of PLE to the extraction of UV filters from sludge
employed rather different conditions as regards the extraction solvent and the
temperature of the cell [17-18]. Likely, the selected in-cell clean up strategies,
based on the use of permeable non-porous membranes [17] or a normal-phase
sorbent [18], conditioned the optimum extraction parameters.
In this study, in order to prevent the influence of clean-up conditions on
the yield of PLE, a first series of extractions was carried out considering just
diatomaceous earth as inert dispersant of sludge (0.5 g of sludge plus 2 g of
diatomaceous earth) and filling material in the extraction cell. Flush volume,
pressure and extraction time were set at 100%, 1500 psi and 5 min, respectively.
Samples were first extracted with n-hexane (50 ºC, 1 cycle) followed by
dichloromethane (60 ºC, 3 cycles). Extracts were collected in separated vessels,
adjusted to 25 mL, filtered (0.45 m) and injected in the GC-MS system. N-
hexane, at low temperature, has been proposed for the removal of low polar
interferences, previously to analytes extraction from complex samples [20-21].
On the other hand, Chu and co-workers [22] described the use of
dichloromethane, under above instrumental conditions, for the PLE extraction
of medium-polar personal care compounds from sludge.
Analysis of n-hexane and dichloromethane extracts revealed that BP-3
and OC were distributed between both fractions, whereas 95% of the responses
measured for the rest of UV filters corresponded to the n-hexane fraction, data
not shown. Despite the high dilution of sample extracts (25 mL), their GC-MS
chromatograms showed a considerable complexity. Visually, the n-hexane
extract was colourless whereas the dichloromethane fraction presented a
yellowish appearance and turbidity. This preliminary data indicate the
suitability of n-hexane:dichloromethane mixtures for the extraction of UV filters
from sludge at relatively low temperatures. However, there is no possibility to
improve the selectivity of the process using n-hexane as pre-extraction solvent.
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
267
3.2. Clean-up conditions
Several SPE sorbents were tested in order to reduce the complexity of
PLE extracts from sludge samples. In all cases, extractions were carried out at 70
ºC using a n-hexane:dichlorometane (8:2) mixture with the cells pressurized at
1500 psi. Two static cycles of 2 min each and a flush volume of 100% were
employed. Extracts were concentrated to 1 mL and loaded on top of the
considered SPE cartridge, previously conditioned as described in the
experimental section. Thereafter, 1 mL of n-hexane was passed through the
sorbent and discarded. Subsequently, analytes were eluted using 5 mL of a n-
hexane: ether (1:1, v:v) mixture. The extract was mixed with 1 mL of isooctane
and evaporated to a final volume of 1 mL. No differences were noticed between
the turbidity and the colour of raw PLE extracts versus those purified with
alumina and silica cartridges. Florisil and PSA cartridges rendered transparent,
although yellowish, extracts and graphitized carbon transparent, colourless
ones. The efficiency of the above clean-up sorbents was evaluated operating the
GC-MS system in the SCAN mode. PSA was the only sorbent able to remove
two broad chromatographic bands (tentatively identified as fatty acids)
overlapping the peaks of HMS, BP-3, IAMC and 4-MBC, and to reduce
significantly the baseline level of the GC-MS chromatograms, Fig. 1. Likely,
fatty acids and other interferences with anionic moieties remain strongly
retained in the PSA cartridge, as it has been early described for the purification
of extracts from vegetal samples [23]. Except graphitized carbon, the rest of
sorbents failed to remove pigments contained in the raw extract. Although
pigments exerted a little effect in the complexity of the GC-MS chromatograms
(Fig. 1), they might impair the efficiency of the GC column due to irreversible
contamination of the stationary phase. Thus, PSA and graphitized carbon were
selected as clean-up sorbents.
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268
8.5 9.5 10.5 11.5 12.5 13.5 14.5 15.5 16.5 17.5 18.5 19.50
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
Time (min)
Abundance (x 107)
A
BCD
No clean-up PSA clean-up Carbon clean-up
Fig. 1. GC-MS chromatograms (SCAN mode) and pictures corresponding to PLE
extracts from sludge purified with different SPE cartridges: A, without clean-up; B,
Florisil; C, graphitized carbon; D, PSA. Extraction conditions: n-
hexane:dichloromethane (80:20), 70 ºC, 2 cycles of 2 min,100 % flush volume, 1500 psi.
In a second series of extractions, the feasibility of integrating extraction
and clean-up steps, placing a layer of the above sorbents (from 0.5 to 2 g) inside
the PLE cell, was investigated. Using the above described extraction
parameters, graphitized carbon (0.5 g) allowed an efficient removal of
pigments; however, the purification efficiency of PSA underwent a dramatic
reduction. Probably, the ability of this sorbent to retain fatty acids interferences
is reduced due to the temperature of the PLE cell (70 ºC versus room
temperature in the off-line modality), as well as the differences in the volume
and the composition of the organic mixture flowing through the layer of PSA,
packed inside the cell, versus those used in the SPE mode [24]. Graphitized
carbon (0.5 g) was introduced in the PLE cell for on-line removal of pigments;
thereafter, the extract was submitted to an additional off-line clean-up with a
SPE cartridge containing 0.5 g of PSA.
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
269
3.3. PLE parameters
3.3.1. Time, dichloromethane percentage and temperature
The influence of the above factors on the efficiency of the extraction step
was simultaneously investigated using a Box-Behnken experimental factorial
design with each variable considered at three levels, Table 2. The flush volume
was 100%, the pressure 1500 psi and two extraction cycles were applied. The
purified extracts were injected in the GC-MS system, operated in the SIM
acquisition mode, and peak areas were used as the response variable in order to
calculate the main effects associated with each experimental factor, their
quadratic terms and the two-factor interactions. Table 3 summarizes the
numerical values of the standardized main effects and their quadratic terms.
The absolute value of a main effect is proportional to the influence of the
associated factor on the efficiency of the PLE extraction. A positive sign
indicates an improvement in the yield of the process when the factor varies
from the low to the high level, within the domain of the design, and a negative
one the opposite trend.
Table 2. Experimental domain of the Box-Behnken design.
Factor Code Level
Low Medium High
Time (min)
CH2Cl2 (%)
Temperature (ºC)
A
B
C
2
5
40
6
22.5
65
10
40
90
Metodología desarrollada-Muestras sólidas
270
Table 3. Standardized main effects and quadratic terms provided by the Box-Behnken
design.
Compound Main effects Quadratic terms
A B C AA BB CC
EHS
HMS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OC
-0.42
-2.7a
-1.0
-0.37
-2.8a
-1.7
-2.2
-4.0a
-2.7a
-3.2a
-1.5
7.7a
-3.5a
-4.1a
-1.7
9.4a
0.03
-0.52
0.82
6.3a
-1.7
-0.60
0.96
-0.75
-2.0
-2.8a
-1.4
-0.96
-3.6a
-3.3a
-3.0a
-2.5
-3.7a
-3.6a
-3.3a
-7.6a
-4.3a
-2.5
-3.8a
-14.4a
-0.94
-1.0
-0.84
-1.1
-0.98
-0.78
-1.5
0.56
a Statistical significant factors and quadratic terms
Data summarized in Table 3 show that the percentage of
dichloromethane (code B) played a positive and statistical significant effect
(95% confidence level) in the extraction of BP-3 and OC, whereas the opposite
trend was observed for the rest of analytes. The temperature of the cell (code C)
affected positively to the extraction of BP-3 and the extraction time (code A)
showed a negative influence on the yield of the extraction, being statistically
significant for three (HMS, 4-MBC and OC) of the investigated species.
Quadratic terms associated with the extraction time (AA) and the percentage of
dichloromethane (BB) also presented statistically significant effects for many
compounds (Table 3). These data suggest a non-linear variation in the efficiency
of the extraction within the domain of the design. The main effect plots for
selected compounds confirmed that maximum yields were achieved at
intermediate extraction times and dichloromethane percentages, Fig. 2. Finally,
two-factor interactions remained below the statistical significance threshold,
data not shown.
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
271
2.0CH2Cl2 (%)
40.0 90.0
HMS
11
12
13
14
15
16
(x104)
Time (min)
10.0 5.0Temperature (ºC)
40.0
Pea
k ar
eaBP-3
14
19
24
29
34
39
44
(x104)
Pea
k ar
ea
2.0 40.0 90.010.0 5.0 40.0CH2Cl2 (%)Time (min) Temperature (ºC)
4-MBC
18
19
20
21
22
(x104)
Pea
k ar
ea
2.0 40.0 90.010.0 5.0 40.0CH2Cl2 (%)Time (min) Temperature (ºC)
OC
10
12
14
16
18
(x104)
Pea
k ar
ea
2.0 40.0 90.010.0 5.0 40.0CH2Cl2 (%)Time (min) Temperature (ºC)
Fig. 2. Main effect graphs provided by the experimental factorial design for selected
compounds.
The best compromise conditions, which maximized the efficiency of the
extraction for all analytes, were calculated with a global desirability (D)
function. D is defined as the geometric mean of the normalized (between 0 for
the minimum and 1 for the maximum) individual responses (di) predicted by
the Box-Behnken design for each UV-filter. The maximum value of D (0.89) was
obtained at 75 ºC, using a n-hexane:dichloromethane (80:20, v:v) mixture and
considering an extraction time of 5 min, Fig. 3.
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Extraction time, 5 min
5 10 15 20 25 30 35 40CH2Cl2 (%)
4050
6070
8090
0
0.2
0.4
0.6
0.8
1D
esir
abili
ty (
D)
Extraction temperature, 75 ºC
2 4 6 8 10Time (min)
1020
3040
0
0.2
0.4
0.6
0.8
1
Des
irab
ility
(D
)
Fig. 3. Plots of the global desirability function.
3.3.2. Extraction cycles, flush volume and purge time
The potential influence of these parameters on the efficiency of the
extraction was evaluated with an univariant approach. No differences were
observed using 1, 2 or 3 extraction cycles of 5 min. Thus, a single cycle was
considered to speed up the extraction step. Fig. 4 shows the responses (peak
areas) for three different flush percentages, referred to the volume of the PLE
cell (11 mL). Similar responses were measured for flush values of 100% and
140%, whereas a slight reduction was appreciated for several analytes using a
percentage of 60%. This factor was set at 100%. Operating under above
conditions (1 cycle and 100% flush), the volume of the extract collected from the
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
273
PLE cell remained around 20 mL. Purge times higher than 2 min were also
studied without significant changes in the extraction efficiency; thus, 2 min was
maintained as working value for this variable.
Pe
ak
are
a (
x 1
05 )
0
5
10
15
20
25
EHS HMS IAMC BP-3 4-MBC EHPABA EHMC OC
60% flush 100% flush 140% flush
Fig. 4. Effect of flush percentage on the responses obtained for a spiked (5 g g-1) sludge
sample, n=3 replicates.
3.4. Recoveries and quantification limits
The recoveries of the method were evaluated using two freeze-dried
samples of primary and biological sludge spiked at two different concentration
levels (300 and 1000 ng g-1). Non-spiked fractions of each matrix and procedural
blanks were also processed, Fig. 5. Found recoveries ranged from 73% to 112%,
with relative standard deviation values below 12%, Table 4. The above data are
similar to those reported by Plagellat et al. [14] for 4-MBC, EHMC and OC
using liquid-liquid extraction of wet sludge samples and Nieto et al. [18] for BP-
3, EHPABA and OC considering methanol:water mixtures for PLE of several
personal care compounds from freeze-dried sludge.
Metodología desarrollada-Muestras sólidas
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9.40 9.80 10.20 10.600
2000
4000
6000
8000
10000
12000
Time (min)
Abundancem/z 120
EHS
HMS
11.60 12.00 12.40 12.80 13.20 13.600
2000
4000
6000
8000
10000
12000
14000
Time (min)
Abundancem/z 178
IAMC(E) EHMC
(Z) EHMC
11.25 11.35 11.45 11.55 11.650
500
1000
1500
2000
2500
3000
3500
4000
Time (min)
Abundancem/z 227
BP-3
11.00 11.20 11.40 11.60 11.80 12.000
500
1000
1500
2000
2500
3000
3500
4000
4500
Time (min)
Abundancem/z 254
(E) 4-MBC
(Z) 4-MBC
15.80 16.00 16.20 16.40 16.600
2000400060008000
10000120001400016000180002000022000
Time (min)
Abundancem/z 360
OC
12.40 12.80 13.20 13.60 14.000
200400600800
1000120014001600180020002200
Time (min)
Abundancem/z 277
EHPABA
A
B
C
Fig. 5. Selected ion monitoring chromatograms corresponding to a procedural blank
(A), a non-spiked sample of biological sludge (B), and same sample fortified with 300 ng
g-1 of each analyte (C).
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
275
Table 4. Recoveries of the method for spiked samples (n=3 replicates) and estimated
limits of quantification (LOQs) of the method.
Analyte Primary sludge (TC 30%) Biological sludge TC (35%) LOQs (ng g-1)
a300 ng g-1 a1000 ng g-1 a300 ng g-1 a1000 ng g-1
EHS
HMS
IAMC
BP-3
4-MBC
EHPABA
EHMC
OC
101 7
96 6
107 6
89 11
86 5
93 7
90 5
85 5
95 7
78 5
80 4
106 6
79 4
88 6
73 5
84 12
102 8
103 1
90 4
112 4
91 7
83 7
90 9
112 5
103 3
100 7
98 6
100 4
107 3
104 3
88 7
98 8
17
34
34
61
26
22
24
33
a Added concentration
The reproducibility of the method was investigated with a sample of
biological sludge fortified at 500 ng mL-1. The relative standard deviations
(RSDs, %) for nine extractions in three consecutive days varied between 6 and
13%.
As shown in Fig. 5, analytes were not detected in the procedural blanks;
therefore, the LOQs of the method (defined for a S/N of 10) were estimated
from chromatographic peaks of UV filters in non-spiked samples, or in the low
level spiked fraction for those species not detected in sludge (IAMC and
EHPABA). The achieved LOQs varied between 17 ng g-1 for EHS and 61 ng g-1
for BP-3, Table 4. They are similar to the LOQs (from 7 to 67 ng g-1) reported for
same compounds in sediment samples with TC below 0.2%, using GC-MS as
detection technique and a less elaborated clean-up procedure [16]. Plagellat and
co-workers [14] achieved LOQs between 9 and 18 ng g-1 for 4-MBC, EHMC and
OC considering a three times larger sample intake (60 g of fresh sludge at 3%)
and using also GC-MS as determination technique.
Metodología desarrollada-Muestras sólidas
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3.5. Application to real samples
The proposed method was applied to freeze-dried sludge from different
urban sewage plants. EHPABA and IAMC were not detected in any of the
processed samples. The concentrations measured for the rest of species are
compiled in Table 5. 4-MBC and EHMC appear in sludge as mixtures of E and
Z forms. The sum of peak areas for both isomers was compared with calibration
curves obtained for the commercial available E forms. Samples code 1 and 2, on
Table 5, corresponded to wet sludge from a plant receiving the wastewater
from a 100000 inhabitants city, located in the Northwest of Spain. Both samples
were obtained in March of 2010 and lyophilized in our laboratory. The rest of
specimens (codes 3 to 9) were from STPs in the same geographical area,
although their exact locations are not revealed due to a confidentiality
agreement. They were collected between February and May 2010, in-situ
lyophilized and further submitted to the laboratory for analysis. Samples code 5
to 9 (Table 5) are mixtures of primary and biological sludge.
Table 5. Summary of concentrations (ng g-1) measured in sludge samples, n=3
replicates. IAMC and EHPABA were not detected in any sample.
Code Type Concentration (ng g-1) SD
EHS HMS BP-3 4-MBC EHMC OC
1
2
3
4
5
6
7
8
9
Primary
Biological
Primary
Biological
Mixture
Mixture
Mixture
Mixture
Mixture
n.d.
270 14
n.d.
133 26
298 5
n.d.
200 36
188 9
268 11
n.d.
207 31
n.d.
110 10
401 35
n.d.
240 8
256 18
180 ± 27
n.d.
n.d.
n.d.
93 11
n.d.
n.d.
n.d.
n.d.
295 14
1543 26
1439 49
106 5
97 8
223 9
120 3
372 10
351 41
1579 51
3287 98
856 98
213 3
104 5
160 7
192 15
125 5
100 10
2776 137
2242 16
3263 176
1039 50
377 30
1766 72
1038 63
1934 222
523 58
2240 45
Meana 226 232 194 648 868 1602
aAverage value of quantified concentrations
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
277
4-MBC, EHMC and OC were ubiquitous pollutants in sludge, with
average concentrations increasing the following order: 4-MBC< EHMC < OC,
Table 5. The mean levels of 4-MBC and OC were lower than those reported for
sewage sludge samples collected in Switzerland; however, a higher value was
obtained for EHMC [14]. EHS and HMS were quantified in six of nine samples
with maximum values below 400 ng g-1, and BP-3 showed a lower detection
frequency, Table 5. Globally, the occurrence frequency and the relative
concentrations of UV filters in sludge followed the same pattern as in water
samples taken in the same geographic area [13]. Moreover, they are in
agreement with the high sorption coefficients reported for 4-MBC, EHMC and
OC in sludge [6].
4. Conclusions
PLE extraction, combined with the use of graphitized carbon for in-cell
retention of pigments and additional clean-up with a PSA cartridge, constitutes
a suitable approach in terms of extraction efficiency and selectivity for the GC-
MS determination of a broad group of UV filters in sludge samples. As far as
we could trace, this study reports the first application of both materials for the
clean-up of PLE extracts from sludge samples, achieving an improved
selectivity in comparison with the commonly used normal-phase sorbents. The
analysis of sludge samples confirmed the significant accumulation of three UV
filters (4-MBC, EHMC and OC) in this matrix, with average concentrations
higher than 600 ng g-1. This information must be considered in order to (1)
properly calculate their removal rates during wastewater treatments and (2) to
evaluate the risk of re-introducing the above species in the terrestrial
environment through the disposal of sludge as fertilizer in agriculture fields.
Acknowledgements
Financial support from the Spanish Government and E.U. FEDER funds
(project CTQ2009-08377) is acknowledged. N.N. is grateful for a FPU grant from
Metodología desarrollada-Muestras sólidas
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the Spanish Ministry of Education and Science. We are also in debt with Dr.
Carballa and Labaqua for supplying, or providing access, the sludge samples.
Journal of Chromatography A, in press (doi:10.1016/j.chroma.2010.11.028)
279
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[12] R. Rodil, M. Moeder, J. Chromatogr. A 1179 (2008) 81.
[13] N. Negreira, I. Rodriguez, E. Rubí, R. Cela, Anal. Bioanal. Chem. 398 (2010) 995.
[14] C. Plagellat, T. Kupper, R. Furrer, L.F. de Alencastro, D. Grandjean, J. Tarradellas,
Chemosphere 62 (2006) 915.
[15] A. Nieto, F. Borrull, E. Pocurull, R.M. Marcé, Trends Anal. Chem. 29 (2010) 752.
[16] R. Rodil, M. Moeder, Anal. Chim. Acta 612 (2008) 152.
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2.4. Publicación:
DETERMINATION OF
SELECTED UV FILTERS IN INDOOR DUST
USING MATRIZ SOLID-PHASE DISPERSION
AND GAS CHROMATOGRAPHY
TANDEM MASS SPECTROMETRY
N. Negreira, I. Rodríguez, E. Rubí, R. Cela
Journal of Chromatography A 1216 (2009) 5895
(doi:10.1016/j.chroma.2009.06.020)
Journal of Chromatography A 1216 (2009) 5895
283
Determination of selected UV filters in indoor dust using matrix solid-phase
dispersion and gas chromatography tandem mass spectrometry
N. Negreira, I. Rodríguez*, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto de
Investigación y Análisis Alimentario, Universidad de Santiago de Compostela,
Santiago de Compostela 15782, Spain.
Abstract
A simple and inexpensive sample preparation procedure, based on the
matrix solid-phase dispersion (MSPD) technique, for the determination of six
UV filters: 2-ethylhexyl salicylate (EHS), 3,3,5-trimethylcyclohexyl salicylate
(Homosalate, HMS), 3-(4-methylbenzylidene) camphor (4-MBC), isoamyl-p-
methoxycinnamate (IAMC), 2-ethylhexyl-p-methoxycinnamate (EHMC) and
octocrylene (OCR) in dust from indoor environments is presented and the
influence of several operational parameters on the performance of the
extraction discussed. Under final working conditions, sieved samples (0.5 g)
were mixed with the same amount of anhydrous sodium sulfate and dispersed
with 2 g of C18 in a mortar with a pestle. This blend was transferred to
polypropylene solid-phase extraction cartridge containing 2 g of activated silica,
as clean-up co-sorbent. The cartridge was first rinsed with 5 mL of n-hexane
and then analytes were recovered with 4 mL of acetonitrile. This extract was
adjusted to 1 mL, filtered and compounds were determined by gas
chromatography combined with tandem mass spectrometry (GC-MS/MS).
Recoveries for samples spiked at two different concentrations ranged from 77 to
99% and the limits of quantification (LOQs) of the method remained between 10
and 40 ng g-1. Analysis of settled dust from different indoor areas, including
private flats, public buildings and vehicle cabinets, showed the ubiquitous of
EHMC and OCR in this matrix, with maximum concentrations of 15 and 41 g
g-1, respectively. Both UV filters were also quantified in dust reference material
Metodología desarrollada-Muestras sólidas
284
SRM 2585 for first time. EHS, 4-MBC and IAMC were detected in some of the
analyzed samples, although at lower concentrations than EHMC and OCR.
Keywords: UV filters; dust; indoor atmospheres; matrix solid-phase
dispersion; gas chromatography tandem mass spectrometry.
1. Introduction
UV filters are compounds designed to mitigate the deleterious effects of
sunlight. Most of them are organic substances, characterized by single or
multiple aromatic structures, often with attached hydrophobic groups [1]. One
of their most known uses is in sunscreens, which are applied directly on the
skin as protection against UV radiation. The number and maximum allowable
concentrations of UV filters in these products have been legislated in many
countries. As example, in the European Union 26 organic compounds have been
approved to be incorporated in sunscreens at individual concentrations up to
10%, for most of them [2,3]. Moreover, they are also included in the formulation
of other personal care products (cosmetics, hair dyes, shampoos) and used in
the protection of goods, plastics, varnishes and clothes [4-7]; however, no data
could be traced related to the type of UV absorbers and concentrations
incorporated in these materials.
As many other daily usage compounds, UV filters are continuously
discharged in the aquatic environment. Washing off from the skin, during
bathing or swimming, and indirect releases from towels and clothes contribute
significantly to their presence in surface and wastewater samples [8]. Their
further behavior in the aquatic media depends on a number of factors such as
(1) their stability during wastewater treatments [4,9-10], (2) physicochemical
properties, particularly their polarity, and (3) potential transformation through
photochemical and/or oxidation reactions [11,12]. Medium and highly polar
UV filters (e.g. 2-hydroxy-4-methoxybenzophenone, BP-3, and 2-hydroxy-4-
methoxybenzophenone-5-sulphonic acid, BP-4) have been often detected in
Journal of Chromatography A 1216 (2009) 5895
285
surface and wastewater [6,7,13]; moreover, BP-4 is not effectively removed
during conventional sewage treatments [6]. More lipophilic species, such as 3-
(4-methylbenzylidene) camphor (4-MBC), 2-ethylhexyl-p-methoxycinnamate
(EHMC) and octocrylene (OCR) have been found in river and lake sediments
[14], sludge [15] and even in biota [4,16]. In fact, concentrations up to 2 g g-1
have been reported for 4-MBC and OCR in fish [16] and as high as 15 g g-1 for
OCR in sludge [15], suggesting that certain UV filters can be bio-accumulated
and concentrated on solid matrices. This information added to reports about the
estrogenic activity of some species, such as 4-MBC and EHMC [17-21], have
risen the concern about their effects on wildlife and humans.
Conversely to above studies dealing with environmental samples, as far
as we could trace, no data are available in relation to the levels of UV filters in
indoor areas. Breath and oral intake of suspended particulate matter and settled
dust are considered as sources of continuous exposure to organic and inorganic
chemicals used in building materials and daily activities [22,23]. Most
compounds present an excellent stability in indoor and confined areas, since
removal through photochemical reactions as well as airborne dispersion have a
minor importance. Some of these species might contribute to the increasing
incidence of allergies, asthma and other respiratory diseases in developed
countries, where citizens spend more and more time in indoor areas [24]. In this
sense, recent studies have shown that certain personal care products reach
higher concentrations in dust than in other heavily polluted environmental
matrices, such as sludge [25,26]. The same trend has been observed for other
compounds added to building materials and furniture, e.g. organophosphorous
[27] and polybrominated flame retardants [28].
The goal of this work was to develop a procedure for the determination
of a group of selected UV filters: EHS, HMS, isoamyl p-methoxycinnamate
(IAMC), 4-MBC, EHMC, and OCR, in dust from different indoor environments.
Matrix solid-phase dispersion (MSPD) was considered as sample preparation
Metodología desarrollada-Muestras sólidas
286
technique on the basis of its low cost, use of mild conditions and feasibility to
integrate extraction and clean-up in the same step, which can be achieved with
a suitable combination of elution solvents, dispersant and clean-up co-sorbents
[29,30]. Although MSPD was initially centered on biota samples [31], recent
works have proven its applicability to the extraction of organic compounds
from dust [25,32] and other solid matrices, such as freeze-dried sludge [33,34].
After extraction, analytes were determined by gas chromatography in
combination with tandem mass spectrometry (GC-MS/MS). The second goal of
this work was to provide a first overview of the levels for some UV filters in
dust samples collected from different confined environments, including a
reference material of indoor dust.
2. Experimental
2.1. Standards and material
HPLC-grade methanol, acetone, n-hexane, dichloromethane, ethyl
acetate and acetonitrile (trace analysis grade) were supplied by Merck
(Darmstadt, Germany). Standards of EHS, HMS, 4-MBC, EHMC and OCR were
acquired from Aldrich (Milwaukee, WI, USA) and Merck. IAMC was kindly
provided by Dr. R. Rodil (University of La Coruña, Spain). Chemical structures
and octanol-water partition coefficients (log Kow) of above compounds are
depicted in Fig. 1. As observed, all of them show moderate to high lipophilic
character with the subsequent risk of being adsorbed on dust particles.
Individual solutions of each compound were prepared in methanol. Further
dilutions and mixtures of them were made in acetone and acetonitrile.
Journal of Chromatography A 1216 (2009) 5895
287
EHS
log Kow 5.77
HMS
log Kow 5.82
IAMC
log Kow 4.06
4-MBC
log Kow 4.95
EHMC
log Kow 5.66
OCR
log Kow 7.53
Fig. 1. Chemical structures and octanol-water partition coefficients (Log Kow) of selected
UV filters.
Anhydrous sodium sulphate, Florisil (60-100 mesh), C18 (70-230 mesh),
silica (230-400 mesh) and alumina (150 mesh) sorbents, used in the MSPD
extraction process, were acquired from Aldrich and Merck. Normal-phase
materials were activated at 130 ºC, for 24 h, and then allowed to cool down in a
desiccator before being used in the extraction process. C18 was used as
received. Polypropylene solid-phase extraction cartridges (15 mL capacity) and
20 µm polyethylene frits were purchased from International Sorbent
Technology (Mid Glamorgan, UK). Syringe filters (Millex GV, 13 mm, 0.22 µm)
were obtained from Millipore (Billerica, MA, USA).
Metodología desarrollada-Muestras sólidas
288
2.2. Samples
Dust from private houses, public buildings and vehicles cabins were
collected using domestic vacuum cleaners equipped with paper filter bags.
After opening, the content of bags was sieved and the fraction with a particle
size below 60 µm was retained for this study. Sieved samples were stored at 4
ºC, in amber glass vessels and their total carbon (TC) content characterized.
Obtained TC values ranged from 13.5 to 33.6%. Reference material SRM 2585,
organic compounds in house dust, was acquired from NIST (Gaithersburg, MD,
USA). MSPD extraction conditions were optimised using a pooled dust matrix
with a TC of 20.8%. Fractions of this sample were fortified with different UV
filters. More details about added compounds and their concentrations are given
in the results and discussion section. During method validation, individual dust
samples were also spiked with all compounds at different levels. In all cases,
the spiking procedure consisted of mixing the sieved samples with a standard
solution of UV filters in acetone. Approximately, 1 mL of standard was added
per g of dust in order to obtain an homogeneous slurry, which was left at room
temperature until complete evaporation of the solvent. After that, spiked
samples were aged, at 4 ºC, for at least 2 weeks before extraction.
2.3. Sample preparation
Dust (0.5 g) was mixed with anhydrous sodium sulphate, 0.5 g, and
dispersed with C18 in a glass mortar, with a pestle, until a visually
homogeneous blend was obtained. Then, it was transferred to a cartridge
containing a polyethylene frit and a given amount of co-sorbent. A second frit
was placed over the dispersed sample before slight compression. Cartridges
were eluted by gravity. Under optimised conditions, 2 g of C18 and the same
mass of silica were used as dispersant and co-sorbent, respectively. MSPD
cartridges were first rinsed with 5 mL of n-hexane to remove organic
compounds with a lower polarity than analytes. Then, analytes were recovered
using just 4 mL of acetonitrile.
Journal of Chromatography A 1216 (2009) 5895
289
2.4. Determination
Analytes were determined by GC-MS/MS using a Varian (Walnut Creek,
CA, USA) CP 3900 gas chromatograph connected to an ion-trap mass
spectrometer (Varian Saturn 2100). Separations were carried out in a HP-5ms
capillary column (30 m x 0.25 mm i.d., df: 0.25 m) supplied by Agilent
(Wilmington, DE, USA). Helium (99.999 %) was used as carrier gas at a constant
flow of 1.2 mL min-1. The GC oven was programmed as follows: 70 ºC (held for
1 min), at 12 ºC min-1 to 280 ºC (held for 10 min). GC-MS interface and ion-trap
temperatures were set at 280 ºC and 220 ºC, respectively. Standards and sample
extracts in acetonitrile (1-2 L) were injected in the splitless mode (splitless time
1 min), with the injector port at 280 ºC. The mass spectrometer was operated in
the electron impact ionisation mode (70 eV). MS spectra were recorded in the
range from 70 to 400 m/z units. The base peak in the spectra of each compound
was isolated with a window of 3 m/z units and subjected to collision induced
dissociation. GC-MS and GC-MS/MS were considered as detection techniques
during the development of this work. Concentrations of target species in non-
spiked samples and evaluation of method performance were done by GC-
MS/MS, considering external calibration, against standards prepared in
acetonitrile, as quantification technique.
3. Results and discussion
3.1. GC-MS/MS conditions
Fig. 2 shows the MS/MS spectra obtained for considered compounds
and the structures proposed for their most intense product ions. In the case of
salicylates (EHS and HMS), the parent ion (138 m/z), corresponding to
replacement of the aliphatic chain bonded to the ester moiety by hydrogen,
underwent a removal of water leading to a single product at 120 m/z units. The
precursor ion in the MS spectra of both cinnamates (178 m/z units), which
reflected also the substitution of the branched alkyl chain attached to the ester
moiety by hydrogen, rendered two main transitions corresponding to the
removal of the hydroxyl and carbonyl groups (178 > 161 and 161 > 133 m/z
Metodología desarrollada-Muestras sólidas
290
units, respectively). The main transition from the [M-H]+ OCR ion, appearing at
360 m/z units, reflected the loss of organic chains (butyl and ethyl groups)
bonded to the tertiary carbon, Fig. 2. Finally, the MS/MS spectra of 4-MBC
showed multiple transitions, the most intense one, indicating the loss of a
methyl group (254 > 239 m/z units), was used for quantitative purposes. Table
1 summarizes optimal MS/MS detection parameters, as well as correlation
coefficients (R2) values and instrumental limits of quantification (LOQs),
defined for a signal to noise ratio (S/N) of 10, corresponding the injection of
standards prepared in acetonitrile. In comparison to the single MS mode,
MS/MS detection provided around twice lower LOQs values, except for 4-
MBC. Retention times given in Table 1 for 4-MBC and both cinnamates
correspond to their E-isomers, which are the forms incorporated in sunscreens
and other personal care products. It must be kept on mind that Z-isomers,
detected in some environmental samples due to photochemical isomerisation
[14,15], appear at lower retention times; thus, the multiple reaction monitoring
(MRM) mode was used to record the MS/MS transitions of 4-MBC, IAMC and
EHMC in the same segment.
Journal of Chromatography A 1216 (2009) 5895
291
225 250 275 300 325 350 m/z
0%
25%
50%
75%
100%
232246258
264
276
288304
318
330
342
360
OCR
m/z 276
100 125 150 175 200 225 m/z
0%
25%
50%
75%
100%
103
121
133
149
161
178
EHMC, IAMC
m/z 133
MeO
(E)
O
m/z 161
125 150 175 200 225 250 m/z
0%
25%
50%
75%
100%
132
149
162
170
183
197
211
226
239
254
4-MBC
m/z 239
100 150 200 250 300 350 m/z
0%
25%
50%
75%
100%120
138
HMS, EHS
m/z 120
Fig. 2. MS/MS spectra and structures proposed for most intense product ions.
Metodología desarrollada-Muestras sólidas
292
Tabl
e 1.
Ret
enti
on ti
mes
, MS/
MS
oper
atio
nal c
ondi
tion
s an
d pe
rfor
man
ce o
f GC
-MS/
MS
and
GC
-MS
for
stan
dard
s of
UV
filte
rs p
repa
red
in a
ceto
nitr
ile. L
OQ
s va
lues
def
ined
for
a S/
N r
atio
of 1
0.
Com
poun
d
Ret
. tim
e
(min
)
Pare
nt io
n
(m/z
)
Prod
uct i
ons
(m/z
)
Exc
itati
on
ampl
itud
e (V
) St
orag
e le
vel (
m/z
) C
orre
lati
on
Coe
ffic
ient
(R2 )
b
LO
Qs
(ng
mL
-1)b
Cor
rela
tion
Coe
ffic
ient
(R2 )
c
LOQ
s
(ng
mL
-1)c
EH
S
HM
S
IAM
C
4-M
BC
EH
MC
OC
R
12.7
13.3
,13.
5
14.7
a
14.9
a
16.8
a
19.2
138
138
178
254
178
360
120
120
161+
133
239
161+
133
276
0.45
0.45
0.91
0.82
0.91
1.29
53
53
68
97
68
137
0.99
9
0.99
8
0.99
9
0.99
7
0.99
9
0.99
9
5 6 5 20
5 6
0.99
6
0.99
7
0.99
6
0.99
6
0.99
3
0.99
9
11
10
15
17
15
12
a Ret
enti
on ti
mes
cor
resp
ondi
ng to
E-i
som
ers.
b Dat
a fo
r G
C-M
S/M
S.
c Dat
a fo
r G
C-M
S.
b,c V
alue
s ob
tain
ed fo
r st
and
ard
s at
six
leve
ls b
etw
een
LO
Qs
and
3
g m
L-1 .
Journal of Chromatography A 1216 (2009) 5895
293
3.2. Optimisation of MSPD parameters
Preliminary extraction assays were performed with fractions of a pooled
dust sample (TC 20.8%) spiked with all compounds at 3 g g-1, except IAMC,
which was not considered in the earliest steps of this research. On the basis of a
previous application of MSPD to the extraction of bactericides from dust [25],
samples (0.5 g) were mixed with 0.5 g of sodium sulphate and dispersed using
1.25 g of C18 in a mortar with a pestle. The resulting blend was transferred to a
polypropylene cartridge containing 2 g of different normal-phase materials,
which worked as clean up co-sorbents retaining species with a higher polarity
than target UV filters through adsorption processes. N-hexane,
dichloromethane, ethyl acetate and acetonitrile were tested as extraction
solvents. In all cases, 5 mL of solvent were recovered from the MSPD cartridge
and analyzed by GC using single MS detection. None of the compounds could
be eluted with n-hexane. 4-MBC, EHMC and OCR were noticed in
dichloromethane extracts; however, this solvent failed to recovered EHS and
HMS, figure not shown. Probably, the presence of a phenolic group in the
structure of these two species (Fig. 1) resulted in a stronger interaction with the
dust matrix and/or with the normal-phase co-sorbent than the rest of analytes.
Responses (peak areas) obtained using ethyl acetate and acetonitrile are plotted
in Fig. 3. Whatever the type of co-sorbent, for EHMC, OCR and, in a lesser
extension, 4-MBC acetonitrile provided higher responses than ethyl acetate. For
EHS and HMS, the co-sorbent played a more important effect on the extraction
yield than the elution solvent. Both salicylates were strongly retained by
alumina, particularly when combined with acetonitrile as eluent. On the basis
of these results, Florisil, silica, ethyl acetate and acetonitrile were selected as co-
sorbents and elution solvents, respectively, for further experiments. Extraction
of non-spiked fractions of the pooled dust sample, using above conditions,
demonstrated the existence of noticeable concentrations of OCR and EHMC in
this matrix, figure not shown.
Metodología desarrollada-Muestras sólidas
294
0,E+00
4,E+05
8,E+05
1,E+06
2,E+06
2,E+06
EHS HMS 4-MBC EHMC OCR
Pe
ak
are
aAlumina ACN Florisil-ACN Silica-ACN Alumina-AcOEt Florisil-AcOEt Silica-AcOEt-
Fig. 3. Comparison of responses for different combinations of elution solvents and co-
sorbents in the MSPD process, n=3 replicates. Data for 0.5 g of spiked dust dispersed
with 1.25 g of C18.
In a second series of extractions, the effects of co-sorbent, elution solvent
and mass of dispersant (C18), on the efficiency of the MSPD process, were
simultaneously assessed using a 23 type experimental factorial design, Table 2.
The mass of sample was 0.5 g and the volume of extraction solvent 10 mL in all
experiments; moreover, MSPD cartridges were rinsed with 5 mL of n-hexane to
remove less polar interferences, previously to the extraction of target species.
Obviously, n-hexane extracts were discarded. The study was accomplished
with a different fraction of the same pooled matrix used in the preliminary
extraction experiments. In this case, it was fortified with EHS, HMS, 4-MBC and
IAMC (3 g g-1, each). Extracts were filtered and processed directly using GC-
MS as detection technique. Peak areas obtained for above species, as well as
EHMC and OCR (already present in the sample) were considered as variable
response in the experimental factorial design. For EHMC the sum of responses
for E and Z isomers was taken. On the other hand, for IAMC and 4-MBC just
one peak corresponding to the isomer added to the sample (E-form) was
Journal of Chromatography A 1216 (2009) 5895
295
noticed in the GC-MS chromatograms. This fact suggests the absence of
isomerisation reactions during aging of the spiked sample and through the
extraction process. Standardized main effects associated to each factor are
reported in Table 2. Their absolute values are proportional to the variation in
the efficiency of the extraction when the considered factor changes from the low
to the high level, within the domain of the design. A positive sign indicates an
increase in the yield of the process and, a negative one the opposite behaviour.
Data on Table 2 suggest to different trends. On one hand, for 4-MBC, OCR,
IAMC and EHMC, the extraction solvent was the only factor playing a
significant influence (95% confidence level) on the yield of the extraction, with
the highest responses corresponding to acetonitrile. On the other hand, EHS
and HMS followed a different pattern. In this case, the most relevant factor was
the type of co-sorbent placed at the bottom of the extraction cartridge. It seemed
that both species were strongly retained on Florisil, thus lower responses were
observed for this material than for silica. A similar trend has been reported by
Rodil and co-workers [14] in the extraction of hydroxylated UV filters from
sediments using pressurized solvents. Moreover, the extraction of EHS and
HMS was favoured significantly by the use of large masses of C-18 as
dispersant. Regarding the elution solvent, HMS was more efficiently recovered
with acetonitrile, whereas EHS presented a higher affinity for ethyl acetate,
although without achieving the statistically significance boundary. In general,
two factor interactions exerted a minor influence on the yield of the extraction
process, figure not shown. As regards the effect of extraction parameters on the
complexity of the total ionic current (TIC) GC-MS chromatograms, little
changes were noticed among different conditions explored in the experimental
design. The most intense chromatographic peaks corresponded to phthalates,
although they appeared at different retention times than target UV filters. On
the basis of above comments, acetonitrile and silica were selected as extraction
solvent and co-sorbent in the MSPD process, respectively; whereas, the mass of
dispersant was fixed at 2 g.
Metodología desarrollada-Muestras sólidas
296
Table 2. Domain and standardized main effects of factors considered in the experimental
factorial design.
Factor Level Standardized value
Low High EHS HMS 4-MBC IAMC EHMC OCR
Solvent
Co-sorbent
C-18 mass (g)
Ethyl acetate
Florisil
0.5
Acetonitrile
Silica
2.0
-5.8
18a
9.1
19a
88a
42a
26a
0.06
3.8
15a
1.2
3.2
17a
0.96
1.4
12a
0.11
0.15
a Statistically significant factors at the 95% confidence level
The following parameter to optimise in the extraction process was the
volume of acetonitrile. The study was carried out by collecting consecutive
fractions of 2 mL from the MSPD cartridge. Normalised responses for duplicate
experiments are shown in Fig. 4. As observed, compounds could be eluted
using just 4 mL of acetonitrile. Thus, the whole sample preparation method
required a total of 9 mL of two different organic solvents, a volume significantly
smaller than that reported for the extraction of same compounds from other
solid matrices with lower carbon contents, such as sediments [14]. Evaporation
of acetonitrile extracts to a final volume of 1 mL, under mild conditions (room
temperature using a gentle stream of nitrogen), did not lead to noticeable losses
of any compound, thus this step was incorporated in the sample preparation
scheme.
Journal of Chromatography A 1216 (2009) 5895
297
0%
20%
40%
60%
80%
100%
120%
EHS HMS IAMC 4-MBC EHMC OCR
Compound
No
rmal
ised
res
po
nse
Fraction 1 Fraction 2 Fraction 3 Fraction 4
Fig. 4. Normalised signals for consecutive acetonitrile fractions (2 mL) eluted from the
MSPD cartridge.
3.3. Performance evaluation
Recoveries of the optimized method were estimated with dust spiked at
two different concentration levels: 0.3 and 3 g g-1, per compound. This study
was carried out with a sample, obtained from a public building (TC 13.5%),
containing relatively low levels of target species. After sieving, it was divided in
three fractions, one was used as blank and the others spiked at the above
referred levels and aged for 2 weeks. Each fraction was processed (n=4
replicates) and the corresponding extracts concentrated to 1 mL. Table 3 shows
the recoveries obtained using GC-MS and GC-MS/MS as detection techniques.
The first led to values over 100% for some analytes in the sample spiked at the
lower level, probably due the co-elution of other compounds which contributed
to peak areas of UV filters recorded in chromatograms monitored for the most
intense m/z ions in their MS spectra (see 3rd column on Table 1), this drawback
was overcome using MS/MS detection, which was chosen as quantification
technique for the analysis of non-spiked samples. Globally, the proposed
Metodología desarrollada-Muestras sólidas
298
method provided recoveries between 77 and 99%, with associated standard
deviations below 10, for all species.
Table 3. Summary of recoveries obtained for a spiked dust sample, n=4 replicates, and
LOQs of the method defined for a S/N ratio of 10.
Compound
Percentage of recovery (%) ± SD LOQs (ng/g)b
GC-MS GC-MS/MS
300 ng g-1a 3000 ng g-1a 300 ng g-1a 3000 ng g-1a
EHS
HMS
IAMC
4-MBC
EHMC
OCR
89 ± 3
91 ± 4
115 ± 6
73 ± 10
120 ± 5
109 ± 2
82 ± 5
81 ± 5
90 ± 3
97 ± 6
100 ± 5
96 ± 5
92 ± 2
87 ± 4
85 ± 4
99 ± 9
93 ± 4
79 ± 1
81 ± 6
83 ± 2
80 ± 4
77 ± 7
89 ± 5
77 ± 5
10
12
10
40
10
12
a Added concentration
b Referred to GC-MS/MS detection
Considering a sample intake of 0.5 g and adjusting the volume of the
final extract to 1 mL, the limits of quantification for the proposed method,
defined as the concentration of analyte which produced a signal to noise ratio of
10, ranged from 10 to 40 ng g-1, Table 3. Analysis of procedural blanks
demonstrated the absence of contamination problems (Fig. 5), thus above
values were mainly controlled by the instrumental LOQs of the GC-MS/MS.
Journal of Chromatography A 1216 (2009) 5895
299
Procedural blankUn-spiked sampleSpiked at 300 ng g-1
13.0 13.5 min.
0.0
2.5
5.0
7.5
10.0
12.5
15.0
m/z 120
EHS
HMS
kCounts
19.0 19.5 20.0 min.
0.0
2.5
5.0
7.5
m/z 276
OCR
kCounts
14.5 15.0 15.5 min.
0.0
2.5
5.0
7.5
10.0m/z 239
4-MBC
kCounts
14 15 16 17 min.
0
25
50
75
100
125
kCounts
m/z 161+133
E-IAMC
E-EHMC
Z-EHMC
Fig. 5. Overlay of GC-MS/MS chromatograms for a procedural blank, a non-spiked
dust sample (code 1, Table 4) and the same matrix fortified at 300 ng g-1.
3.4. Real samples analysis
The proposed method was applied to the analysis of several dust
samples collected in confined environments, from two different geographic
areas: Galicia (Northwest Spain) and La Rioja (North Spain). Obtained
concentrations are given in Table 4. Codes 1 to 8 correspond to dust from
private flats and public building, samples 9-10 were obtained from vehicles
Metodología desarrollada-Muestras sólidas
300
cabins and code 11 corresponds to SRM 2585. HMS remained below the LOQs
of the method in all samples, whereas the rest of species were noticed at
different levels from 35 ng g-1 up to 41 g g-1, depending on the compound and
the sample. In some cases, extracts needed to be diluted up to 10 times to fit
within the linear response range of the optimised method (up to 3 g mL-1
referred to the acetonitrile extract). Data reported on Table 4 for IAMC, 4-MBC
and EHMC corresponded to the sum of E and Z isomers. For the first two
species, the ratio between peak areas of Z and E forms remained under 0.05;
however, EHMC showed a higher percentage of isomerisation with (Z/E) ratios
from 0.4 to 1, depending on the sample. In general, the highest UV filters
concentrations corresponded to EHMC and OCR, which were found in
practically all the processed samples. The average concentration of OCR in dust
(11.4 g g-1) was twice higher than that reported by Plagellat and co-workers in
freeze-dried sludge [15]. As regards EHMC, from the best of our knowledge,
these are the highest levels (up to 15 g g-1) ever found in any environmental or
biota sample. Concentrations of both UV filters, OCR and EHMC, in reference
material SRM 2585 (code 11, Table 4) are in the same range than the certified
values of BDEs 99 and 209 in this sample [28]. As this reference material was
prepared from dust samples obtained in North America, a different geographic
region than that for the rest of samples considered in this study, the presence of
high levels of EHMC and OCR in dust seems to be a world wide reality. EHS,
IAMC and 4-MBC were also found in some of the processed samples; however,
their occurrence frequency and found concentrations were lower than those
corresponding to EHMC and OCR.
Journal of Chromatography A 1216 (2009) 5895
301
Table 4. Concentrations of UV filters in non-spiked dust samples, n=3 replicates.
Code TC (%) Concentrations (ng g-1) with their standard deviations
EHS IAMC 4-MBC EHMC OCR
1
2
3
4
5
6
7
8
9
10
11ª
13.5
21.7
33.6
22.4
18.6
17.9
20.2
26.5
15.7
15.1
23.3
35 (1)
650 (20)
1440 (50)
2747 (104)
n.d.
144 (14)
114 (17)
n.d.
114 (13)
n.d.
n.d.
n.d.
1290 (60)
n.d.
n.d.
n.d.
n.d.
58 (1)
92 (1)
n.d.
n.d.
n.d.
n.d.
700 (50)
1990 (240)
n.d.
n.d.
n.d.
n.d.
116 (52)
n.d.
n.d.
n.d.
560 (40)
15000 (500)
2400 (30)
1680 (34)
1460 (20)
2050 (80)
1080 (30)
6220 (330)
6570 (30)
177(12)
6460 (150)
700 (40)
7482 (12)
19300 (760)
34400 (3000)
458 (8)
1400 (300)
n.d.
41000 (1100)
7700 (300)
450 (50)
880 (30)
Meanb 750 480 935 3970 11400
a SRM 2585 b Average of values over the LOQs of the method n.d. below detection limits
Likely, different sources are responsible for the presence of UV filters in
indoor dust. On one hand, personal care products contribute directly (sprays)
and indirectly (accidental spillage, volatilisation, skin cells) to the levels of UV
filters in confined areas; moreover, furniture, upholstery, paints and polymeric
materials used in vehicles cabins, flats and public buildings, as well as clothes
lint, might content some of the investigated compounds as UV absorbers.
Metodología desarrollada-Muestras sólidas
302
4. Conclusions
MSPD followed by GC-MS/MS constitutes a fast, straightforward
approach for the determination of six UV filters in dust samples. The proposed
sample preparation method is particularly attractive since it does not require
the acquisition of dedicated instrumentation and it consumes low volumes of
organic solvents, providing recoveries over to 77% for all compounds and a
ready to inject extract. The performance of the extraction was mainly controlled
by the elution solvent and the type of co-sorbent placed at the bottom of the
MSPD cartridge. Results obtained for dust samples from different confined
environments demonstrated the presence of several UV filters in this matrix.
Particularly, mean values in the g g-1 range were detected for two of the
investigated species: EHMC and OCR. To the best of our knowledge, this is the
first report of both compounds in indoor areas. Further studies are necessary in
order to establish whether they proceed just from personal care products or
they might diffuse out from clothes, upholstery and building materials used in
indoor areas.
Acknowledgments
This study has been supported by Spanish Government, Xunta de
Galicia, and E.U. FEDER funds (projects DGICT CTQ2006-03334 and
PGIDIT06PXIB237039PR). N. N thanks a FPU grant to the Spanish Ministry of
Science and Innovation. We also acknowledge Dr. P. Canosa for the supply of
most dust samples.
Journal of Chromatography A 1216 (2009) 5895
303
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[17] H. Klammer, C. Schlecht, W. Wuttke, C. Schmutzler, I. Gotthardt, J. Köhrle, H. Jarry,
Toxicology 238 (2007) 192.
[18] M. Heneweer, M. Muusse, M. Van den Berg, J. T. Sanderson, Toxicol. Appl. Pharmacol. 208
(2005) 170.
[19] M. Schlumpf, B. Cotton, M. Conscience, V. Haller, B. Steinmann, W. Lichtensteiger,
Environ. Health Perspect. 109 (2001) 239.
[20] K. Maerkel, W. Lichtensteiger, S. Durrer, M. Conscience, M. Schlumpf, Environ. Toxicol.
Pharmacol. 19 (2005) 761.
[21] D. Seidlová-Wuttke, J. Christoffel, G. Rimoldi, H. Jarry, W. Wuttke, Toxicol. Appl.
Pharmacol. 214 (2006) 1.
[22] S. Harrad, C. Ibarra, M. Diamond, L. Melymuk, M. Robson, J. Douwes, L. Roosens, A.C.
Dirtu, A. Covaci, Environ. Intern. 34 (2008) 232.
[23] H.A. Jones-Otazo, J.P. Clarke, M.L. Diamond, J.A. Archbold, G. Ferguson, T. Harner, M.G.
Richardson, J.J. Ryan, B. Wilford, Environ. Sci. Technol. 39 (2005) 5121.
[24] K. Bröms, K. Svärdsudd, C. Sundelin, D. Norbäck, Indoor Air 16 (2006) 227-235.
Metodología desarrollada-Muestras sólidas
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[25] P. Canosa, I. Rodríguez, E. Rubí, R. Cela, Anal. Chem. 79 (2007) 1675.
[26] S. Morales, P. Canosa, I. Rodríguez, E. Rubí, R. Cela, J. Chromatogr. A 1082 (2005) 128.
[27] G. Ingerowski, A. Friedle, J. Thumulla, Indoor Air 11 (2001) 145.
[28] H.M. Stapleton, T. Harner, M. Shoeib, J.M. Keller, M.M. Schantz, S.D. Leigh, S.A. Wise,
Anal. Bional. Chem. 384 (2006) 791.
[29] M. García-López, P. Canosa, I. Rodríguez, Anal. Bioanal. Chem. 391 (2008) 963.
[30] E.M.Kristenson, L. Ramos, U.A.Th. Brinkman, Trends Anal. Chem. 25 (2006) 96.
[31] S.A. Barker, J. Chromatogr. A 885 (2000)115.
[32] M. García, I. Rodríguez, R. Cela, Anal. Chim. Acta 590 (2007) 17.
[33] M.T. Pena, M.C. Casais, M.C. Mejuto, R. Cela, Anal. Chim. Acta 626 (2008) 155.
[34] C. Sánchez-Brunete, E. Miguel, J.L. Tadeo, Talanta 74 (2008) 1211.
Metodología desarrollada-Fotoiniciadores en alimentos
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B. Fotoiniciadores
1. FOTOINICIADORES EN ALIMENTOS
1.1. Introducción
El uso de fotoiniciadores en la cara externa de los envases de alimentos
puede provocar la contaminación de éstos. De hecho, diversos fotoiniciadores
han sido detectados en leche infantil, yogur, té y zumos. Sus efectos en la salud
humana son desconocidos; sin embargo, son compuestos indeseables cuya
presencia en alimentos debe ser controlada.
Los métodos de preparación de muestra desarrollados hasta el momento
requieren elevados tiempos de análisis, consumo de grandes volúmenes de
disolvente y exhaustivos procesos de concentración, con los posibles riesgos de
pérdida de analitos. Por ello, es necesario el desarrollo de nuevas metodologías
que minimicen la manipulación de la muestra, el empleo de disolventes
orgánicos, el coste y el tiempo de análisis. En esta Tesis, se estudió la
posibilidad de emplear la SPME como técnica de extracción y concentración, en
una única etapa, para la determinación de varios fotoiniciadores en muestras de
leche distribuidas en envases del tipo brick [Negreira, 2010-B]. La elección de la
matriz se ha realizado en base a las alertas sanitarias y a los datos relativos a la
presencia de varios compuestos orgánicos, usados como fotoiniciadores en la
polimerización de tintas, en muestras de leche. También se ha tenido en cuenta
la importancia y el elevado consumo de este alimento, en especial, por parte de
la población infantil. Además de la optimización de las condiciones de
extracción, se ha evaluado la influencia del contenido graso de las muestras en
la eficacia del proceso de concentración.
Metodología desarrollada-Fotoiniciadores en alimentos
308
1.2. Esquema del método desarrollado para fotoiniciadores en
alimentos
Muestra leche 1,5 mL
SPME:Inmersión, fibra PDMS-DVB, 100ºC, 40 min con agitación
Retirada de la fibra, secar con papel
Adicción de agua milli-Q (8,5 mL)
+
Desorción de la fibra 2 min a 270 ºC
GC-MS
Figura 23: Esquema de trabajo seguido para la determinación de fotoiniciadores en leche
mediante SPME y GC-MS.
1.3. Publicación:
SOLID-PHASE MICROEXTRACTION FOLLOWED BY
GAS-CHROMATOGRAPHY MASS SPECTROMETRY
FOR THE DETERMINATION OF
INK PHOTO-INITIATORS IN PACKED MILK
N. Negreira, I. Rodríguez, E. Rubí, R. Cela
Talanta 82 (2010) 296
(doi:10.1016/j.talanta.2010.04.037)
Talanta 82 (2010) 296
311
Solid-phase microextraction followed by gas chromatography mass
spectrometry for the determination of ink photo-initiators in packed milk
N. Negreira, I. Rodríguez*, E. Rubí, R. Cela
Departamento de Química Analítica, Nutrición y Bromatología, Instituto de
Investigación y Análisis Alimentario, Universidad de Santiago de Compostela,
Santiago de Compostela 15782, Spain.
Abstract
A novel, single step method for the determination of seven ink photo-
initiators in carton packed milk samples is described. Solid-phase
microextraction (SPME) and gas chromatography (GC), combined with mass
spectrometry (MS), were used as sample preparation and determination
techniques, respectively. Parameters affecting the performance of the
microextraction process were thoroughly evaluated using uni- and multivariate
optimization strategies, based on the use of experimental factorial designs. The
coating of the SPME fibre, together with the sampling mode and the
temperature were the factors playing a major influence on the efficiency of the
extraction. Under final conditions, 1.5 mL of milk and 8.5 mL of ultrapure water
were poured in a glass vessel, which was closed and immersed in a water
boiling bath. A poly(dimethylsiloxane)-divinylbenzene (PDMS-DVB) coated
fibre was exposed directly to the diluted sample for 40 min. After that, the fibre
was desorbed in the injector of the GC-MS system for 3 min. The optimized
method provided limits of quantification (LOQs) between 0.2 and 1 g L-1 and a
good linearity in the range between 1 and 250 g L-1. The inter-day precision
remained below 15% for all compounds in spiked whole milk. The efficiency of
the extraction changed for whole, semi-skimmed and skimmed milk; however,
no differences were noticed among the relative recoveries achieved for milk
samples, from different brands, with the same fat content.
Metodología desarrollada-Fotoiniciadores en alimentos
312
Keywords: ink photo-initiators, packed food, milk, solid-phase
microextraction, gas chromatography, mass spectrometry.
1. Introduction
Photo-initiators are low molecular weight compounds added to some
printing inks. Under ultraviolet (UV) irradiation these substances are
decomposed generating reactive free radicals, which activate the
polymerization of ink components allowing a fast drying of the printing film.
Generally, photo-initiators present one or two aromatic rings in their molecules.
Some of them, case of benzophenone and amino benzoate derivatives, show
similar chemical structures to those compounds used as UV-filters in
sunscreens and personal care products [1].
The use of ink photo-initiators in the external face of multilayered
packaging cartons can lead to their occurrence in food. In 2005, the European
Food Safety Authority (EFSA) reported the presence of 2-
isopropylthioxanthone (ITX) in several liquid foods, particularly packed milk,
and solid infant formula [2]. Further studies confirmed the presence of ITX not
only in milk [3-5], but also in yoghurts [6], fruit juices [7] and even wine [8], at
concentrations up to several hundreds of g L-1. In addition, other photo-
initiators such as 2-ethylhexyl-4-dimethylaminobenzoate (EHPABA) [4,8] and
benzophenone [8] have been also found in carton packed foods and beverages.
Migration through multilayer materials and/or contamination of the inner face
during storage of rolled bobbins of printing packages may lead to the presence
of ink photo-initiators in food [4,9,10]. Potential long-term effects of photo-
initiators exposure on human health remain unknown; however, they are
considered as undesirable compounds, whose presence in packed foodstuff has
to be controlled. Milk is a particularly concerning matrix, since it is considered a
basic, worldwide consumed nourishment.
Talanta 82 (2010) 296
313
Gas chromatography (GC) and liquid chromatography (LC) based
techniques are often applied to the determination of ink photo-initiators in
packed food. In most cases, they are combined with single or tandem mass
spectrometry (MS) detection [4,8,11]. Moreover, the complexity of foodstuff
matrices makes necessary a previous step to isolate target compounds from the
rest of constituents. Liquid-liquid extraction (LLE), using acetonitrile as
extractant, is one of the most popular approaches for the extraction of photo-
initiators (particularly ITX) from liquid and powdered milk, as well as from
other fatty samples [4-6, 9]. Acetonitrile is compatible with the use of LC, in the
reversed-phase mode, as separation technique and provides extracts with a low
level of lipids. N-hexane extraction followed by a further clean-up using a silica
cartridge has also been reported [8]. Solid-phase extraction (SPE) of diluted
milk samples [3], or of the primary LLE extract from the same matrix [7,12],
allows a reduction in the consumption or organic solvents and a further
improvement in the selectivity of the sample preparation process, respectively.
Solid-phase microextraction (SPME) is a valuable, solvent-free alternative
for the extraction and concentration of organic compounds from different
matrices. SPME integrates extraction and concentration in the same step;
therefore, it competes with multistep strategies in terms of cost and, many
times, SPME provides a higher selectivity since it is based on equilibrium
processes rather than in exhaustive extractions. Liquid foodstuffs, such as milk,
constitute complex matrices, limiting the yield of SPME extractions in
comparison with water samples. In spite of this, several authors have
demonstrated the suitability of SPME for the determination of g per L levels of
different organic compounds in milk [13-16]. When sample pre-treatment
and/or headspace sampling are not possible, dilution of the matrix with water
is a straightforward solution to preserve the integrity of the SPME coating and
to limit the co-extraction of interfering compounds which might damage the GC
column [15-16].
Metodología desarrollada-Fotoiniciadores en alimentos
314
The aim of this study is to develop a single step sample preparation
method, based on the SPME technique, for the determination of a group of
seven ink photo-initiators in packed milk samples. To the best of our
knowledge this is the first application of SPME to the determination of this
family of compounds in milk. Parameters affecting the performance of the
extraction were systematically evaluated using univariant and also
experimental factorial designs studies. After extraction, fibres were thermally
desorbed and compounds determined by GC-MS, in the selected ion
monitoring (SIM) mode.
2. Experimental
2.1. Solvents, standards and SPME equipment
Methanol, acetonitrile (HPLC-grade) and ethyl acetate (trace analysis
grade) were obtained from Merck (Darmstadt, Germany). Sodium chloride was
provided by Aldrich (Milwaukee, WI, USA). Ultrapure water was obtained
from a Milli-Q system (Millipore, Billerica, MA, USA). Standards of
benzophenone (BP), 1-hydroxycyclohexyl-phenylketone (CPK), ethyl-4-
dimethylaminobenzoate (EDMAB), 4-methylbenzophenone (4-MBP), 2,2-
dimethoxy-2-phenylacetophenone (2,2-DMPA), EHPABA and ITX were
acquired from Aldrich. Their chemical structures and some properties of
relevance to predict their behaviour during extraction are summarized in Table
1. In general, target compounds show medium to low polarities and those with
ionisable moieties (CPK, EDMAB and EHPABA) remain in the neutral form at
the pH of milk (6.6-6.8 units). Individual solutions of each species were
prepared in methanol, further dilutions and mixtures of them were also made
in methanol, when used to fortify milk samples, and in ethyl acetate when
considered to optimize GC-MS determination conditions.
Talanta 82 (2010) 296
315
Table 1. Abbreviated names, CAS numbers, structures, log Kow and vapour pressure
(Pv) values of target species.
AAbbbbrreevviiaattiioonn NNaammee CCAASS
nnuummbbeerr SSttrruuccttuurree
aaLLoogg
KKooww
aaPPvv
((mmTToorrrr))
BP Benzophenone 119-61-
9
3.18 0.82
CPK 1-Hydroxycyclohexyl-
phenylketone
947-19-
3
2.34 0.037
EDMAB Ethyl-4-
dimethylaminobenzoate
10287-
53-3
3.14 1.43
4-MBP 4-Methylbenzophenone 134-84-
9
3.64 0.19
2,2-DMPA 2,2-dimethoxy-2-
phenylacetophenone
24650-
42-8
4.76 0.011
EHPABA
2-Ethylhexyl-4-
dimethylaminobenzoate
21245-
02-3
6.15 0.0046
ITX 2-
Isopropylthioxanthone
5495-
84-1
5.33 0.0014
a Values obtained from SciFinder Scholar Database, http://www.cas.org/products/sfacad/
A manual SPME holder and fibres coated with different polymers:
poly(dimethylsiloxane) (PDMS, 100 m film thickness), polyacrylate (PA, 85
m film thickness), Carboxen-PDMS (CAR-PDMS, 75 m film thickness),
PDMS-divinylbenzene (PDMS-DVB, 65 m film thickness) and DVB-CAR-
PDMS (50/30 m film thickness) were obtained from Supelco (Bellefonte, PA,
USA). Before being used for first time, SPME fibres were thermally conditioned
following conditions recommended by the supplier.
Metodología desarrollada-Fotoiniciadores en alimentos
316
2.2. Samples and SPME procedure
The whole (3.6% fat), half-skimmed (1.55% fat) and skimmed (0.30% fat)
milk samples were bought in local supermarkets. All samples were
commercialized in multilayered Tetra Pak or Combibloc type carton packages.
Optimization of SPME conditions was carried out with spiked aliquots (100 g
L-1) of whole milk. The percentage of methanol in this matrix was maintained at
1%. In further experiments, samples with different fat contents were spiked at
increased levels in the range between 1 and 250 g L-1. Spiked samples were
thoroughly homogenized and stored overnight at 4 ºC to simulate the
interactions between analytes and matrix occurring in polluted samples.
SPME experiments were carried out in 10 and 22 mL glass vials
furnished with a PTFE-faced septum and an aluminium crimp cap. A given
volume of milk (from 1.5 to 3 mL) and the corresponding amount of ultrapure
water were poured in the vessels, which contained a magnetic stir bar (10 mm x
4 mm). After being closed, vessels were stirred for 5 min and then stabilized at
the selected temperature for the same period. Then, a SPME fibre was exposed
to the headspace (HS) of the vial or dipped directly into the liquid matrix for a
pre-established period. In some experiments, sodium chloride (NaCl) was also
added to the SPME vessel in order to assess the effect of the ionic strength on
the yield of the extraction. Under optimized conditions, extractions were carried
out in 10 mL (nominal volume) vessels containing 1.5 mL of milk and 8.5 mL of
ultrapure water, without addition of NaCl. A PDMS-DVB fibre was exposed
directly to the stirred sample (700 rpm), previously thermostated at 100 ºC, for
40 min. After this time, the fibre was retracted into the SPME holder. Drops of
sample attached to the outlet surface of the metallic needle were removed with
a soft paper tissue and the fibre was desorbed for 3 min at 270 ºC in the injector
of the GC-MS system.
Talanta 82 (2010) 296
317
2.3. Carton packages
Additionally to the optimization of SPME conditions for milk samples,
ink photo-initiators were also investigated in a limited number of carton
packages. Samples were extracted using the method previously developed and
validated by Sanches-Silva and co-workers [17]. In brief, packages were opened,
the internal side was rinsed with ultrapure water and then, they were cut in
small pieces (around 2 mm x 2 mm). One gram of the above matrix was
accurately weighed and extracted with 10 mL of acetonitrile, at 70 ºC for 24 h.
After filtration, the supernatant solution was concentrated to 2 mL, using a
gentle stream of nitrogen, and injected directly in the GC-MS system.
2.4. Determination
Analytes were determined using a GC-MS system consisting of an
Agilent (Wilmington, DE, USA) 7890A gas chromatograph connected to a
quadrupole type mass spectrometer (Agilent MS 5975C), furnished with an
electron-impact (EI) ionization source. The mass analyzer was operated in the
selected ion monitoring (SIM) mode. Separations were carried out in a HP-5ms
type capillary column (30 m x 0.25 mm i.d., df: 0.25 m) supplied by Agilent.
Helium (99.999 %) was used as carrier gas at a constant flow of 1.0 mL min-1.
The GC oven was programmed as follows: 70 ºC (held for 3 min), at 10 ºC min-1
to 280 ºC (held for 10 min). Ionisation source, mass analyzer and transfer line
temperatures were set at 230, 150 and 290 ºC, respectively. Standards prepared
in ethyl acetate were injected in the splitless mode (splitless time 3 min), with
the injector port at 280 ºC. SPME fibres were desorbed at 270 ºC, case of PDMS-
DVB, PDMS and DVB-CAR-PDMS, or 290 ºC for PA and CAR-PDMS. A
desorption step of 3 min, maintaining the injector in the splitless mode during
this time, was used in all cases. Retention times and m/z ratios of ions used to
monitor the signal of each compound are summarized in Table 2.
Metodología desarrollada-Fotoiniciadores en alimentos
318
Table 2. Performance of the GC-MS quadrupole system for direct injection (2 L) of
photo-initiators standards prepared in ethyl acetate.
Compound Ret.time
(min)
aSelected
ions (m/z)
Linearity,
R2
(10-2000
g L-1)
bRepeatability
(RSDs, %),
50 g L-1
cReproducibility
(RSD, %),
50 g L-1
LOQs
(g L-1)
BP
CPK
EDMAB
4-MBP
2,2-DMPA
EHPABA
ITX
14.98
15.67
15.86
16.41
17.71
21.24
22.29
105,77,182
99,81
148,193,164
119,196
151,105
165,277
239,254
0.9995
0.9995
0.9993
0.9993
0.9999
0.9999
0.9997
1.5
0.6
2.6
5.3
2.0
3.2
2.8
3.4
5.4
2.8
3.4
1.6
4.8
3.6
1.6
2.3
0.7
1.9
0.7
1.1
3.5
a Underlined ions were used for quantification purposes. b data for n=3 consecutive replicates. c data for n=9 replicate injections in 3 different days.
An ion-trap type Varian (Walnut Creek, CA, USA) 240 mass
spectrometer (MS), furnished with an EI source and connected to 450 model GC
instrument, from the same supplier, was also used to confirm the presence of
photo-initiators in some milk cartons. The system was also equipped with a
Factor Four (Varian) BP-5 type capillary column (30 m x 0.25 mm i.d., df: 0.25
m). Injector, transfer line temperatures and rest of chromatographic conditions
were the same as those reported in the above paragraph for the quadrupole GC-
MS instrument; however, the splitless time was reduced to 1 min. Source and
trap temperatures were set at 200 and 150 ºC, respectively. MS spectra were
recorded in the range from 50 to 400 m/z.
3. Results and discussion
3.1. GC-MS determination conditions
Table 2 summarizes some relevant features of the GC-MS (quadrupole)
system for the determination of ink photo-initiators. Under conditions reported
in the experimental section, all compounds were baseline separated showing
retention times comprised between 15 and 23 min. The plots of peak area,
Talanta 82 (2010) 296
319
corresponding to quantification ions (see Table 2), versus concentration fitted a
linear model with determination coefficients (R2) higher than 0.999, within the
interval between 10 and 2000 g L-1. Limits of quantification (LOQs), defined as
the concentration of each species producing a signal 10 times higher than the
baseline noise, remained between 0.7 and 3.5 ng mL-1. Relative standard
deviations (RSDs) corresponding to injections performed within the same day
and in consecutives days stayed below 6%.
3.2. Optimization of SPME parameters
3.2.1. Preliminary experiments
In the initial steps of this study, microextraction experiments were
carried out in 10 mL volume vessels, which contained 2 mL of spiked (100 g L-
1) whole milk plus 8 mL of ultrapure water. Vessels were equilibrated at 100 ºC
and fibres were exposed directed to the diluted samples for 25 min. After that,
they were desorbed using conditions provided in the experimental section. Fig.
1 depicts the responses (peak areas) obtained for triplicate assays using three
SPME coatings. With the only exception of EHPABA, the most hydrophobic of
the considered species, the PDMS fibre provided much lower responses than
PDMS-DVB and PA ones. The latter two coatings showed similar extraction
efficiencies for EHPABA and ITX, whereas PDMS-DVB was preferred for the
rest of photo-initiators. CAR-PDMS and DVB-CAR-PDMS fibres showed a very
low affinity for ITX, as well as a poor repeatability for the rest of species (data
not shown). Thus, PDMS-DVB and PA were selected for additional
experiments. Carry-over effects were evaluated by desorbing each fibre twice at
270 ºC (PDMS-DVB) and 290 ºC (PA). Relative responses in the second
desorption remained below 0.2% of those observed in the first one. In order to
eliminate any risk of cross contamination, they were additionally desorbed (3
min), at the above temperatures, in the split injector of a non operative GC
instrument under a nitrogen stream of 30 mL min-1.
Metodología desarrollada-Fotoiniciadores en alimentos
320
0
1
2
3
BP CPK EDMAB 4-MBP 2,2-DMPA EHPABA ITX
Pea
k ar
ea
PDMS-DVB PDMS PA
(x 105)
Fig. 1. Responses obtained with different fibres for spiked whole milk samples, n=3
replicates. Direct sampling at 100 ºC for 25 min.
3.2.2. Multifactor optimization of SPME conditions
Efficiency of SPME methods is affected by a considerable number of
factors, which are sometimes correlated. A strategy based on the use of
experimental factorial designs was adopted to identify those parameters
playing a major effect on the performance of the SPME process, and to achieve
optimal conditions with a minimum effort and cost.
Initially, a two levels 25-1 type fractional factorial design was used to
assess the effects of temperature, sampling mode, ionic strength, dilution factor
and fibre coating on the efficiency of the extraction. Low and high values for
each of these parameters are given Table 3. Previous assays showed poor
efficiencies operating at room temperature; therefore, the domain of this factor
was established between 55 and 100 ºC. The dilution ratio was chosen according
to the information reported in a previously work, dealing with the application
of SPME to the determination of pesticides in cow milk [16]. The volume of
milk used in each experiment was 1.5 or 3 mL, depending on the dilution ratio,
and the total volume in the SPME vessel was made up to 15 mL with ultrapure
water in all cases. Extractions were performed in 22 mL vials to allow working
Talanta 82 (2010) 296
321
in direct and HS modes, depending on the conditions defined by the
experimental design. Peaks areas obtained for each compound in the 16
extractions involved in the above design were used as variable responses.
Standardized values for main effects corresponding to each factor were
calculated with the Statgraphics Centurion XV software (Manugistics,
Rockville, MD, USA). Their graphical representation is summarized in the main
effect plots grouped in Fig. 2. The length of depicted lines is proportional to the
variation in the response of the investigated species when a given factor
changes from the low (-) to the high (+) level, in the domain of the design. A
positive slope indicates an improvement in the efficiency of the extraction and a
negative one the opposite effect.
Table 3. Experimental domain of the 25-1 fractional design.
Factor Level
Low (-) High (+)
Fibre
Temperature (ºC)
Sampling mode
NaCl (%)
Dilution factor
PDMS-DVB
55
Direct
0
1:5
PA
100
HS
15
1:10
Metodología desarrollada-Fotoiniciadores en alimentos
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Temperature Salt
EDMAB
31
41
51
61
71
81(x 103)
Fibre Mode Dilution
-
+ -
+-
+
- + -+
Pea
k ar
ea
Temperature Salt
BP
93
113
133
153
173
(x 103)
Fibre Mode Dilution
-
+ -
+
- + - + -+P
eak
area
Temperature Salt
2,2-DMPA
43
63
83
103
123(x 103)
Fibre Mode Dilution
-
+ -
+-
+
- + -+P
eak
area
Temperature Salt
EHPABA
02
4
68
1012
(x 103)
Fibre Mode Dilution
- +
-
+-
+
-
+
-+
Pea
k ar
ea
Temperature Salt
ITX
0
3
6
9
12
15(x 103)
Fibre Mode Dilution
- +
-
+ -
+
-
+
-
+Pea
k ar
ea
Temperature Salt
4-MBP
46
66
86
106
126
(x 103)
Fibre Mode Dilution
-
+ -
+
-+
-
+
-
+Pea
k ar
ea
Temperature Salt
CPK
28
38
48
58
68
(x 103)
Fibre Mode Dilution
-
+-
+-
+-
+-
+Pea
k ar
ea
Fig. 2. Main effect plots corresponding to the 25-1 fractional design.
Again, similar responses were attained for EHPABA and ITX with both
fibres, whereas for the rest of species the PDMS-DVB coating provided higher
yields. Direct exposure of the fibre to the diluted samples was preferred to HS
Talanta 82 (2010) 296
323
sampling and higher extraction efficiencies were attained at 100 ºC versus 55 ºC.
The effect of the ionic strength was compound dependent and the dilution
factor affected negatively to the obtained responses. Considering these features,
direct sampling using a PDMS-DVB fibre was fixed for further experiments.
Extractions were carried out in 10 mL vessels containing 1.5 mL of milk plus 8.5
mL of water. This dilution factor (ca. 6.6 times) constitutes a reasonable
compromise between extraction efficiency and stability of the SPME coating,
avoiding the existence of free HS in the vessel.
Optimal values corresponding to temperature, sodium chloride and
extraction time were evaluated with more detail using a 31 x 22 type factorial
design, with three replicates of the central point. The domain of this second
design, the standardized values of main effects associated to each factor and
some relevant two-factor interactions are compiled in Table 4. The absolute
values of the main effects are proportional to the variation in the efficiency of
the extraction when the considered factor changes from the low to the high
level. A positive sign points out to an increase in the yield of the process and a
negative one indicates the opposite behaviour. The temperature was the most
important of the considered factors, with a positive and statistically significant
effect (95% confidence level) in the efficiency of the extraction for all
compounds. In the case of ITX and EHPABA, the quadratic term associated to
this factor (AA) was also significant, Table 4. The corresponding main effect
plot suggested an exponential increase in the efficiency of the SPME with the
temperature of the sample for both species, figure not shown. The sampling
time also affected positively to the extraction process, although it overpass the
level of statistically significance only for EHPABA and ITX. Finally, sodium
chloride exerted a negative effect on the extraction and it was statistically
significant for 4-MBP, EHPABA and ITX. The exception to the above pattern
corresponded to CPK, which is the most polar of the considered compounds.
Consequently, the thermodynamic of CPK extraction increases significantly
with the ionic strength of the solution, whereas the kinetics of its migration
Metodología desarrollada-Fotoiniciadores en alimentos
324
from the bulk of the solution to the surface of the SPME coating is reduced in a
minor extension than for the more hydrophobic compounds [18]. The
interaction temperature-salt (AB) was also statistically significant for CPK,
EHPABA and ITX. The corresponding interaction plots (figure not given)
demonstrated that, for EHPABA and ITX, the negative effect of salt was more
relevant at 100 than at 60 ºC. Taking these comments into account, further
extractions were performed at 100 ºC, without addition of sodium chloride to
the SPME vessel.
Talanta 82 (2010) 296
325
Tab
le 4
. Exp
erim
enta
l dom
ain,
sta
ndar
dize
d m
ain
effe
cts
and
rele
vant
inte
ract
ions
of f
acto
rs in
volv
ed in
the
31
x 22
exp
erim
enta
l
desi
gn. Fa
ctor
C
ode
Lev
el
Stan
dar
diz
ed v
alu
e
Low
M
ediu
m
Hig
h B
P
CP
K
ED
MA
B
4-M
BP
2,2-
DM
PA
E
HP
AB
A
ITX
Tem
per
atu
re (º
C)
A
60
80
100
6.3*
4.
6*
6.7*
8.
2*
6.6*
18
* 21
*
NaC
l (%
) B
0
- 10
-2
.2
0.26
-2
.3
-2.6
* -2
.3
-10*
-9
.0*
Tim
e (m
in)
C
20
- 40
1.
5 1.
0 1.
8 1.
5 1.
7 3.
4*
4.5*
A
A
-
- -0
.32
-0.5
5 0.
20
1.24
1.
23
7.86
* 8.
81*
A
B
-
- 1.
52
3.11
* 1.
58
0.69
0.
77
-8.0
2*
-5.9
6*
*Sig
nifi
cant
fact
ors
at th
e 95
% c
onfi
den
ce le
vel.
Metodología desarrollada-Fotoiniciadores en alimentos
326
A detailed study of the extraction kinetics showed that EHPABA and ITX
required more than 2 hours of direct sampling at 100 ºC to achieve equilibrium
conditions, Fig. 3. For the rest of species, the efficiency of the extraction
normally reached a maximum between 30 and 45 min and then it decreased
slightly, Fig. 3. This trend might be the result of competitive adsorption
processes on the surface of the PDMS-DVB coating. An exposure time of 40 min
was adopted.
0
1
2
3
4
5
0 20 40 60 80 100 120
Pe
ak
are
a
Time (min)
BP
CPK
EDMAB
(x 106)
0
1
2
3
4
5
6
7
0 20 40 60 80 100 120
Pe
ak
are
a
Time (min)
4-MBP
2,2-DMPA
(x 106)
0
1
2
3
0 20 40 60 80 100 120
Pe
ak
are
a
Time (min)
EHPABA
ITX
(x 106)
Fig. 3. Kinetics of the SPME for whole milk. Direct sampling at 100 ºC using a PDMS-
DVB fibre.
Talanta 82 (2010) 296
327
3.2.3. Stirring and methanol addition
Agitation is expected to increase the kinetics of the extraction, improving
the transport of the compounds between the liquid sample and the interface
with the SPME coating. On the other hand, PTFE covered stirrers are a potential
source of cross contamination problems. Experiment data (Fig. 4) demonstrated
that stirring improved significantly the yield of the extraction for all
compounds except CPK. As previously commented, this is the most polar of the
species involved in this study; therefore, it is expected to be that showing the
higher diffusion rates towards the interface between the solution and the SPME
fibre and thus the less affected by stirring. This type of dependence between the
efficiency of stirring and the polarity of analytes has been previously reported
in the literature [19]. In order to avoid the risk of cross contamination problems,
stir bars were wrapped with PTFE tape, which was removed after each
extraction.
0%
20%
40%
60%
80%
100%
120%
BP CPK EDMAB 4-MBP 2,2-DMPA EHPABA ITX
Rel
ativ
e re
spo
nses
With stirring Without stirring
Fig. 4. Effect of stirring (700 rpm) on the relative efficiency of the SPME. Sampling
time 40 min, n= 3 replicates.
Addition of an organic solvent to the SPME vessel contributes to reduce
competitive adsorptions of hydrophobic species on the walls of glass vessels,
improving the efficiency of microextraction for these compounds. On the other
hand, the yield of the process decreases for hydrophilic species, which turn out
Metodología desarrollada-Fotoiniciadores en alimentos
328
more soluble in the sample. During optimization of SPME conditions, fortified
aliquots of milk containing a 1% of methanol were employed. After dilution (1.5
mL to 10 mL), the percentage of methanol in the sample was 0.15%. A series of
extractions was performed using samples containing also 1% and 3% of
methanol. Except in the case of EHPABA and ITX, the efficiency of the
extraction was negatively affected by the addition of methanol, figure not
shown. To limit the contribution of this parameter to the variability of the
extraction, the percentage of methanol in the sampling vessel was always
maintained below 0.5%.
3.2.4. Fat content
In general, the performance of microextraction techniques is affected by
the characteristics of the sample. As a general rule, the higher is the complexity
of the matrix, the lower the yield of the extraction. The complexity of milk
samples is mainly related with their lipidic content. Fig. 5 depicts the responses
obtained for samples of whole, half-skimmed and skimmed cow milk spiked
with target photo-initiators at 50 g L-1. Their declared fat percentages were
3.6%, 1.55% and 0.3%, respectively. The yield of the extraction was inversely
proportional to the fat content in the sample.
0
1
2
3
BP CPK EDMAB 4-MBP 2,2-DMPA EHPABA ITX
Pea
k ar
ea
full-fat half-skimmed skimmed(x 105)
Fig. 5. Comparison of peak areas for spiked (50 g L-1) milk samples with different fat
contents, n= 3 replicates.
Talanta 82 (2010) 296
329
3.3. Analytical figures of merit
The developed method was characterized in terms of linearity,
repeatability and reproducibility using aliquots of whole milk fortified at
different concentrations in the range between 1 and 250 g L-1. Plots of peak
areas versus added concentrations fitted a linear model with determination
coefficients higher than 0.993, Table 5. RSDs of extractions carried out in the
same day (repeatability study) for samples spiked at three different
concentrations (between 5 and 50 g L-1), remained below 7%, only slightly
higher than the repeatability of the GC-MS system for direct injection of
standards in ethyl acetate, see Table 2. RSDs for extractions performed during 3
consecutive days varied between 8 and 15%, Table 5. LOQs of the proposed
method, defined as the concentration of each specie providing a
chromatographic peak with a signal to noise ratio (S/N) of 10, were estimated
from the lowest addition level in the linearity study. Obtained values varied
between 0.2 g L-1 for BP and 1 g L-1 for EDMAB, Table 5. In the case of ITX,
the photo-initiator more often investigated in packed milk, a LOQ of 0.4 g L-1
was achieved. Globally, these values are similar to those attained using LLE,
with additional SPE purification, followed by GC-MS [8] or LC-MS/MS [12];
they are also equivalent to LOQs reported for ITX using SPE with GC-MS/MS
determination [3], and LLE followed by LC-MS/MS [6]. Gallart-Ayala and co-
workers [11] achieved a 100-fold reduction in the LOQ of ITX at the expense of
using a last generation LC-MS/MS instrument providing accurate mass
measurements [11]. Taking into account the trend depicted in Fig. 5, LOQs
reported in Table 5 can be further reduced for semi-skimmed and skimmed
milk. The only exception was BP. For this compound the lower attainable LOQ
was limited by the presence of this specie in procedural blanks corresponding
to the extraction of ultrapure water.
Metodología desarrollada-Fotoiniciadores en alimentos
330
Table 5. Linearity, repeatability (n= 3 replicates), reproducibility (n= 9 replicates in 3
different days) and LOQs of the proposed method for full-fat milk samples.
Compound Linearity (R2)
1-250 g L-1 a
Repeatability (RSDs, %) Reproducibility
(RSDs, %),
16 g L-1 a
LOQs
g L-1 5 g L-1 a 25 g L-1 a 50 g L-1 a
BP
CPK
EDMAB
4-MBP
2,2-DMPA
EHPABA
ITX
0.9962
0.9966
0.9964
0.9969
0.9977
0.9931
0.9969
7.0
4.7
0.6
5.0
6.5
1.0
5.4
5.5
5.3
4.8
3.5
1.9
4.1
4.7
5.1
1.3
4.4
3.3
4.2
3.4
2.4
8.2
8.3
9.5
11.9
15.1
9.6
9.4
0.2
0.6
1
0.4
0.3
0.8
0.4
a Added concentration
Accuracy is a major issue during the validation of any analytical
procedure. Microextraction methodologies are prone to variations in their
efficiency depending on the characteristics of the matrix as it has been already
proved for whole, semi-skimmed and skimmed milk samples, Fig. 5. A further
series of assays was carried out to investigate whether the yield of the SPME
method varies also among samples, with the same fat content, from different
brands, or not. Aliquots corresponding to specimens of whole, semi-skimmed
and skimmed milk, from three different suppliers, were fortified at
concentrations in the range from 10 to 20 g L-1. Non-spiked aliquots of each
specimen were also processed, Fig. 6. One of the specimens was used as
reference and the responses measured for the other two normalized to the first.
Significant differences were not observed among milk samples with the same
fat content, Table 6. Thus, levels of ink photo-initiators in unknown samples can
be quantified by external calibration, using matrix-matched standards. The only
requirement is that both, unknown samples and calibration standards, present
the same fat content.
Talanta 82 (2010) 296
331
21.0 21.1 21.2 21.3 21.460
100
140
180
220
260
300
340
380
Time (min)
Abu
nd
ance
EHPABAm/z 277
14.6 14.8 15.0 15.20
1
2
3
4
5
Time (min)
Abu
nd
ance
(x 104)
BPm/z 105
(x 103)
22.0 22.2 22.40
1
2
3
4
5
Time (min)
Abu
nd
ance
ITXm/z 254
17.5 17.6 17.7 17.8
1
2
3
4
(x 104)
Time (min)
Abu
nd
ance
0
2,2-DMPAm/z 151
15.5
(x 103)
15.8 15.90
123456789
10111213
Time (min)
Abu
nd
ance
EDMABm/z 148
16.015.7 16.30 16.40 16.50
1
2
3
4
(x 104)
Time (min)
Abu
nd
ance
0
4-MBPm/z 119
16.60
15.6 15.7 15.815.4
2
4
6
8
10
12
14
16
18
20
(x 103)
Abu
nd
ance
Time (min)
CPKm/z 99
Fig. 6. GC-MS chromatograms for half-skimmed milk. Dotted line, un-spiked sample.
Solid line, same matrix fortified at 10 g L-1.
Metodología desarrollada-Fotoiniciadores en alimentos
332
Table 6. Relative recoveries provided by the SPME method for milk samples with
different fat contents, n= 3 replicates.
Compound
Relative recoveries (%) ± SD
a Skimmed a Semi-skimmed b Whole
Brand 1 Brand 2 Brand 1 Brand 2 Brand1 Brand 2
BP
CPK
EDMAB
4-MBP
2,2-DMPA
EHPABA
ITX
101±1
94±1
100±4
100±3
101±3
108±3
101±6
101±12
95±11
94±8
104±2
98±4
100±0.2
106±7
102±4
97±1
106±3
97±2
100±2
106±9
100±8
98±7
94±3
94±7
93±7
98±4
101±6
97±3
103±2
101±1
101±5
101±5
105±3
89±5
98±2
94±9
98±3
99±6
100±13
98±14
92±4
96±1
a Added concentration 10 g L-1.
b Added concentration 20 g L-1.
3.4. Real samples
The presence of ink photo-initiators was investigated in a total of 20
samples corresponding to whole, semi-skimmed and skimmed milk. Around
60% of them were distributed in Tetra Pak cartons and the rest in Combibloc ones.
All species remained under the LOQs of the method, although BP was detected,
at concentrations below the LOQ of this species, in some of them. In order to
establish if these results indicate the phase-out of ink photo-initiators in milk
cartons, or simply the lack of significant migration problems, six packages
(Tetra Pak and Combibloc types) were extracted under conditions reported in the
experimental section. A peak at the retention time of BP was found in the
chromatograms provided by the quadrupole GC-MS instrument for the six
samples, whereas the rest of photo-initiators were not noticed. The identity of
this compound was confirmed with the ion-trap GC-MS system, Fig. 7. A
detailed quantification of BP levels in milk packages was beyond the scope of
this research. A rough estimation points out to concentrations below 1 g g-1 in
the six processed samples.
Talanta 82 (2010) 296
333
12.75 13.00 13.25 13.50 13.75 minutes
0.0
0.5
1.0
1.5
2.0
2.5
3.0
kCounts
m/z 105
50 100 150 200 250 300 350 m/z
0%
25%
50%
75%
100%
51
77105
182
50 100 150 200 250 300 350 m/z
0%
25%
50%
75%
100%
51
77
105
182
A
B
BPNIST spectra
Fig. 7. GC-MS (ion-trap) chromatograms and spectra showing the signal of BP in the
extract from a Tetra Pak package. A, procedural blank. B, carton extract.
4. Conclusions
The suitability of SPME for the extraction of seven ink photo-initiators in
packed milk has been demonstrated for first time. Its major advantages over
previously published methods are integration of extraction and concentration in
the same step and null consumption of organic solvents. Although the yield of
the microextraction is conditioned by the fat content in the sample, no
differences were noticed among milk specimens from different brands, with the
same nominal fat content. When combined with GC-MS detection, the
developed procedure provides adequate figures of merit to screen the potential
contamination of milk with photo-initiators used during the printing of
Metodología desarrollada-Fotoiniciadores en alimentos
334
packaging materials. Although this possibility has not been explored in this
study, it is expected that the proposed method can be also adapted to other
liquid foodstuffs, such as fruit juices and wine. Target compounds could not be
quantified in any of the processed milk samples; however, BP was found in
carton packages.
Acknowledgments
This study has been supported by the Spanish Government and E.U.
FEDER funds (project CTQ2009-08377). N.N. thanks a FPU contract to the
Spanish Ministry of Science and Innovation.
Talanta 82 (2010) 296
335
References
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[2] R. Anton, S. Barlow, D. Boskou, L. Castle, R. Crebelli, W. Dekant, K.H. Engel, S. Forsythe, W.
Grunow, M. Heinonen, J.C. Larsen, C. Leclercq, W. Mennes, M.R. Milana, I. Pratt, I. Rietjens,
K. Svensson, P. Tobback, F. Toldrá, EFSA J 293 (2005) 1.
[3] G. Allegrone, I. Tamaro, S. Spinardi, G. Grosa, J. Chromatogr. A 1214 (2008) 128.
[4] A. Gil-Vergara, C. Blasco, Y. Picó, Anal. Bioanal. Chem. 389 (2007) 605.
[5] R. Bagnati, G. Bianchi, E. Marangon, E. Zuccato, R. Fanelli, E. Davoli, Rapid Commun. Mass
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[6] C. Benetti, R. Angeletti, G. Binato, A. Biancardi, G. Biancotto, Anal. Chim. Acta 617 (2008)
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[7] C. Sun, S.H. Chan, D. Lu, H.M.W. Lee, B.C. Bloodworth, J. Chromatogr. A 1143 (2007) 162.
[8] G. Sagratini, G. Caprioli, G. Cristialli, D. Giardiná, M. Ricciutelli, R. Volpini, Y. Zuo, S.
Vittori, J. Chromatogr. A 1194 (2008) 213.
[9] A. Sanches-Silva, S. Pastorelli, J.M. Cruz, C. Simoneau, I. Castanheira, P. Paseiro-Losada, J.
Agric. Food Chem. 56 (2008) 2722.
[10] G. Morlock, W. Schwack, Anal. Bioanal. Chem. 385 (2006) 586.
[11] H. Gallart-Ayala, E. Moyano, M.T. Galceran, J. Chromatogr. A 1208 (2008) 182.
[12] D. Shen, H. Lian, T. Ding, J. Xu, C. Shen, Anal. Bioanal. Chem. 395 (2009) 2359.
[13] M.J. González Rodríguez, F.J. Arrebola Liébanas, A. Garrido Frenich, J.L. Martínez Vidal,
F.J. Sánchez López, Anal. Bioanal. Chem. 382 (2005) 164.
[14] C. Mardones, D. Von Baer, J. Silva, M.J. Retamal, J. Chromatogr. A 1215 (2008) 1.
[15] N. Aguinaga, N. Campillo, P. Viñas, M. Hernández-Córdoba, Anal. Chim. Acta 596 (2007)
285.
[16] M. Fernandez-Alvarez, M. Llompart, J.P. Lamas, M. Lores, C. García-Jares, R. Cela, T.
Dagnac, Anal. Chim. Acta 617 (2008) 37.
[17] A. Sanches-Silva, S. Pastorelli, J.M. Cruz, C. Simoneau, P. Paseiro-Losada, J. Food Sci. 73
(2008) C92.
[18] J. Pawliszyn, Solid Phase Microextraction, Theory and Practice, Wiley-VCH, New York,
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[19] P. Canosa, I. Rodríguez, E. Rubí, M.H. Bollaín, R. Cela, J. Chromatogr. A 1124 (2006) 3.
Conclusiones
339
IV. CONCLUSIONES
Tras la presentación y discusión de los resultados obtenidos es
conveniente remarcar las conclusiones más relevantes y revisar si se han
alcanzado los objetivos inicialmente planteados.
El objetivo fundamental a la hora de optimizar cada una de las
metodologías recogidas en esta Tesis ha sido el desarrollo de métodos sencillos,
rápidos y de bajo coste, que minimicen al máximo la generación de residuos y
proporcionen unos bajos límites de cuantificación, así como prestaciones
adecuadas en lo referente a su precisión y exactitud. Además, se ha trabajado
con matrices complejas, ej. polvo, agua residual, lodos y alimentos, intentando
combinar la etapa de extracción con la de clean-up. Finalmente, la aplicación de
dichas metodologías a muestras reales evidencia sus posibilidades y sus
limitaciones, además de aportar información relativa a la distribución y el
comportamiento de los analitos en las matrices consideradas.
A continuación se exponen las conclusiones más significativas de cada
uno de los estudios realizados:
1) “Dispersive liquid-liquid microextraction followed by gas
chromatography-mass spectrometry for the rapid and sensitive
determination of UV filters in environmental water samples”.
DLLME constituye una alternativa sensible, rápida y sencilla para la
concentración y extracción de 9 filtros solares (EHS, HMS, BzS, BP-3, 4-MBC,
IAMC, EHMC, EHPABA y OCR) de muestras de agua superficial y residual. El
método propuesto no requiere instrumentación dedicada, el volumen de
muestra es pequeño (10 mL), el tiempo de preparación de muestra es de tan
sólo 5 minutos y la fase extractante es compatible con el uso de cromatografía
de gases como técnica de determinación. La eficacia de extracción no se ve
afectada por la fuerza iónica, ni por el pH, ni por la naturaleza de la muestra,
Conclusiones
340
mientras que el volumen y el tipo de extractante y dispersante sí mostraron una
gran influencia. Por otro lado, las principales limitaciones del método
desarrollado son la dificultad para su automatización, así como el empleo de
disolventes halogenados en la etapa de extracción.
2) “Silicone discs as disposable enrichment probes for gas
chromatography-mass spectrometry determination of UV filters in
water samples"
El uso de absorbentes desechables de silicona resultó adecuado para la
extracción de filtros solares en muestras de agua superficial y residual,
proporcionando LOQs en la región de los bajos ng L-1, cuando se combina con
detección mediante GC-MS, empleando inyección de grandes volúmenes. Las
condiciones óptimas de extracción son similares a las publicadas en la
bibliografía para Twisters recubiertos de PDMS; además, las eficacias de
extracción son equivalentes en ambos casos. La principal limitación del método
es la lenta cinética del proceso de extracción; sin embargo, la utilización de una
placa agitadora multiposición permite procesar de manera simultánea, y con
una mínima atención por parte del operador, un número muy importante de
muestras.
La aplicación de las anteriores técnicas de microextracción a muestras de
agua superficial y residual ha puesto de manifiesto que los filtros UV más
frecuentemente detectados han sido OCR, EHMC y 4-MBC, mientras que
EHPABA y IAMC no han sido encontrados en prácticamente ninguna de las
muestras procesadas.
3) “Sensitive determination of salicylate and benzophenone type UV
filters in water samples using solid-phase microextraction,
derivatization and gas chromatography tandem mass spectrometry".
La combinación de la SPME con una etapa posterior de sililación on-fiber
hace posible la determinación de BP-1 y BP-8 mediante GC-MS, además de
Conclusiones
341
mejorar la forma de pico y los límites de cuantificación alcanzados para los
filtros UV: BP-3, EHS y HMS. Las condiciones óptimas de trabajo
correspondieron a muestreo directo, a temperatura ambiente, empleando una
fibra de PDMS-DVB y considerando MSTFA como agente sililante en la etapa
de derivatización. El método propuesto alcanzó LOQs en el nivel de los bajos
ng L-1, claramente inferiores a los descritos previamente en la bibliografía para
SPME, y una repetibilidad aceptable (RSD ≤ 14%). En el caso de muestras de
agua de río y residual tratada, la eficacia de la extracción fue similar a la
obtenida para agua ultrapura. El método, se aplicó a muestras
medioambientales encontrando concentraciones significativas de BP-3 y BP-1 en
aguas residuales.
4) “Solid-phase extraction followed by liquid chromatography
tandem mass spectrometry for the determination of hydroxylated
benzophenone UV absorbers in environmental water samples".
En este trabajo, se optimizaron las condiciones experimentales para la
determinación simultánea de 2 filtros solares (BP-3 y BP-4) y 4 benzofenonas
relacionadas (BP-1, BP-2, BP-6 y BP-8) mediante LC-MS/MS, empleando la SPE
como técnica de concentración. Las diferencias de polaridad y acidez entre la
BP-4 y el resto de benzofenonas han requerido una optimización exhaustiva
tanto de las condiciones de extracción en SPE, usando un adsorbente en fase
reversa (Oasis HLB, 60 mg), como de separación en la columna de LC. Además,
ha sido preciso combinar de modo simultáneo la ionización mediante ESI en
modo positivo y negativo, dado que, a diferencia de lo que ocurre con el resto
de compuestos, la BP-3 genera el ión precursor [M+H]+ con una eficacia muy
superior al [M-H]-.
Los LOQs alcanzados en este trabajo (desde 0,4 ng L-1 para agua de río
hasta 32 ng L-1 para agua residual sin tratar) permiten la determinación de las
seis benzofenonas consideradas en muestras de agua superficial y residual. La
BP-4 se cuantificó en todas las muestras analizadas, incluso en agua residual
Conclusiones
342
tratada, lo que indica que no es eliminada en las estaciones depuradoras. Este
hecho, añadido a su alta polaridad y elevada solubilidad en agua, indica que la
BP-4 podría alcanzar incluso los suministros de agua potable. Por su parte, la
BP-3 y la BP-1 también fueron detectadas en agua residual sin tratar y en río; sin
embargo, los datos obtenidos apuntan a una eliminación considerable de ambos
compuestos en las estaciones depuradoras de aguas residuales urbanas.
En futuros trabajos es necesario mejorar la selectividad de la etapa de
extracción, con objeto de reducir la influencia de la matriz en la eficacia de la
ionización mediante electrospray. Aprovechando el carácter ácido de estos
analitos parece factible que adsorbentes en modo mixto (ej. Oasis MAX)
permitan llevar a cabo un fraccionamiento de las benzofenonas y otros
compuestos con carácter neutro y básico.
5) "Study of some UV filters stability in chlorinated water and
identification of halogenated by-products by gas chromatography-
mass spectrometry".
En primer lugar, se desarrolló un método para la determinación de BP-3,
BP-1, EHPABA y EHS en muestras de agua usando GC-MS como técnica de
determinación, SPE en la etapa de concentración y derivatización de los
compuestos más polares.
BP-3, BP-1 y EHPABA presentan una elevada reactividad en muestras de
agua que contienen bajas concentraciones de cloro libre, similares a las
existentes en agua de grifo o de piscina. Por el contrario, el EHS presenta una
mayor estabilidad. En exceso de cloro libre, tanto BP-1 como BP-3 y EHPABA se
degradan siguiendo cinéticas de orden 1. La vida media de los compuestos es
función de variables tales como la concentración de cloro libre, el pH del agua y
la presencia de trazas de bromuro. Como productos de estas reacciones, se han
identificado compuestos halogenados (clorados y bromados) resultado de
reacciones de sustitución de hidrógenos por halógenos y, en el caso de la BP-3,
también productos halogenados donde no se conserva la estructura química de
Conclusiones
343
la benzofenona. Algunos de los productos de transformación son resistentes a
reacciones posteriores de oxidación, por lo que se podrían encontrar en matrices
medioambientales, ej. agua y sedimentos.
6) "Optimization of pressurized liquid extraction and purification
conditions for gas chromatography-mass spectrometry
determination of UV filters in sludge".
La extracción mediante PLE combinada con la utilización de carbón
grafitizado en la celda de PLE para la retención de pigmentos y una limpieza
adicional con un cartucho de PSA, constituye una mejora en términos de
eficacia de extracción y selectividad para la determinación mediante GC-MS de
un amplio número de filtros solares (EHS, HMS, BP-3, 4-MBC, IAMC, EHMC,
EHPABA y OCR) en muestras de lodo. Este estudio representa la primera
aplicación de ambos materiales para la purificación de extractos de PLE
correspondientes a muestras de lodo. Los factores con una mayor influencia
sobre el rendimiento global del proceso de preparación de muestra fueron los
disolventes empleados en la etapa de extracción y en la elución posterior de los
analitos del cartucho de PSA.
La aplicación del método a muestras de lodo confirmó la acumulación
significativa de tres filtros solares (4-MBC, EHMC y OCR) en esta matriz, con
concentraciones promedio superiores a 600 ng g-1.
7) "Determination of selected UV filters in indoor dust using matrix
solid-phase dispersion and gas chromatography tandem mass
spectrometry".
La combinación de MSPD con GC-MS/MS ha permitido determinar, por
primera vez, los niveles de seis filtros UV (EHS, HMS, 4-MBC, IAMC, EHMC y
OCR) en muestras de polvo. El método desarrollado integra las etapas de
extracción y clean-up en un único paso, con un consumo moderado de
Conclusiones
344
disolventes orgánicos y sin necesidad de utilizar instrumentación costosa en la
etapa de preparación de muestra. Las recuperaciones obtenidas fueron
superiores al 77% para todos los compuestos, y los LOQs se mantuvieron en la
región de los bajos ng g-1. Su aplicación a muestras de polvo, procedentes de
viviendas particulares y edificios públicos, puso de manifiesto la presencia de
filtros UV en esta matriz, a niveles de concentración un orden de magnitud
superiores a los existentes en lodos, alcanzando valores máximos de 15 y 41 μg
g-1 para EHMC y OCR, respectivamente.
8) "Solid-phase microextraction followed by gas chromatography
mass spectrometry for the determination of ink photo-initiators in
packed milk".
Con este estudio se demostró la adecuación de la SPME para la
extracción de 7 fotoiniciadores, con estructuras similares en muchos casos a
filtros UV derivados de la benzofenona y el ácido p-amino benzoico, en leche
envasada y su determinación mediante GC-MS. Sus ventajas respecto a trabajos
publicados anteriormente son el nulo consumo de disolventes orgánicos y la
integración de extracción y concentración en la misma etapa. Aunque, la
eficacia de la microextracción se vió afectada por el contenido graso de la
muestra, no se apreciaron diferencias entre distintas marcas de leche con el
mismo contenido graso nominal. El método proporcionó LOQs en el rango de
valores recogidos en la bibliografía, empleando metodologías de preparación de
muestra más complejas; sin embargo, no se detectaron niveles significativos de
estos fotoiniciadores en ninguna de las muestras de leche analizadas.
Bibliografía
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ABREVIATURAS
Y ACRÓNIMOS
Abreviaturas y acrónimos
359
VI. ABREVIATURAS Y ACRÓNIMOS
A ADN Ácido desoxirribonucleico
APCI Atmospheric pressure chemical
ionization Ionización química a presión atmosférica
API Atmospheric pressure ionization Ionización a presión atmosférica
APPI Atmospheric pressure photo-
ionization Fotoionización a presión atmosférica
ARN Ácido ribonucleico
ASE Accelerated solvent extraction Extracción acelerada con disolventes
B
BDE Brominated diphenyl ether Difenil éter polibromado
BSTFA N,O-bis-
(trimethylsilyl)trifluoroacetamide N,O-bis-(trimetilsilil)trifluoroacetamida
C
CAR-PDMS Carboxen-Polydimethylsiloxane Carboxen- Polidimetilsiloxano
CW-DVB Carbowax-Divinylbenzene Carbowax-Divinilbenceno
C18 Octadecyl modified silica phase Sílica modificada con grupos octadecilos
D
DDT Dichlorodiphenyltrichloro ethane Diclorodifeniltricloro etano
DC Direct current Corriente continua
DLLME Dispersive liquidliquid
microextraction Microextracción líquidolíquido dispersiva
E
EU European Union Unión Europea
EI Electronic impact Impacto electrónico
EPA Environmental Protection Agency Agencia de protección medioambiental
ESI Electrospray Electrospray
ESL Solid-liquid extraction Extracción sólido-líquido
Abreviaturas y acrónimos
360
F
FDA Food and Drug Administration Administración de alimentos y
medicamentos (USA)
G
GC Gas chromatography Cromatografía de gases
H
HCX Hydrophobic Cation Exchange Intercambio catiónico y fase reversa
HF-LPME Hollow fiber-Liquid phase
microextraction
Microextracción en fase líquida con fibra
hueca
HLB Hydrophilic-Lipophilic Balance Balance hidrófilo-lipófilo
HPLC High performance liquid
chromatography Cromatografía líquida de alta resolución
HS Headspace Espacio de cabeza
I
IL Ionic liquid Líquido iónico
IUPAC International Union of Pure and
Applied Chemistry
Unión internacional de Química pura y
aplicada
K
Kow Octanol-Water partition constant Constante de partición octanol-agua
L
LC Liquid chromatography Cromatografía líquida
LD Liquid desorption Desorción líquida
LDPE Low-density polyethylene Polietileno de baja densidad
LLE Liquidliquid extraction Extracción líquidolíquido
LLME Liquidliquid microextraction Microextracción líquidolíquido
LOD Limit of detection Límite de detección
LOQ Limit of quantification Límite de cuantificación
LPME Liquid-phase microextraction Microextracción en fase líquida
M
MAX Mixed-Mode Anion Exchange Modo mixto de intercambio aniónico y fase
reversa
Abreviaturas y acrónimos
361
MCX Mixed-Mode Cation Exchange Modo mixto de intercambio catónico y fase
reversa
MEPS Microextraction by packed Sorbent Microextracción con adsorbente
empaquetado
min Minutes Minutos
MR Reference material Material de referencia
MS Mass spectrometry Espectrometría de masas
MS-MS Tandem mass spectrometry Espectrometría de masas en tándem
MSPD Matrix solid-phase dispersion Dispersion de matriz en fase sólida
MSTFA N‐Methyl‐N‐(trimethylsilyl)trifuoroace
tamide N‐metil‐N‐(trimetilsilil)trifuoroacetamida
MTBE Methyl tert-butyl ether t-butilmetil éter
MTBSTFA N-Methyl-N-[tert-butyldimethyl-
silyl]trifluoroacetamide
N‐(tert‐butildimetilsilil)‐N‐metiltrifluoroacet
amida
m/z Mass/charge Masa/carga
P
PA Polyacrylate Poliacrilato
PAH Polycyclic aromatic hydrocarbon Hidrocarburo policíclico aromático
PCB Polychlorinated biphenyl Bifenilo policlorado
PDMS Polydimethylsiloxane Polidimetilsiloxano
PDMS-DVB Polydimethylsiloxane-Divinylbenzene Polidimetilsiloxano-Divinilbenceno
PE Polyethylene Polietileno
PEG Poliethylenglycole Polietilenglicol
PLE Pressurized liquid extraction Extracción con disolventes presurizados
PSA Primary-secundary amine Amina primaria y secundaria
R
R2 Determination coefficient Coeficiente de determinación
RP-HPLC Reversed phase-high pressure liquid
chromatography Cromatografía líquida en fase reversa
RSD Relative standard deviation Desviación estándar relativa
S
SBSE Stir bar sorptive extraction Extracción con barra agitadora
SDME Single drop microextraction Microextracción con gota
Abreviaturas y acrónimos
362
SEC Size exclusion chromatography Cromatografía de exclusión por tamaños
SIM Selected ion monitoring
S/N Signal-to-noise ratio Relación señal/ruido
SPE Solid-phase extraction Extracción en fase sólida
SPME Solid-phase microextraction Microextracción en fase sólida
T
TBDMS Tert-butyldimethylsilyl Tert-butildimetilsilil
TD Thermal desorption Desorción térmica
U
UPLC Ultra- performance liquid
chromatography
USAEME Ultrasound-assisted
emulsificationmicroextraction
Microextracciónemulsificación asistida por
ultrasonidos
UV Ultraviolet Ultravioleta
V
Vm Volumen de muestra
W
WAX “Weak” Mixed-Mode Anion Exchange Modo – mixto de intercambio aniónico débil
y fase reversa
WCX “Weak” Mixed-Mode Cation Exchange Relleno de modo – mixto de intercambio
cationico débil y fase reversa
VII. ANEXO: OTRAS PUBLICACIONES
Analytica Chimica Acta 575 (2006) 106–113
Formation of halogenated by-products of parabens in chlorinated water
P. Canosa, I. Rodrıguez ∗, E. Rubı, N. Negreira, R. CelaDepartamento de Quımica Analıtica, Nutricion y Bromatologıa, Instituto de Investigacion y Analisis Alimentario,
Universidad de Santiago de Compostela, Santiago de Compostela 15782, Spain
Received 31 March 2006; received in revised form 19 May 2006; accepted 22 May 2006Available online 27 May 2006
Abstract
Chemical transformations of four alkyl esters of p-hydroxybenzoic acid, parabens, in chlorinated water samples are investigated. Quantification ofthe parent species and identification of their reaction by-products were performed using gas chromatography in combination with mass spectrometry.Experiments were accomplished considering free chlorine and paraben concentrations at the mg L−1 and �g L−1 level, respectively. Concentrationof water samples, using solid-phase extraction, and silylation of the target species were carried out in order to improve the detectability of parentspecies and their possible transformation products, achieving quantification limits at the low ng L−1 level. Under employed experimental conditions,the decrease in the concentrations of parabens followed pseudo-first-order kinetics. Half-lives values obtained for model ultrapure water solutionswere in good agreement with those observed in tap water samples. For the first type of sample, only two by-products were detected for eachparaben. They corresponded to chlorination of the aromatic ring in one or two carbons situated in ortho-positions to the hydroxyl group. Bothspecies were also generated after the addition of parabens to chlorinated tap water. Moreover, three new transformation products were noticedfor each parent compound. They were identified as bromo- and bromochloro-parabens, formed due to the existence of traces of bromide in tapwater sources. Experiments carried out by mixing paraben-containing personal care products with tap water, containing free chlorine, confirmedthe formation of all above described halogenated by-products. In addition, the presence of the di-chlorinated forms of methyl and propyl parabenhas been detected for first time in raw sewage water samples.© 2006 Elsevier B.V. All rights reserved.
Keywords: Parabens; Chlorine; Halogenated by-products; Water samples; Gas chromatography–mass spectrometry
1. Introduction
Parabens, esters of p-hydroxybenzoic acid, are extensivelyemployed as bactericides and preservative agents in antiperspi-rants, sunscreen creams, bath gels, shampoos and toothpaste[1]. As in the case of many personal care chemicals, they arecontinuously released in the aquatic media through domesticwastewater, therefore, a growing concern has arisen in relationto their potential long term effects on wildlife. Nowadays, itis known that parabens are weak endocrine disrupters [2] and,although they are removed in a considerable extension duringconventional sewage water treatments [3], their presence hasbeen detected in river water samples at the low ng L−1 level[4]. Moreover, recent studies have suggested a possible relation-ship between breast cancer and prolonged dermal expositions toparaben-containing products [5].
∗ Corresponding author. Tel.: +34 981 563100x14387; fax: +34 981 595012.E-mail address: qnisaac@usc.es (I. Rodrıguez).
The possibility that personal care chemicals and non-prescription drugs might reach potable water sources has fos-tered different studies aiming to test their stability and to studythe formation of undesirable by-products, during water chemicaldisinfection treatments [6]. This information is also relevant toimprove their removal efficiency in sewage water plants usingadvanced oxidation processes. On the other hand, less atten-tion has been paid to the reactivity of personal care productswhen mixed with tap water in our homes. Independently ofthe primary disinfection treatment applied in production plants,tap water is normally amended with free chlorine to insureits bacteriological quality throughout the distribution system.Chlorine is a potent oxidant able to react with natural organicmatter and anthropogenic chemicals rendering different halo-genated by-products. Some of them are toxic species, whichmight suppose a potential risk for human health [7,8]. Particu-larly, compounds containing phenolic groups in their structuresshow favourable chlorination kinetics [9–12]. As an example,recent studies have demonstrated that personal care productscontaining triclosan (5-chloro-2-(2,4-dichlorophenoxy) phenol)
0003-2670/$ – see front matter © 2006 Elsevier B.V. All rights reserved.doi:10.1016/j.aca.2006.05.068
P. Canosa et al. / Analytica Chimica Acta 575 (2006) 106–113 107
produce several toxic and persistent by-products, such as 2,4-dichloro and 2,4,6-trichlorophenol, when they are mixed withchlorinated water [13,14].
Aims of this work are: (i) to evaluate the stability of fouralkylated parabens (methyl, ethyl, propyl and butyl paraben)in water samples containing free chlorine (sodium hypochlo-rite plus hypochlorous acid) at the low mg L−1 level, (ii) todetermine their half-lives under different experimental condi-tions and (iii) to identify their potential by-products as well as toinvestigate their further stability under oxidative conditions. Pre-liminary studies were carried out using buffered ultrapure watersamples spiked with known amounts of chlorine and parabenstandard solutions. The relevance of the obtained results wasfurther assessed by mixing chlorinated tap water with paraben-containing personal care products. Determination of parabenremaining concentrations in the above experiments and iden-tification of their by-products were performed using gas chro-matography with mass spectrometry. In order to achieve quan-tification limits at the low ng L−1 level, for the parent species,and also to identify their oxidation by-products, even if theyare formed in a minor extension, a sample preparation proce-dure was developed. It consisted of water concentration usinga solid-phase extraction (SPE) cartridge, followed by elution ofthe target species and derivatization with a silylation reagent toimprove the performance of gas chromatography separations.
2. Experimental
2.1. Standards and material
HPLC-grade methanol, ethyl acetate for trace analysisand sodium thiosulphate were supplied by Merck (Darm-stadt, Germany). Methyl paraben (MeP), ethyl paraben (EtP),propyl paraben (PrP) and butyl paraben (BuP), as well asthe derivatization reagent N-methyl-N-(tert-butyldimethylsilyl)-trifluoroacetamide (MTBSTFA), were purchased from Aldrich(Milwaukee, WI, USA). Standards of 3-chloromethylparaben(3-ClMeP) and 3-chloroethylparaben (3-ClEtP) were alsoobtained from Aldrich. Individual solutions of each analyte wereprepared in methanol. Further dilutions and mixtures of the fournon-halogenated parabens were made in methanol and ethylacetate. The first were employed for preparing spiked water sam-ples. Optimisation of GC–MS conditions was performed usingthe silylated derivatives of the analytes. They were obtained byaddition of 40 �L of MTBSTFA to aliquots (0.5 mL volume) ofparaben solutions in ethyl acetate [13].
Sodium hyphochlorite with a nominal free chlorine contentof 4% (w/v) was purchased from Aldrich. This solution wasstored at 4 ◦C and its exact concentration determined weekly byiodometric titration [15]. OASIS HLB (60 mg) SPE cartridgeswere acquired from Waters (Milford, MA, USA). Glass woolfilters were purchased from Millipore (Billerica, MA, USA).
2.2. Concentration of water samples
Recoveries of the sample preparation method for non-halogenated parabens were evaluated using spiked and non-
spiked samples of ultrapure, tap and raw wastewater. Wastewatersamples were taken from the main sewer of a 100,000 inhabi-tants city, filtered and stored for a maximum of 1 week at 4 ◦Cbefore analysis. Ultrapure and tap water were obtained directlyin the laboratory when needed. Recoveries of the SPE procedurewere calculated using spiked aliquots (1 L volume) of the abovedescribed water samples adjusted at pH 2.5. After the extractionstep, the SPE sorbent was dried using a nitrogen stream and thenanalytes were eluted with 2 mL of ethyl acetate. A fraction of thisextract (0.5 mL volume) was poured in a 1.5 mL GC autosam-pler vial together with 40 �L of MTBSTFA. The mixture washeated for 1 h at 70 ◦C and injected in the GC–MS system afterreturning to room temperature. Standard solutions of parabensin ethyl acetate were silylated using the same protocol. Quantifi-cation was performed using external calibration by comparisonof peak areas obtained for silylated standards and extracts fromSPE cartridges.
2.3. Chlorination experiments
The potential degradation of parabens and the formationof disinfection by-products was investigated using samples(100–250 mL) of ultrapure (Milli-Q) and tap water contain-ing a known amount of chlorine. Experiments were performedat room temperature, unless otherwise is stated, using light-protected 300 mL glass bottles. Samples were adjusted at dif-ferent pHs, in the range from 6.3 to 8.6, with 0.1 M phosphateor bicarbonate buffer solutions. After that, they were spikedwith a methanolic solution containing one or several parabensand with the corresponding volume of a sodium hypochlo-rite solution, which exact concentration was previously deter-mined by titration. Obviously, the last step was not necessaryfor experiments accomplished using chlorinated tap water. Inthis case the water free chlorine content was determined usingthe N,N-diethyl-p-phenylenediamine procedure with photomet-ric detection [16]. Bottles were closed, shaken manually for1–2 min and allowed to stand during an established reactiontime. Then, the excess of chlorine was removed by addition ofsodium thiosulphate (10 mg per 100 mL of water) and sampleswere adjusted at pH 2.5 and concentrated as described in theprevious paragraph. Initial concentrations of free chlorine andparent compounds in the above experiments ranged from 0 to5 mg L−1 and from 10 to 40 �g L−1, respectively. The percent-age of methanol in the spiked samples was maintained below0.04%.
2.4. GC–MS determination
Parabens and their transformation by-products were deter-mined by GC–MS. The system was a Varian CP 3900 GasChromatograph (Walnut Creek, CA, USA) connected to an ion-trap mass spectrometer (Varian Saturn 2100). Separations werecarried out using a HP-5 MS capillary column (30 m × 0.25 mmi.d., df: 0.25 um) supplied by Agilent (Wilmington, DE, USA).Helium (99.999%) was used as carrier gas at a constant flow of1 mL min−1. The GC oven was programmed as follows: 2 min at50 ◦C, at 10 ◦C min−1 to 270 ◦C (held for 10 min). The GC–MS
108 P. Canosa et al. / Analytica Chimica Acta 575 (2006) 106–113
interface and the ion trap temperatures were set at 270 and220 ◦C, respectively. The temperature of the injector was main-tained at 280 ◦C and injections (1 �L volume) were made in thesplitless mode (purge time 2 min) using an autosampler device.The mass spectrometer was operated in the electron impact (EI)mode (70 eV). Spectra were recorded in the range from 90 to550 m/z units.
3. Results and discussion
3.1. Performance of the analytical procedure
In order to investigate the possible reactions of parabens con-sidering levels of free chlorine and parent compounds whichmimic those present in real life water samples, a concentrationstep was necessary. Samples were extracted using reversed-phase SPE cartridges (OASIS HLB, 60 mg). Extraction con-ditions were optimised to maximize the recoveries for thefour parabens considered as parent species in this study. Ethylacetate was chosen to elute the analytes from the SPE car-tridge. This solvent presents a medium polarity allowing thequantitative recovery of phenolic species from reversed-phasesorbents [13] and, moreover, it does not contain hydroxylgroups, which might consume the silylation reagent. There-fore, the derivatization reaction was accomplished directly inthe cartridge extract, without introducing an additional solventexchange step. Parabens were recovered quantitatively from theOASIS sorbent using only 2 mL of ethyl acetate. After that,0.5 mL of this extract were mixed with 40 �L of MTBSTFAto obtain the silylated derivatives. Table 1 summarizes the m/zratios employed for quantification of parabens, the linearity inthe response of the GC–MS for silylated standards contain-ing increasing concentrations, at seven different levels, from10 to 2000 ng L−1, and the recoveries of the sample prepara-tion procedure for different water samples, 1 L volume, spikedat the 0.5 �g L−1 level. In the case of tap water, previouslyto the addition of parabens, the available free chlorine wasremoved using sodium thiosulphate. Raw sewage water sam-ples contained significant levels of some parabens, particularlyMeP and PrP; therefore, recoveries were calculated by divid-ing the difference between concentrations measured for spikedand non-spiked aliquots of the same water sample by the addedone. Considering a sample intake of 1 L, quantification limitsof the developed procedure remained at the low ng L−1 level(Table 1).
Fig. 1. Influence of pH and free chlorine concentration on the stability of PrP.Reaction time 10 min.
Fig. 2. Plots of MeP concentration, logarithmic values, vs. the reaction time forwater samples buffered at pH 7.3. (A) Ultrapure water spiked with chlorine at0.4 and 1.6 mg L−1. (B) Tap water samples containing 0.46 mg L−1 of chlorineand spiked with sodium hypochlorite to achieve a chlorine concentration of1.60 mg L−1.
Table 1Quantification ions, linearity, recoveries (n = 4 replicates, 1 L volume samples) and quantification limits of the analytical procedure for paraben species
Compound Retentiontime (min)
Quantificationion (m/z)
Regressioncoefficient (R2)
%Recovery ± R.S.D. QL (S/N 10)(ng L−1)
Ultrapure water Tap water Raw wastewater
MeP 15.59 209 0.998 106.2 ± 2.3 80.4 ± 4.6 104.7 ± 4.1 10EtP 16.35 223 0.998 109.2 ± 2.2 94.3 ± 2.9 91.1 ± 5.8 10PrP 17.38 237 0.998 111.9 ± 1.0 97.9 ± 3.5 96.4 ± 4.1 6BuP 18.39 251 0.999 106.9 ± 1.3 97.3 ± 3.8 103.4 ± 2.3 3
P. Canosa et al. / Analytica Chimica Acta 575 (2006) 106–113 109
Table 2Calculated half-lives (t1/2) for parabens at different chlorine concentrations
Sample type Chlorine concentration (mg L−1)a t1/2 (min)
MeP EtP PrP BuP
Ultrapure water 0.40b 33.8 (0.990) 26.1 (0.996) 30.9 (0.990) 27.6 (0.991)Ultrapure water 1.60b 4.6 (0.999) 5.3 (0.997) 4.2 (0.998) 5.0 (0.991)Tap water 0.46c 31.5 (0.990) 31.5 (0.990) 30.3 (0.999) 29.9 (0.997)Tap water 1.60b 4.7 (0.998) 4.7 (0.992) 4.6 (0.994) 4.4 (0.991)
Values in parentheses correspond to the correlation coefficients of graphs plotting the logarithmic value of paraben concentrations vs. the reaction time.a Initial concentration at zero time.b Water samples spiked with sodium hypochlorite.c Tap water without any extra addition of chlorine.
3.2. Behaviour of parabens in presence of chlorine
3.2.1. Effect of sample pH and chlorine concentrationThe relevance of reactions between phenolic species and free
chlorine (hypochlorous acid and hypochlorite) depends on sev-eral factors such as the concentration of chlorine, the pH of themedia and the kinetic of the process. In order to assess whetherparabens react at a significant extent with low chlorine concen-trations, such as those contained in tap and water samples usedin recreational activities, a first set of experiments was designed.Aliquots of ultrapure water (100 mL), buffered at different pHsin the range between 6 and 9, were spiked with a fixed con-centration of parabens (40–45 �g L−1) and increasing levels offree chlorine from 0 to 2.4 mg L−1. Independent series of exper-iments were carried out for each parent compound. After a reac-tion time of 10 min, the excess of chlorine was quenched usingsodium thiosulphate and samples concentrated as described inSection 2. Results obtained for PrP are shown in Fig. 1. Anoticeable diminution in the signal of the parent compoundwas observed for free chlorine concentrations over 0.4 mg L−1.
Within the investigated chlorine concentrations, the higherdegradation rates took place at pHs 7.3 and 8.0. This behaviourconfirms that, as it has already been described for other phenolicspecies, the anionic form of PrP (pKa 8.2) and the hypochlor-ous acid (pKa 7.5) react faster than the neutral paraben and thehypochlorite anion [6,9,17]. A similar trend to that presented inFig. 1 was observed for methyl, ethyl and butyl paraben, data notgiven.
3.2.2. Reaction kineticsTime course of paraben concentrations were followed using
ultrapure and tap water samples buffered at pH 7 and spiked,when necessary, with sodium hypochlorite at two different lev-els: 5.63 and 22.53 �M (equivalent to free chlorine concentra-tions of 0.4 and 1.6 mg L−1). Experiments were carried out using250 mL volume aliquots containing an initial concentration ofeach paraben between 9 and 10 �g L−1 (45–70 nM, dependingon their molecular weights). After a given time, the reaction wasstopped and the remaining concentration of the parent speciedetermined. Data at zero time were obtained by adding sodium
Table 3Retention times, proposed identities and most intense ions in MS spectra of paraben disinfection by-products
Parent compound By-products Retention time (min) m/z ratios of most intense ions with their relative abundances
MeP
3-ClMeP 17.19 243 (100), 245 (35), 163 (22)3,5-DClMeP 18.54 277 (100), 279 (71), 183 (12)3-BrMeP 18.06 287 (89), 289 (100), 119 (15)3,5-DBrMeP 20.21 363 (72), 365 (100), 367 (30)3-Br-5-ClMeP 19.38 321 (68), 323 (100)
EtP
3-ClEtP 17.88 257 (100), 259 (38), 149 (37)3,5-DClEtP 19.17 291 (100), 293 (71), 183 (34)3-BrEtP 18.69 301 (100), 303 (99), 149 (26)3,5-DBrEtP 20.79 379 (52), 381 (100), 383 (53), 313 (12)3-Br-5-ClEtP 20.01 335 (70), 337 (100)
PrP
3-ClPrP 18.82 271 (100), 273 (38), 149 (21)3,5-DClPrP 20.07 305 (100), 307 (70), 183 (15)3-BrPrP 19.60 315 (94), 317 (100), 149 (15)3,5-DBrPrP 21.60 393 (61), 395 (100), 397 (60)3-Br-5-ClPrP 20.83 349 (72), 351 (100)
BuP
3-ClBuP 19.76 285 (100), 287 (36), 149 (14)3,5-DClBuP 20.95 319 (100), 321 (74), 183 (10)3-BrBuP 20.50 329 (90), 331 (100), 149 (12)3,5-DBrBuP 22.43 407 (52), 409 (100), 411 (57)3-Br-5-ClBuP 21.69 363 (70), 365 (100)
110 P. Canosa et al. / Analytica Chimica Acta 575 (2006) 106–113
thiosulphate to chlorinated water samples before parabens. As ahigh molar excess of chlorine was used, the reaction rates onlydepended on the instantaneous concentration of the consideredparaben. Fig. 2 shows the graphs obtained for MeP. The loga-rithmic value of its concentration was plotted versus the reactiontime for ultrapure (Fig. 2A) and drinking water (Fig. 2B). Oneof the drinking water samples was obtained directly from thetap and contained 0.46 mg L−1 of free chlorine. The other wasprepared in the laboratory by addition of sodium hypochloriteto a tap water sample containing 0.2 mg L−1 of chlorine. In allcases, experimental data fitted quite well a straight line meaningthat, in presence of an excess of chlorine, the removal of MePfollows a pseudo-first-order kinetic. Moreover, similar half-liveswere obtained for ultrapure and tap water containing similar lev-els of free chlorine. The decrease in the concentrations of theother three parabens followed also pseudo-first-order kineticsand, moreover, they presented similar half-life values to thoseobtained for MeP (Table 2). As experiments were performedin different days, without controlling exactly the temperatureof the water samples (15 ± 2 ◦C) and, in addition, several stepsare involved in the sample preparation procedure, the observeddifferences among half-lives of the four parabens are probablynon-significant.
3.2.3. Disinfection by-productsTable 3 summarizes retention times and m/z ratios for the
most intense ions in the MS spectra of paraben by-products.Five transformation species were identified for each parent com-pound. MS spectra of the two major by-products presentedthe typical isotopic pattern of molecular ions correspondingto mono- and di-chlorinated compounds. Moreover, their basepeaks were shifted 34 and 68 m/z units in comparison to thoseappearing in the spectra of the respective parent compounds,figure not shown. Thus, they correspond to substitution of oneor two atoms of hydrogen per chlorine in the aromatic ring ofparabens. It is expected that chlorination takes place in both car-bons located in ortho- to the phenolic group, since the para-position is blocked with the ester moiety [17]. Comparison ofretention times obtained for the mono-chlorinated by-productsof MeP and EtP with those corresponding to standards of 3-ClMeP and 3-ClEtP, purchased from Aldrich and submitted tothe same silylation procedure, confirmed the above assump-tion, figure not shown. It was also verified that both standardsreact with free chlorine following pseudo-first-order kineticsand rendering the same di-chlorinated by-products than methyland ethyl paraben. Their half-lives for a chlorine concentrationof 0.4 mg L−1 were 31.2 and 31.9 min for 3-ClMeP and 3-ClEtP, respectively. Conversely to the limited stability of mono-chlorinated paraben by-products, the di-chlorinated ones wererather resistant to undergo further chlorine substitution reactionsor cleavage of the aromatic ring, even in presence of relativelyhigh concentrations of chlorine, Fig. 3. Therefore, if they aregenerated in a real life situation, their presence in the aquaticenvironment appears as feasible.
Experiments with tap water revealed the presence of threeadditional by-products for each paraben. On the basis of theinformation contained on their MS spectra, they were identi-
Fig. 3. Time-course of EtP and its chlorinated by-products in ultrapure waterbuffered at pH 7.3. Initial chlorine concentrations 0.4 mg L−1 (A), 1.6 mg L−1
(B) and 5 mg L−1 (C).
fied as brominated compounds, Fig. 4. As in the case of parentspecies and chlorinated transformation products, the bromi-nated species react with MTBSTFA rendering the correspondingdimethyl-tert-butylsilyl derivatives. Base peaks in their spectracorresponded to the loss of tert-butyl moiety and therefore theyappear at [M-57] m/z units. The most probable source of thesenew by-products is the existence of bromide traces in tap watersources. In presence of free chlorine, bromide is oxidized tobromine, which reacts with aromatic compounds through elec-trophilic substitution reactions. In order to verify this hypothesis,250 mL aliquots of a tap water sample (measured chlorine con-centration 0.53 mg L−1) were spiked with a mixture of the fourconsidered parabens, at 10 �g L−1, and allowed to react for20 min. The experiment was repeated using another aliquot ofthe same water sample spiked also with 1 �g L−1 of bromide,as potassium bromide. As depicted in Fig. 5 for ethyl paraben,
P. Canosa et al. / Analytica Chimica Acta 575 (2006) 106–113 111
Fig. 4. MS spectra and proposed structures for brominated by-products of EtP.
a diminution in the peak areas for the chlorinated species wasobserved, whereas, signals for the brominated by-products rosesignificantly. In seems that both, chlorine and bromine, competeto react with parabens through substitution reactions.
Further experiments were accomplished using ultrapurewater spiked with a fixed concentration of free chlorine(0.4 mg L−1) and increasing levels of bromide: 1, 10 and65 �g L−1. Ratios between peak areas of the mono and di-halogenated by-products showed that, for all parabens, thebromination reaction is much more favourable than the chlori-nation one (Table 4). As a consequence, brominated species areexpected to be the major paraben transformation by-products inchlorinated tap water prepared from aquifers placed in coastalregions or areas with salt deposits, which can contain up to1 mg L−1 of bromide [18–20]. Although, data regarding the toxi-
Fig. 5. Influence of bromide (addition level 1 �g L−1) on the formation of chlo-rinated and brominated derivatives of EtP. Experiments performed using tapwater (free chlorine concentration 0.53 mg L−1) buffered at pH 7.3. Reactiontime 20 min.
city of halogenated paraben derivatives were not found, it is wellaccepted that, normally, brominated by-products generated dur-ing disinfection of tap water represent a higher health risk thanthe chlorinated ones [21].
3.2.4. Paraben by-products from personal care productsand occurrence in the aquatic media
From toxicological and environmental perspectives, it is rel-evant to know if the above halogenation reactions can take placealso when paraben-containing products get in contact with chlo-rinated tap water. In this situation, the available chlorine couldbe exhausted by other chemicals included in the formulation ofpersonal care products and thus parabens might remain unaf-fected. In order to evaluate this possibility, a bath gel, purchased
Table 4Ratios between peak areas of chlorinated and brominated paraben by-productsin ultrapure water containing an initial chlorine concentration of 0.4 mg L−1 andincreasing levels of bromide
Parentspecies
Peak area ratios Initial bromideconcentration (�g L−1)
1 10 65
MeP3-ClMeP/3-BrMeP 2.19 0.13 <10−5
3,5-DClMeP/3,5-DBrMeP 9.67 0.02 2.3 × 10−4
EtP3-ClEtP/3-BrEtP 2.51 0.16 <10−5
3,5-DClEtP/3,5-DBrEtP 10.32 0.02 <10−5
PrP3-ClPrP/3-BrPrP 2.43 0.12 <10−5
3,5-DClPrP/3,5-DBrPrP 12.72 0.02 1.7 × 10−4
BuP3-ClBuP/3-BrBuP 2.71 0.15 <10−5
3,5-DClBuP/3,5-DBrBuP 10.53 0.02 <10−5
Reaction time 20 min; pH 7.3.
112 P. Canosa et al. / Analytica Chimica Acta 575 (2006) 106–113
Fig. 6. Time-course of BuP brominated by-products after mixing a bath gel withtap water containing 0.82 mg L−1 of free chlorine.
from a local market, was first diluted 10 times, using ultrapurewater, and then added to 250 mL aliquots of ultrapure (reference)and tap water samples (containing 0.82 mg L−1 of free chlorine),buffered at pH 7.3. Two series of experiments were carried outwith samples thermostated at 15 and 38 ◦C. After a given time,from 0 to 30 min, chlorine was removed and samples processedas described previously. Reference experiments revealed that thebath gel contained only non-halogenated parabens. Their con-centrations ranged from 200 to 2240 �g g−1. On the other hand,assays with tap water confirmed the formation of all describedchlorinated and brominated by-products, for the four parabenspresented in the gel, even considering reaction times as short as2 min. The decrease in the signals of parent species was about1.5-folds faster at 38 ◦C (average half-life 5.9 min) than at 15 ◦C(average half-life 8.5 min). Moreover, as shown in Fig. 6 forBuP, the formation of the brominated derivatives was enhancedat 38 ◦C, a rather common temperature in a water bath. Anyhow,they remained as minor by-products in comparison to the chlo-rinated ones. Additional experiments using different personalcare products, e.g. mouth rinse solutions, confirmed the halo-genation of those parabens included in their composition, when
Fig. 7. Chromatogram for a non-spiked wastewater sample (code 3, Table 5).GC–MS plots for the di-chlorinated by-products of MeP (m/z 277 + 279) andPrP (m/z 305 + 307).
mixed with tap water, data not given. Standards for most of theby-products identified in this study are not commercially avail-able; therefore, the yield of the described reactions could notbe calculated. However, considering a semi-quantitative esti-mation, chlorinated parabens might justify up to 80% of thedecrease observed in the concentrations of the correspondingparent compounds.
From the best of our knowledge, halogenated parabens do nothave any commercial application; therefore, their presence in theaquatic environment will serve as a further evidence of trans-formation reactions described in this work. Fig. 7 depicts theGC–MS traces of the di-chlorinated MeP and PrP by-products,
Table 5Concentrations of MeP and PrP paraben (�g L−1) in different raw sewage watersamples
Samplecode
Concentration (�g L−1) ± S.D. Peak area ratios (%)
MeP PrP 3,5-DClMeP/MeP (%)
3,5-DClPrP/PrP (%)
1 n.q. 0.98 ± 0.09 n.d. 18.42 0.06 ± 0.01 10.84 ± 0.64 520.7 6.63 0.49 ± 0.05 12.96 ± 0.65 323.0 15.94 n.q. 0.94 ± 0.08 n.d. 32.05 1.88 ± 0.12 0.67 ± 0.06 4.2 3.4
Ratios between peak areas for their di-chlorinated by-products and parentspecies. Sample volume: 1 L, n = 3 replicates; n.q. under quantification limits;n.d. non-detected.
P. Canosa et al. / Analytica Chimica Acta 575 (2006) 106–113 113
together with their MS spectra, for a raw wastewater sample con-centrated 500-folds (sample number 3, Table 5). Table 5 gives theconcentrations of MeP and PrP in five raw wastewater samples,as well as the ratios between responses (peak areas) obtained fortheir di-chlorinated by-products and the corresponding parentspecies. In two of the considered samples, peak areas for 3,5-DClMeP were higher than those corresponding to MeP. Otherby-products of both parabens were not detected. EtP and BuPremained around the detection limits of the method in all sam-ples and none of their halogenated derivatives was found.
4. Conclusions
Parabens react with free chlorine rendering several halo-genated by-products. Chlorine levels usually contained in tapwater are enough to produce significant amounts of their chlo-rinated by-products in a few minutes. Therefore, consideringthe extensive employment of parabens in personal care prod-ucts, daily activities such as showering and bathing constitute asource of dermal exposition to paraben chlorinated by-products.Moreover, the di-chlorinated derivatives are highly resistant toundergo further oxidation reactions and their presence has beenfound for first time in sewage water samples. If minimal amountsof bromide are presented in tap water sources, halogenationreactions are shifted to the production of brominated parabens.Further studies are necessary to: (i) evaluate the potential heathrisks and possible endocrine disrupter activity of halogenatedparaben by-products, (ii) to know their fate in the environmentand (iii) to study the kinetic of paraben oxidation reactions usingalternative disinfectants such as ozone or chlorine dioxide.
Acknowledgements
Financial support from project DGICT CTQ2005-00425 andtechnical assistance from LABAQUA with free chlorine deter-
minations in water samples is acknowledged. PC acknowledgesa FPU grant from the Spanish Ministry of Education.
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