Plant-microbe-soil interactions and their role in phytotechnologies applied to trace metal-rich soils

DEPARTAMENTO DE EDAFOLOXÍA E QUÍMICA AGRÍCOLA Universidade de Santiago de Compostela INSTITUTO DE INVESTIGACIÓNS AGROBIOLÓXICAS DE GALICIA Consejo Superior de Investigaciones Científicas Plant-microbe-soil interactions and their role in phytotechnologies applied to trace metal-rich soils Vanessa Álvarez López Tese de Doutoramento, Xuño de 2016 Las Dra. Petra S. Kidd y Dra. Ángeles Prieto Fernández, Científicas Titulares del Consejo Superior de Investigaciones Científicas, adscritas al Instituto de Investigaciones Agrobiológicas de Galicia y la Dra. Mª del Carmen Monterroso Martínez, Profesora Titular del Departamento de Edafología y Química Agrícola de la Universidad de Santiago de Compostela INFORMAN: Que la presente memoria titulada “Plant-microbe-soil interactions and their role in phytotechnologies applied to trace metal-rich soils”, presentada por Dña Vanessa Álvarez López para optar al Grado de Doctora en Biología, fue realizada bajo nuestra dirección. Y considerando que representa trabajo de Tesis de Doctorado, autorizamos su presentación ante el Tribunal correspondiente. Y para que así conste, firmamos el presente informe en Santiago de Compostela a 1 de junio de 2016. Fdo.: Dra. Petra S. Kidd Fdo.: Dra. Ángeles Prieto Fernández Fdo.: Dra. Mª del Carmen Monterroso Marnez A meus pais A Cristina 5(680(1 iii Resumen INTRODUCCIÓN El término elemento traza se refiere a aquellos elementos que se encuentran en concentraciones menores a 0.1% (1000 mg kg-1) en el medio ambiente o menores a 0.01% (100 mg kg-1) en tejidos animales y vegetales (Adriano, 2001). Las denominaciones metales traza (que hace referencia a su abundancia y que incluye metales y metaloides) o metales pesados (que indica que sus densidades son > 5 g cm-3) también se usan con frecuencia para estos elementos y se consideran sinónimos (Kabata-Pendias, 2010). Algunos metales traza (como el Cu, Fe, Mn, Ni o Zn) son micronutrientes esenciales y son imprescindibles en determinaGRVSURFHVRVFHOXODUHVSRUHOFRQWUDULRDRWURVFRPR$V&G+JR3EQRVHOHV conoce ninguna función biológica y se consideran no-esenciales. En ambos casos, un exceso de estos elementos afecta severamente a procesos biológicos en organismos vivos lo que causa un riesgo importante para el medio ambiente y la salud humana (Cuypers et al., 2010). Suelos naturalmente ricos en metales Sin tener en cuenta la actuación del ser humano, la concentración de elementos traza en los suelos depende del material de partida y de sus procesos de meteorización. Los tipos más frecuentes de suelos naturalmente ricos en metales son: (i) suelos serpentiníticos que derivan de rocas ultramáficas ricas en Fe y Mg y enriquecidas en Ni, Cr y Co (ii) suelos calamínicos, ricos en Pb y Zn (iii) suelos enriquecidos en Cu y Co que derivan de argilitas y dolomitas y (iv) suelos derivados de rocas enriquecidas en Se. Los suelos serpentiníticos se forman a partir de la meteorización y disolución de rocas ultramáficas ricas en minerales ferromagnesianos (70%) y con un contenido en sílice (SiO2) menor del 45% (Kruckeberg, 2002). Estos suelos ocupan alrededor del 1% de toda la corteza terrestre (Proctor, 1999) y todos comparten algunas características específicas como son una relación Ca:Mg < 1, déficit de nutrientes esenciales (N, K o P) y elevadas cantidades de elementos traza como Ni, Cr o Co (Brooks, 1987). Estas propiedades junto con sus condiciones físicas dan lugar al denominado síndrome serpentinítico y proporcionan un ambiente inhóspito para la mayoría de las plantas, lo que ha dado lugar al desarrollo de adaptaciones fisiológicas y morfológicas diferentes a las de la flora de las áreas circundantes y a la presencia de altas tasas de endemismos (Whittaker, 1954). Entre la vegetación característica de estos suelos se encuentran las plantas iv hiperacumuladoras, que se caracterizan por alcanzar concentraciones extremas de elementos traza en sus tejidos aéreos cuando crecen en sustratos ricos en metales (Brooks, 1987). Actualmente, se conocen aproximadamente 500 taxones de plantas hiperacumuladoras de las cuales 450 acumulan Ni (Pollard et al.YDQGHU Ent et al., 2015). Una de las familias más ampliamente representada es la Brassicaceae que incluye el género Alyssum, en el cual se engloba aproximadamente el 90% de las especies hiperacumuladoras de Ni (Pollard et al., 2014). En la Península Ibérica se encuentran dos subespecies endémicas de Alyssum serpyllifolium Desf. (Brassicaceae) hiperacumuladoras de Ni: Alyssum serpyllifolium ssp. lusitanicum (conocido como A. pintodasilvae), que crece de forma natural en el afloramiento serpentinítico de Barazón (NO España) y de Trás-os-Montes (NE Portugal) (Díez-Lázaro et al.  )UHLWDV et al.  Pinto da Silva, 1970) y Alyssum serpyllifolium ssp. malacitanum (conocido como A. malacitanum) (López-González, 1975), que crece en las zonas serpentiníticas de Sierra Bermeja y de Sierra de Aguas (S España). Contaminación de suelos por metales traza Debido a que la formación del suelo es un proceso extremadamente lento y a que realiza funciones vitales se puede considerar como un recurso esencial y no renovable1. La Estrategia Temática para la Protección del Suelo de la Unión Europea incluye la contaminación del suelo entre los principales procesos de degradación. Su contaminación se define según la Comisión Europea como “la presencia de contaminantes en el suelo por encima de un cierto nivel que causa un deterioro o pérdida de una o más de sus funciones”2. Los datos recogidos por la Agencia Europea para el Medioambiente (EEA) estiman que el número de lugares potencialmente contaminados en la Unión Europea asciende a 2,5 millones y el de lugares contaminados a 3420003. En general, el sector de la producción contribuye más a la contaminación del suelo (60%) que el sector servicios (32%) y las actividades mineras junto con las industrias de metal son fuentes importantes de contaminación local (13%). En cuanto al tipo de contaminantes, los más frecuentes son aceites (24%) y metales pesados (35%) (Panagos et al., 2013). 1http://ec.europa.eu/environment/soil/index_en.htm 2http://eusoils.jrc.ec.europa.eu/library/themes/contamination/ 3EEA,http://www.eea.europa.eu/data-and-maps/indicators/progress-in-management-of-contaminated -sites-3/assessment v Resumen La mayoría de los países europeos poseen una legislación a nivel nacional o regional para regular la contaminación del suelo. Actualmente, en España, la contaminación del suelo está regulada por la Ley 5/2013 de Suelos Contaminados (BOE no. 140, 12 de Junio de 2013) y el Decreto Real 9/2005, ambos incluidos en el II Plan Nacional de Suelos Contaminados (2007-2015). Estos proporcionan una lista de actividades potencialmente contaminantes y los criterios para declarar un suelo como contaminado además de los Niveles Genéricos de Referencia (NGRs), que se definen como la concentración de un contaminante que no implica un riesgo para el ser humano y naturaleza. Los límites de metales incorporados al suelo en lodos de depuradora y compost destinados a la agricultura se regulan por el Real Decreto 1310/1990 (Ministerio de Agricultura 1990) y por la Ley española de fertilizantes (Real Decreto 824/2005 (Ministerio de Agricultura 2005)). Sin embargo, no existe todavía una legislación para el uso de enmiendas en suelos marginales o degradados. Disponibilidad de metales y absorción por las plantas La (bio)disponibilidad de los metal(oid)es se define como la fracción que puede interactuar con una diana biológica (Geebelen et al., 2003). Las plantas obtienen los nutrientes y los metales traza desde la solución del suelo que está en contacto cercano con las raíces (definido por Hiltner en 1904 como rizosfera) y su disponibilidad está gobernada por un pseudo-equilibrio entre esta solución y la fase sólida (Kidd et al., 2009). La cantidad de metal disponible está determinada, además de por la concentración total en la solución del suelo, por la capacidad de la fase sólida de reponer el elemento a la solución del suelo (Bruemmer et al., 1986). La reposición de los elementos incluye disociación de complejos presentes en la solución del suelo, pero también liberación de elementos lábiles asociados con la fase sólida, y está controlada por el tamaño de la fracción de metal lábil, su transporte y la velocidad a la que se libera desde la fase sólida (Lehto et al., 2006). Las plantas modifican activamente las propiedades físicoquímicas de la rizosfera para movilizar nutrientes y/o metales traza. Los procesos que determinan la especiación de los metales en el suelo, su distribución y su disponibilidad están controlados por una combinación de las propiedades del suelo, las características de las plantas y la interacción con sus microorganismos asociados (Jones et al., 2004). vi La identificación de los mecanismos específicos que controlan la disponibilidad y absorción de metales en la rizosfera de las hiperacumuladoras puede ayudar a la mejora de los métodos de recuperación de suelos contaminados usando estas plantas. Aunque existen estudios que indican que la alta acumulación de metal por hiperacumuladoras es debida a que movilizan activamente los metales desde fracciones no disponibles (McGrath et al..LGG et al.$OIRUG et al., 2010), la mayoría de los estudios demuestran que las plantas hiperacumuladoras acceden a las mismas fracciones que las no hiperacumuladoras (Salt et al., 2000). Otros estudios más recientes sugieren que la liberación de ligandos orgánicos (como citrato) por las raíces de las hiperacumuladoras podría acelerar la disolución de los minerales ricos en Ni y mantener así una alta tasa de reposición GHVGHODVIUDFFLRQHV PHQRVOiELOHV%RRPLQDWKDQ\'RUDQ:HQ]HO et al., .LGG et al., 2009). Finalmente, parece probable que las hiperacumuladoras no posean mecanismos específicos de movilización de metales sino que la disminución pronunciada de Ni que ocurre en su rizosfera como consecuencia de la alta tasa de acumulación por las plantas induzca su reposición desde fracciones menos móviles a otras más móviles. Existe una gran variedad de métodos propuestos para evaluar la disponibilidad de los metal(oid)es en el suelo. Las técnicas tradicionales y más comúnmente utilizadas para estimar la fracción disponible de metales incluyen métodos químicos basados en extracciones simples y/o secuenciales (Michel y Ludwig, +DVV\)LQH6LQHPEDUJRODIUDFFLyQOiELOGHPHWDOHVWLPDGDXVDQGR estas técnicas no se correlaciona con la acumulación por la planta. La técnica de DGT (del inglés Diffusive Gradients in Thin Films) ha sido desarrollada como una técnica pasiva para medir in situ el pseudo-equilibro que ocurre entre los metales en la disolución del suelo y la fracción intercambiable (Davison y Zhang, 1994), imitando el efecto de la planta. La técnica de la DGT se basa en la acción de una resina embebida en un gel que adsorbe continuamente iones desde la solución del suelo, lo que genera un gradiente entre el suelo y la resina. Mientras está en contacto con el suelo, la resina induce un flujo continuo de metales y la disociación desde la fase sólida para compensar la depleción en la solución en contacto con el gel. Se ha observado que la fracción de metal disponible estimada usando esta técnica se correlaciona mejor con la acumulación por parte de las SODQWDV TXH OD PHGLGD FRQ H[WUDFFLRQHV TXtPLFDV 'DYLVRQ \ =KDQJ  Degryse et al., 2009). vii Resumen Las medidas realizadas por DGT se pueden incorporar a modelos numéricos para estimar los parámetros que cuantifican la reposición de los metal(oid)es desde la fase sólida a la solución del suelo (Harper et al.(UQVWEHUJHU et al., 2002). El modelo de DIFS (del inglés DGT Induces Fluxes in Sediments and Soils) se ha desarrollado para cuantificar los procesos que influencian el flujo de elementos hacia el gel de la DGT, es decir, para describir la interacción del gel con el suelo. Este modelo estima la proporción de elementos, en la fase sólida (KD) que pueden liberarse y acumularse en el gel, y el tiempo de respuesta (TC) al que se alcanza el equilibrio entre la fase sólida y la solución del suelo (Ernstberger et al., 2002). Tecnologías de recuperación de suelos respetuosas con el medio ambiente En las últimas décadas se ha documentado un incremento en el uso de tecnologías verdes respetuosas con el medio ambiente (GRO, del inglés Gentle remediation options) para la recuperación de suelos contaminados. Estas opciones incluyen la estabilización del contaminante in-situ y las técnicas de fitocorrección que se basan en el uso de plantas y sus microrganismos asociados para la recuperación de suelos contaminados y se consideran más rentables y menos invasivas que las técnicas convencionales de ingeniería civil (Vangronsveld et al., 0HQFK et al., 2010). Las tecnologías más comúnmente utilizadas en el caso de contaminación con metales pesados son: fitoestabilización, que incluye el uso de plantas exclusoras de metales en combinación con enmiendas del suelo para UHGXFLUODWUDQVIHUHQFLDGHOFRQWDPLQDQWHDODFDGHQDWUyILFD\ILWRH[WUDFFLyQTXH implica el cultivo de plantas acumuladoras de metales que se cosechan al final de su ciclo de crecimiento retirando de esta forma el metal del suelo. Cuando el metal fitoextraído se puede recuperar y comercializar, obteniendo un beneficio económico, el proceso se denomina fitominería (Chaney et al.  7DQJ et al., 2012). El éxito de las técnicas de fitocorrección es altamente dependiente de la solubilidad y la especiación de los metal(oid)es en el suelo y de la tolerancia de las plantas y su capacidad de acumular los elementos, por lo tanto, una adecuada selección de las plantas para cada caso es de vital importancia. En las técnicas de fitoextracción es deseable el uso de plantas con alta capacidad de acumulación de los elementos traza mientras que para la fitoestabilización se requieren plantas que excluyan los metales de su parte aérea. viii Mejora de la eficiencia de técnicas de fitocorrección Las plantas seleccionadas en técnicas de fitocorrección para la revegetación de suelos contaminados deben tolerar, además de la contaminación, otros numerosos factores de estrés tales como deficiencia de agua y nutrientes, acidez o salinidad, erosión o compactación, inundaciones, herbivoría o plagas. Las áreas contaminadas acogen una importante variedad de genotipos vegetales tolerantes a metales y adaptados a las condiciones físico químicas y climatológicas de cada sitio en particular (Batty, 2005). Las plantas hiperacumuladoras tienen una gran relevancia en el campo de la fitoextracción debido a su potencial aplicación en técnicas de fitominería. Actualmente, el Ni y el Co son los metales que presentan HOSUHFLRPiVDOWRHQHOPHUFDGRVLQHPEDUJRGHELGRDODOWRQ~PHUR \DPSOLD distribución de las hiperacumuladoras de Ni, este parece ser el elemento que puede ser más fácilmente explotado mediante fitominería (Chaney et al., 2014). A pesar de que las hiperacumuladoras tiene características muy adecuadas para su empleo en fitoextracción, su aplicación todavía es limitada debido a su baja producción de biomasa o al desconocimiento sobre su cultivo (Kidd et al., 2015). Ante esta limitación, se han descrito cultivos de alta biomasa como especies anuales de alto crecimiento (p.ej. girasol (Helianthus annus) o tabaco (Nicotiana tabacum)) o especies leñosas (p.ej. Salix spp. o Populus spp.), como una alternativa viable, particularmente para su uso en suelos contaminados con Cd, Zn o Se. Entre las ventajas de estas especies cabe mencionar que se pueden cultivar fácilPHQWH XWLOL]DQGR SUDFWLFDV DJURQyPLFDV \D HVWDEOHFLGDV *KRVK \ 6LQJK  Meers et al., 2005) y que su desarrollado sistema radicular permite la exploración de un mayor volumen de suelo y un crecimiento rápido. En cualquier caso, el cálculo del tiempo necesario para la recuperación del suelo es demasiado elevado y por lo tanto el proceso solo sería aceptable si se combina con producción de biomasa para bioenergía o como materia prima. En el caso de las técnicas de fitoestabilización, a menudo incluyen revegetación con plantas tolerantes y también incorporación de enmiendas inorgánicas y/u orgánicas para mejorar las propiedades físicas, químicas y biológicas del suelo (Vangronsveld et al.%DUUXWLD et al., 2011). Las enmiendas que se utilizan de forma más habitual incluyen agentes de encalado, fosfatos, oxihidróxidos de Fe y Mn, materiales orgánicos, zeolitas, cenizas… (Adriano et al.%RODQ et al., 2014). Estas enmiendas pueden reducir la solubilidad de los metal(oid)es por la formación de precipitados insolubles o aumentando la capacidad de adsorción del suelo. Además, en el caso de la materia orgánica, su uso puede mejorar las propiedades físicas del suelo como la ix Resumen infiltración y retención de agua, proporcionar microy macro-nutrientes para el crecimiento vegetal y estimular la actividad microbiana. Debido a que la deficiencia de nutrientes en los suelos contaminados es frecuentemente uno de los factores más limitantes para el desarrollo vegetal (Ye et al., 2002), la fertilización de este tipo de suelos es un paso crucial para el establecimiento de una cobertura vegetal. Los regímenes de fertilización pueden estar diseñados para mejorar el crecimiento pero, en el caso de la fitoextracción, también para aumentar la acumulación de los elementos traza ya que la fertilización inorgánica puede alterar la disponibilidad de los elementos traza en el suelo (McLaughlin et al.6LQJK et al., 2011). El diseño de patrones de cultivo apropiados puede mejorar significativamente el crecimiento de las plantas y su nutrición, además de modificar la solubilidad de los metales en el suelo (Kidd et al., 2015). El cultivo intercalado de plantas se basa en interacciones interespecíficas entre raíces, que puede resultar en una mejora en la disponibilidad de los nutrientes y, por lo tanto, aumentar la productividad de los cultivos (Wieshammer et al., 2007). Por otra parte, el intercalado de distintas especies puede alterar las condiciones en las rizosferas y afectar la disponibilidad de los metales a las plantas vecinas (Tang et al., 2012). Además de por una razón práctica (evitar el agotamiento de nutrientes o pestes relacionadas con una especie determinada), el cultivo el intercalado con legumbres herbáceas (p.ej. Lotus spp.) o leñosas (e.g. Alnus spp.) puede mejorar la fertilidad del suelo ya que son fijadoras de N2 y además algunas especies, como Lupinus spp., son también capaces de solubilizar P (Vance et al./LX et al., 2013). El uso de bacterias promotoras del crecimiento vegetal (PGPB, del inglés plant growth promoting bacteria) se ha descrito también como una opción para mejorar el establecimiento y crecimiento vegetal en suelos contaminados (Weyens et al.  6HVVLWVFK et al., 2013). Las bacterias asociadas a plantas incluyen bacterias endofíticas (que colonizan el interior de los tejidos vegetales sin causar efectos negativos en la planta), bacterias de la filosfera (que se encuentran en las superficies externas de las plantas) y bacterias rizosféricas (que habitan en el suelo en contacto con las raíces e influenciado por la planta). El uso de bacterias como biofertilizantes (p.ej. bacterias fijadoras de N2 atmosférico, o capaces de solubilizar P o Fe) es una práctica común en agricultura para evitar el uso excesivo de fertilizantes químicos (Weyens et al., 2009). Además, existen bacterias también capaces de producir fitohormonas que estimulan el crecimiento xvi de tabaco, sobre todo en suelos enmedados con compost en los que todos los inóculos (P64, P87, P30, P75 y P29) aumentaron significativamente su biomasa. En el caso de S. caprea la inoculación bacteriana tuvo un menor efecto y solo la cepa P30 causó un ligero aumento en su crecimiento. La inoculación bacteriana también mejoró algunos índices de eficiencia fotosintética como CHL o Fv / Fm en plantas de tabaco creciendo en suelo sin compost. Los inóculos bacterianos también influenciaron la disponibilidad de metales en el suelo y este efecto fue más pronunciado en los suelos sin enmienda orgánica. Todas las cepas en combinación con plantas de tabaco aumentaron el Zn y Cd disponible en estos suelos. En el caso de los suelos plantados con Salix, la variación en la disponibilidad de metales dependió del inóculo y de la presencia de compost. Las cepas bacterianas inoculadas también modificaron el fraccionamiento de metales en ambos suelos. En cuanto a la acumulación de metal en planta, en suelos sin enmienda, las cepas P30, P75 y P29 disminuyeron la concentración de metales en las plantas de tabaco y las cepas P64 y P87 aumentaron la concentración de Zn. En las plantas creciendo en suelo con compost la inoculación bacteriana no causó efectos significativos en la concentración de Zn mientras que las cepas P30 y P29 disminuyeron la concentración de Cd. En el caso de Pb, en general las cepas tendieron a aumentar su acumulación. En las plantas de S. caprea, se los efectos inducidos por las bacterias fueron más pronunciados en las hojas. Así por ejemplo en plantas desarrolladas en suelo sin enmienda, la cepa P87 aumentó la acumulación de Cd y Zn en hojas. Los efectos de las bacterias en la cantidad total de metal fitoextraído dependieron del tipo de suelo en el que crecieron las plantas. En el caso de las plantas de tabaco, en de los suelos sin enmendar, la cantidad de metal extraído se vio negativamente afectada por la inoculación bacteriana, pero en las plantas creciendo en suelos enmendados con compost, todos los inóculos mejoraron la capacidad de extracción de metales. Por el contrario, en el caso de S. caprea, los efectos beneficiosos de la inoculación bacteriana fueron más pronunciados en el suelo sin tratar, mientras que en el suelo con compost, en general, los inóculos redujeron la cantidad de Zn y Cd fitoextraído. 2.c. El efecto de la inoculación de la cepa P30 de Rhodococcus erythropolis en el crecimiento de N. tabacum dependió tanto de la densidad celular del inóculo como del modo de inoculación (en semilla, en suelo, aplicada una o dos veces, o aplicada en semilla y en suelo). El tratamiento que produjo un mayor aumento en la biomasa de las plantas de tabaco fue una única inoculación en suelo con la xvii Resumen densidad de 106 UFCs mL-1 (hasta 45% de aumento). La inoculación repetida en suelo usando la densidad de 106 UFCs mL-1 también aumentó la producción de biomasa pero solo hasta un 9%. En cuanto a la acumulación de metales por las plantas de tabaco, los cambios detectados dependieron más del modo de inoculación que de la densidad del inóculo. El aumento de densidad desde 106 UFCs mL-1 a 108 UFCs mL-1 no dio lugar a una mejora en los efectos. La inoculación bacteriana afectó en menor medida a las concentraciones de metales en la parte aérea que a la producción de biomasa, aunque se observó un efecto positivo en la acumulación de Pb con el inóculo 108 UFCs mL-1 aplicado en suelo (29% de aumento) o en semilla y suelo (34% de aumento). La acumulación de Zn también aumentó hasta un 24% después de la inoculación en semilla con una densidad de 106 UFCs mL-1, mientras que en el caso de la acumulación de Cd, ningún tratamiento indujo un aumento significativo. Los tratamientos de inoculaFLyQWDPELpQDIHFWDURQDODGLVSRQLELOLGDGGHPHWDOHQHOVXHORVLQHPEDUJRORV aumentos en la disponibilidad inducidos por las bacterias no siempre se vieron reflejados en un aumento de acumulación de metales por las plantas. En general, la inoculación con la cepa P30 tuvo un efecto más beneficioso en el crecimiento de la planta que en la acumulación de metal. Todos los WUDWDPLHQWRVDXPHQWDURQODFDQWLGDGWRWDOGHPHWDOILWRH[WUDtGRSRUODVSODQWDVVLQ embargo, los mayores incrementos en la cantidad total de metal extraído no se correlacionaron con un aumento en el número de bacterias aplicadas. De hecho, los mejores resultados se obtuvieron después de una inoculación única en el suelo utilizando la densidad celular más baja (106 UFCs mL-1). 3.a. Los suelos de las escombreras de la mina de Touro presentaron un pH ácido y baja fertilidad (reflejada en un bajo contenido total de C y N o P disponible). La capacidad de intercambio catiónico (CIC) fue también baja y dominada por H+ y Al 3+. Las concentraciones de elementos traza fueron muy variables en las escombreras pero, en general, se encontraron elevadas concentraciones de Al, Cr, Fe, Mn o Ni y los valores más elevados de metales disponibles (extraídos con NaNO3) correspondieron al Cu (media de 61 mg kg-1 y rango entre 7 y 219 mg kg-1) y Mn (media de 823 mg kg y rango desde 650 a 1142 mg kg-1). Un año después de la adición de compost, el pH del suelo aumentó significativamente y este aumento continuó observándose durante los dos siguientes años, y también se observó un aumento progresivo en C y N con el tiempo, encontrándose los máximos valores a los dos años de añadir la enmienda de compost. Asimismo, el compost causó cambios significativos de la CIC, que aumentó significativamente xviii un año después del tratamiento y descendió de nuevo después de tres años a valores similares a los que se determinaron en los suelos sin enmienda. A pesar de esto, mientras que antes de la adición del compost, la CIC estaba dominada por H+ y Al3+ (con una saturación en base menor al 20%), tres años después del tratamiento la saturación en bases fue superior al 90% (dominada principalmente por Ca2+). El P disponible también aumentó significativamente después de la adición del compost (hasta 100 veces un año después de la aplicación de la enmienda). Al mismo tiempo, la adición de compost disminuyó significativamente la disponibilidad de elementos traza como Cu, Al, Cd, Co, Cr, Ni, Zn o Mn. La actividad HQ]LPiWLFDIXHSUiFWLFDPHQWHQXODHQHOVXHORVLQHQPHQGDUVLQHPEDUJRXQDxR después de la adición de compost, la actividad de las enzimas hidrolasas aumento hasta 23 veces y los aumentos fueron incluso más importantes en ureasa y arilsulfatasa. En general, las actividades enzimáticas altas se mantuvieron durante los tres años de experimento. El crecimiento vegetal mostró una alta variabilidad entre los individuos de la misma especie a lo largo de la parcela experimental. El máximo crecimiento en la primera temporada de crecimiento se observó en las plantas de S. viminalis (106 cm), que mostraron también la mayor altura en el momento de la cosecha después de cuatro años de crecimiento (media de 219 cm con plantas que alcanzaron hasta 390 cm). Después de cuatro años el porcentaje de mortalidad fue mayor y más variable en Salix spp. que en P. nigra, por consiguiente, la biomasa final obtenida fue muy variable entre sub-parcelas. El valor medio de biomasa obtenida para Salix spp. fue de 7,0 ±3.0 t ha-1 en S. caprea y 4.3 ±2.3 t ha-1 en S. viminalis. La producción de biomasa de P. nigra fue más uniforme con una media de 4.3 ±0.7 t ha-1. La gramínea A. capillaris presentó un buen establecimiento y cobertura después de su siembra y la producción aérea de biomasa fue de 2.2 t ha-1 en la primera cosecha (dos años de crecimiento) y aumentó hasta 4.6 t ha-1 en la segunda cosecha (tres años de crecimiento). La concentración de nutrientes en los tejidos aéreos se mantuvo dentro del rango considerado beneficioso para el crecimiento vegetal y la concentración de Cu no excedió el umbral considerado como fitotóxico. Los principales efectos en los parámetros fisicoquímicos del suelo fueron un resultado directo de la adición de compost, mientras que los efectos inducidos por el crecimiento de las plantas se observaron en los parámetros bioquímicos. A pesar de esto, en comparación con los no plantados, en los suelos plantado también se observaron mejoras en parámetros como el pH, la CIC o la disponibilidad de Cu . 3.b. Los suelos de la mina de Rubiais presentaron un pH neutro o xix Resumen ligeramente alcalino (7.1-8.6) y baja fertilidad, reflejada en una baja cantidad de C, N y P disponible. Las concentraciones de metales en la zona de estudio fueron altamente variables a lo largo de las parcelas pero, en general, se encontraron altas concentraciones de Cd, Pb y Zn. La adición de 5% (p/p) de compost aumentó significativamente la cantidad de nutrientes, a la vez que disminuyó la cantidad de metales disponibles en el suelo (debido a un cambio en la distribución de los metales en el suelo hacia fracciones menos disponibles asociadas a óxidos de Fe y Mn, materia orgánica y la fracción residual). La adición de compost mejoró también las propiedades bioquímicas del suelo (como demuestra el aumento en la actividad enzimática). En los suelos sin enmienda, el crecimiento de ambas especies de Salix spp., provocó un aumento de C y N en el suelo comparado con los suelos sin plantar. En el caso de plantas desarrolladas en los suelos fertilizados con NPK se observó una disminución significativa en el pH del suelo, al mismo tiempo que un aumento en la biodisponibilidad de Cd y Zn. En el caso de plantas de S. smithiana en cocultivo con A. glutinosa, en los suelos enmendados se observó también un aumento en la disponibilidad de metales. En general, las actividades enzimáticas fueron mayores en los suelos con planta y no dependieron de la especie de Salix XWLOL]DGD VLQ HPEDUJR HQFRPSDUDFLyQ FRQ HO PRQRFXOWLYR HO FR-cultivo de S. smithiana con A. glutinosa, en general, aumentó la actividad enzimática. En el caso de la hiperacumuladora N. caerulescens, su crecimiento no modificó la cantidad de C, N o P en los suelos sin enmienda, pero aumentó significativamente estos nutrientes en los suelos con compost (comparado con el suelo sin plantar) y disminuyó la cantidad de Zn disponible tanto en los suelos sin enmienda como en los enmendados. De forma similar a lo observado en el caso de las especies leñosas, N. caerulescens aumentó la actividad enzimática del suelo y, en algunos casos, este aumento fue más pronunciado en el co-cultivo con L. corniculatus que con la hiperacumuladora en monocultivo. Los ejemplares de S. smithiana mostraron un buen establecimiento y baja mortalidad (4-28%), mientras que en el caso de S. atrocinerea, se encontró una menor supervivencia y las plantas fueron solo capaces de establecerse en el suelo sin enmienda y en monocultivo (en este tratamiento se observó un 84% de supervivencia). El crecimiento de individuos de una misma especie fue altamente variable a lo largo de la parcela experimental. Los individuos de S. smithiana (media de 82 cm después de cuatro años de crecimiento) fueron siempre mayores que S. atrocinerea (media de 42 cm después de cuatro años de crecimiento) y se xx observó un efecto positivo de la fertilización al cabo de tres años en el crecimiento de las plantas. El co-cultivo con A. glutinosa provocó también un aumento en la altura de las plantas de S. smithiana. En cuanto a la acumulación de metales por las especies de Salix, las máximas concentraciones de Cd y Zn se observaron en hojas de S. smithiana desarrolladas en suelo sin enmendar (21.4 mg kg-1 y 4248 mg kg-1 respectivamente), y la adición de compost causó una reducción en la cantidad de metales acumulados por las plantas (la máxima concentración en hojas de S. smithiana fue 11.3 y 1782 mg kg-1 de Cd y Zn respectivamente). N. caerulescens mostró un buen establecimiento en ambos tratamientos (con y sin enmienda) y una baja mortalidad (4-24%). El compost tuvo un efecto importante en el crecimiento de las hiperacumuladoras, aumentando significativamente su cobertura y producción de biomasa (desde 72 g m-2 hasta 210 g m-2 en monocultivo), mientras que el co-cultivo con la leguminosa disminuyó la cantidad de biomasa producida en los suelos sin enmendar. En los suelos sin compost, la máxima concentración de Cd y Zn se encontró en las hojas de N. caerulescens durante el periodo de floración (hasta 1350 mg kg-1 de Cd y hasta 22052 mg kg-1 de Zn). Del mismo modo que en el caso de las especies leñosas, la adición de FRPSRVWGLVPLQX\yODDFXPXODFLyQGHPHWDOHVSRUODVSODQWDVVLQHPEDUJRHOFRcultivo con L. corniculatus mejoró tanto el estado nutricional de las plantas como la acumulación de metales. La cantidad total de Zn extraído fue mayor en los suelos con compost (desde 586 mg m-2 hasta 998 mg m-2), mientras que la enmienda no causó un efecto en la cantidad de Cd fitoextraído. Las principales conclusiones de esta tesis son: La selección de microorganismos adaptados a las condiciones del suelo y capaces de colonizar la rizosfera son pre-requisitos importantes para desarrollar técnicas de inoculación exitosas. El estudio de la rizosfera de las hiperacumuladoras permitió el aislamiento de bacterias con potencial en técnicas de re-inoculación. Un pequeño número de aislados se encontraron exclusivamente asociados a las subspecies hiperacumuladoras de Alyssum y podrían ser candidatos interesantes para estudios adicionales relacionados con la aplicación de especies hiperacumuladoras de Ni en fitominería. El estudio de los mecanismos que controlan la disponibilidad de Ni en la rizosfera de las hiperacumuladoras es de vital importancia para el desarrollo de técnicas agronómicas enfocadas a la mejora de la fitoextracción (o fitominería) xxi Resumen mediante el incremento de la disponibilidad de Ni. La hiperacumuladora A. pintodasilvae mostró una mayor capacidad de influenciar las características físicoquímicas del suelo pero, contrariamente a lo esperado, en su rizosfera se apreció una menor capacidad de reposición de metal que en el caso de la exclusora H. lanatus. Los resultados de este estudio sugieren que la mayor disponibilidad de Ni en la rizosfera de la hiperacumuladora está relacionada con una mayor disolución mineral debido a la búsqueda activa de nutrientes por la planta. El carbono orgánico disuelto podría ser un parámetro determinante de la disponibilidad de metal en el suelo por lo que se necesitan más estudios para caracterizar su papel en la disponibilidad de Ni. Los estudios realizados a escala de maceta en invernadero demuestran que el uso de técnicas agronómicas como la fertilización inorgánica o el uso de enmiendas orgánicas pueden mejorar las condiciones fisicoquímicas del suelo para el desarrollo vegetal en técnicas de fitoextracción. Además, el uso de subproductos de la industria como enmiendas en el suelo proporciona una forma económica de reciclaje de estos residuos. Los estudios de bioaumento, también realizados a escala de maceta en invernadero, demostraron que en algunos casos los inóculos bacterianos pueden mejorar el crecimiento vegetal, modificar la disponibilidad del metal en el suelo y aumentar la acumulación de metal por las plantas. Sin embargo, estos efectos fueron altamente específicos dependiendo del tipo de suelo y/o especie vegetal. Además, en alguna de estas asociaciones planta-bacteria, se observaron efectos sinérgicos en el crecimiento vegetal cuando se combinaron enmiendas de suelo (compost) con la inoculación bacteriana. El método de inoculación mostró gran influencia en la respuesta de la planta a la inoculación. Los experimentos realizados permiten concluir que, tanto los modos de inoculación (semilla como en suelo) como la densidad celular del inóculo, son factores importantes a tener en cuenta a la hora de diseñar estrategias de inoculación bacteriana, y que es necesario lograr una optimización de estas estrategias antes de su aplicación a gran escala El estudio de campo realizado en una antigua mina de Cu (Touro) demostró que la fisoestabilización asistida puede ser una técnica adecuada para disminuir los riesgos medioambientales asociados a escombreras de minas altamente contaminadas. En este estudio, el uso de compost de baja calidad (basados en residuos orgánicos y/o industriales) permitió el establecimiento y crecimiento en suelos degradados, tanto de especies de alta biomasa como de gramínea. Los xxii resultados del ensayo mostraron también que enmiendas orgánicas en combinación con crecimiento de especies exclusoras pueden ser útiles en la reducción de la disponibilidad de Cu , la mejora la fertilidad y de la actividad biológica del suelo. El estudio de campo llevado a cabo en una antigua mina de Pb/Zn (Rubiais) en el que se evaluó la eficacia de fitoextracción de Salix spp. y de la hiperacumuladora N. caerulescens, demostró que, tanto el régimen de fertilización (NPK o enmiendas orgánicas), como el patrón de cultivo (monocultivo o intercalado con legumbres), pueden modificar la fertilidad del suelo, la actividad microbiana, el crecimiento vegetal, el estado nutricional y la acumulación de metales. 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Environ Sci Technol 49:4773-4780 Vance CP, Uhde-Stone C, Allan DL (2003) Phosphorus acquisition and use: critical adaptations by plants for securing a nonrenewable resource. New Phytol 157:423 -447 Vangronsveld J, Colpaert JV, Van Tichelen KK (1996) Reclamation of a bare industrial area contaminated by non-ferrous metals: Physico-chemical and biological evaluation of the durability of soil treatment and revegetation. Environ Pollut 94:131-140 Vangronsveld J, Herzig R, Weyens N, Boulet J, Adriaensen K, Ruttens A, Thewys T, Vassilev A, Meers E, Nehnevajova E, van der Lelie D, Mench M (2009) Phytoremediation of contaminated soils and groundwater: lessons from the field. Environ Sci Pollut Res Int 16:765-794 Wenzel WW, Bunkowski M, Puschenreiter M, Horak O (2003) Rhizosphere characteristics of indigenously growing nickel hyperaccumulator and excluder plants on serpentine soil. Environ Pollut 123:131-138 Chapter 1 4 Natural concentrations or background levels of TE in soils are strongly dependent on the soil parent material and acting weathering processes (Adriano, 5RVV$QLQFUHDVLQJO\LQGXVWULDOLVHGVRFLHW\KDVKRZHYHUOHGWRWKH widespread introduction of trace metals and metalloids into our environment, causing acute and diffuse contamination of soil and waters. Natural metal(loid)-enriched soils The most important types of natural metalliferous soils include (i) serpentine soils, which are derived from Feand magnesium (Mg)-rich ultramafic rocks and are significantly enriched in metals such as, Ni, chromium (Cr) and cobalt (Co) (ii) calamine soils, which are derived from non-sulphide Zn ores in carbonate rocks and are enriched in Pb, Zn and usually contain elevated concentrations of Cd, (iii) Cuand Co-containing soils, which are derived from argillites and dolomites, and (iv) soils derived from various selenium (Se)-rich rock types, such DVEODFNVKDOHV5HHYHV:LQNHO et al., 2015). Serpentine soils are formed through the weathering of ultramafic rocks which are comprised of at least 70% ferromagnesian (or mafic) minerals (particularly within the olivine and pyroxene groups) and less than 45% of silica (SiO2) (Kruckeberg, 2002). These soils are ubiquitous, but patchily distributed around the world (occupying approximately 1% of the earth’s crust (Proctor, 1999) and although some variations can occur between sites, serpentine soils are typically characterised by a low calcium (Ca):Mg ratio and elevated concentrations of TE such as Ni, Cr and Co (Table 1.1). Serpentine soils are also often deficient in essential nutrients such as nitrogen (N), potassium (K) or phosphorus 3*RUGRQDQG/LSPDQ9ODPLVDQG-HQQ\:DONHU3URFWRU DQG:RRGHOO%URRNV6LQFHPDQ\VHUSHQWLQHRXWFURSVDUHVWHHSDQG rocky their associated soils are often skeletal, with a low organic matter (OM) conWHQWDQGZDWHUKROGLQJFDSDFLW\:+&%URRNV0DFtDVDQG&DOYRGH$QWD 5REHUWVDQG3URFWRU$VDZKROHWKHSDUWLFXODUSK\VLFDOFKHPLFDO and biological properties of these soils are commonly referred to as the serpentine syndrome and present an inhospitable environment for most plant species. As a result, serpentine soils often support plant communities with high rates of endemism and which have evolved both morphological and physiological adaptaWLRQVGLIIHUHQWLDWLQJWKHPIURPWKHIORUDRIQHLJKERXULQJDUHDV:KLWWDNHU 3URFWRU-HQQ\:LWKLQWKHVHSODQWFRPPXQLWLHVDUDUHJURXSRISODQWV 5 Introduction can be found: the so-called metal-hyperaccumulators which will be discussed in the following section. Metal-hyperaccumulating plants Metalliferous plants can be classified into three groups according to their response (in terms of uptake and accumulation) to elevated TE concentrations in WKHVRLOH[FOXGHUVLQGLFDWRUVDQGK\SHUDFFXPXODWRUV%DNHU%DNHUDQG Brooks, 1989) (Figure 1.1). Excluder plants are able to restrict the uptake of metal (loid)s into their roots and/or their translocation to the shoot, thereby maintaining a low level of metal(loid)s in their aerial tissues over a wide range of soil metal(loid) FRQFHQWUDWLRQV LQGLFDWRU SODQWV WDNH XS PHWDOORLGV RYHU D ZLGHU UDQJH WKDQ ‘normal’ plants and the concentrations in their shoots reflect those of the soil, until SK\WRWR[LFLW\SUHYHQWVIXUWKHUJURZWKDQGFDXVHVGHDWKRIWKHSODQWDQGDFFXPX lator plants are able to actively take up and concentrate metal(loid)s in their aboveground tissues without showing symptoms of toxicity. Within the accumulators are the denominated hyperaccumulators which have a highly specialised physiology permitting them to accumulate extreme amounts of trace metals in their shoots when growing in metal(loid)-enriched substrates (Brooks et al., 1977). These metal(loid)s reach concentrations that would be toxic to most other plant species. The first plant species documented to accumulate such extreme levels of metal(loid)s in their leaves, already at the end of the 19th century, were Thlaspi caerulescens (now renamed as Noccaea caerulescens 0H\HU$O-Shehbaz, 2014)) and Viola calaminaria (Baumann, 1885). Later, in the 1940s, it was reported that Alyssum bertolonii, growing in the Tuscany region of Italy, contained DQH[WUDRUGLQDULO\KLJK1LFRQWHQWRIDERXWௗPJௗNJ−1 in its shoot dry matter Table 1.1. Mean element concentrations (μg g-1) in rocks, soils and vegetation Co Cr Cu Pb Mn Ni Zn Earth´s crust 25 100 55 13 950 75 70 Granite 3 20 13 48 165 1 45 Basalt 47 114 110 8 1280 76 86 Ultramafic rocks 150 1600 10 1 1620 2000 50 Soil (non-ultramafic) 10 60 20 10 850 40 50 Soils (ultramafic) 250 2500 20 10 1000 2500 40 Vegetation (non-ultramafic) 1 1 10 10 80 2 100 Vegetation (ultramafic) * 10 10 10 10 100 80 100  *excluding hyperaccumulators Source: Brooks, (1987) and Baker and Brooks, (1989) Chapter 1 6 (Minguzzi and Vergnano, 1948). However, the specific use of the term “hyperaccumulator” was introduced by Brooks et al. (1977) to describe plants containing a Ni concentration of greater than 1000 mg kg-1 in dry material. Later on the use of the term was widened to include other elements: the same criteria as Ni was used for Cu (Malaisse et al., 1978), Co (Brooks et al., 1980), Pb (Reeves and Brooks, 1983) and As (Reeves, 2006). While a threshold of 10000 mg kg-1 GU\ZHLJKW':ZDVVWLSXODWHGIRU=QDQG0Q5HHYHVDQG%URRNV Baker and Brooks, 1989) and of 100 mg kg-1 (0.01% DW) for Se and Cd (Reeves, 2006). Recently these criteria have been considered somewhat conservative and a lower threshold has been proposed for elements such as Co, Cu and Cr (300 mg kg-1':RUIRU=QPJNJ-1':.UlPHUYDQGHU(QW et al., 2013). Although still a topic of much debate, the hyperaccumulation trait is thought to have evolved as a means of reducing competition from other plant species (‘elemental allelopathy’), or as a defence mechanism against herbivores and SDWKRJHQLF PLFURRUJDQLVPV %R\G DQG 0DUWHQV  %R\G DQG -DIIUp  Davis et al.%R\G7KHUHDUHDSSUR[LPDWHO\WD[DRISODQWVWKDW are known to hyperaccumulate one or more metals or metalloids, and around 450 of these accumulate Ni (Hunt et al.3ROODUG et al.YDQGHU(QW et al., 2015). Within the Ni hyperaccumulators the most represented plant families include the Brassicaceae, Euphorbiaceae, Asteraceae, Flacourtiaceae, Buxaceae and Figure 1.1. Typical plant responses to increasing soil metalloid) concentrations (Baker, 1981). 7 Introduction Rubiaceae, although Ni hyperaccumulators are also present in another 30 families distributed throughout the plant kingdom, indicating that the hyperaccumulation WUDLWPXVWKDYHHYROYHGLQGHSHQGHQWO\VHYHUDOWLPHV5HHYHV3ROODUG et al., 2014). The Brassicaceae contains 25% of known Ni hyperaccumulators and the largest single clade is found in the section Odontarrhena of the genus Alyssum, which contains 48 Ni-hyperaccumulating taxa whose distribution is strongly correlated with the occurrence of ultramafic (serpentine) outcrops around the MediterUDQHDQEDVLQ%URRNV.UlPHU3ROODUG et al., 2014). In fact, the vast majority of Ni-hyperaccumulators appear to be serpentine-endemic species (Reeves and Adigüzel, 2008). The Iberian Peninsula hosts two Nihyperaccumulating subspecies of Alyssum serpyllifolium Desf. (Brassicaceae) which are both serpentine-endemic: Alyssum serpyllifolium ssp. lusitanicum Dudley and P. Silva (frequently referred as A. pintodasilvae) growing at the serpentine outcrops of Barazón (NW Spain) and Trás-os-Montes (NE Portugal) 3LQWRGD6LOYD)UHLWDV et al.'tH]-Lázaro et al., 2006) and Alyssum serpyllifolium ssp. malacitanum Rivas Goday (commonly referred as A. malacitanum) (López-González 1975) growing at the serpentine outcrops of Sierra Bermeja and Sierra de Aguas (S Spain). The subspecies A. serpyllifolium Desf. ssp. serpyllifolium (non-hyperaccumulator) is also found growing in limestonederived soils in NW Spain and S Spain. The list of Zn hyperaccumulators is much shorter than that of Ni hyperaccumulators, with c. 10-20 identified species (Baker et al.%URRNV7KH\ are mainly found growing on calamine soils, although for some species such as N. caerulescens and Arabidopsis halleri, non-metallicolous accessions have DOVR EHHQ UHSRUWHG 0HHUWV DQG 9DQ ,VDFNHU  %HUW et al., 2000). Zn hyperaccumulation appears to be concentrated in the genera Noccaea and Arabidopsis (both within the Brassicaceae family), while just a few species in other families have been found (including members of the Sedum genus &UDVVXODFHDHLQ$VLD%DNHUDQG%URRNV'HQJ et al.*XSWD et al.0DUWRV et al., 2016). Hyperaccumulation of Cd has been reported for the Zn hyperaccumulators N. caerulescens, A. halleri,S. plumbizincicola and S. alfredii (Brown et al..SSHU et al./RPEL et al.'HQJ et al., 2016) and the Ni hyperaccumulator Noccaea goesingensis (Lombi et al., 2000). Some plant species native to seleniferous soils accumulate Se to levels typically 100-fold higher than other vegetation: upwards of 1000 mg kg−1 (DW) and up to 15000 mg kg−1':WKHVHDUHFRQVLGHUHG6HK\SHUDFFXPXODWRUV%HDWK et Chapter 1 8 al., 1939). Selenium hyperaccumulation has been reported in ~30 species in the families Brassicaceae (Stanleya), Fabaceae (Astragalus) and Asteraceae (Xylorhiza,Oonopsis,Symphyotrichum5RVHQIHOGDQG%HDWK&DSSDDQG Pilon-Smits, 2014). Trace metal contamination in soils In addition to natural geogenic sources, elevated concentrations of trace metals may be present in the soil due to anthropogenic activities (Table 1.2). Soil contamination is defined by the European Commission (EC) as “the occurrence of pollutants in soil above a certain level causing a deterioration or loss of one or more soil functions” or “the presence of man-made chemicals or other alteration in the natural soil environment”1. Trace metals are one of the most frequent soil contaminants (24%) present at polluted sites across Europe (Panagos et al., 2013) and excess TE concentrations has been identified by the EC as one of the eight major threats to European soils, and thus food safety, animal and human health, and ecosystem services (COM, 2002). Soil is an extremely complex medium which performs vital functions such as biomass production, storage, filtration and transformation of substances. Soil has a role as a habitat and gene pool, serves as a platform for human activities, landscape and heritage and acts as a provider of raw materials. It contains around twice the amount of carbon in the atmosphere and three times the amount found in vegetation (Lorenz, 2013). As its formation is an extremely slow process and supports many vital functions, soil can be considered as an essential and nonrenewable resource2. Soil generates numerous ecosystem services which are classified as (i) supporting (soil formation, nutrient cycling, habitat), (ii) regulating UHJXODWLRQRIHOHPHQWDOF\FOHV&VHTXHVWUDWLRQ ZDWHUSXULILFDWLRQ DQGVWRUDJH adsorption and transformation of pollutants), (iii) provisioning (raw materials and ELRPDVVSK\VLFDOVWDELOLW\DQGVXSSRUWRISODQWVDQGLYFXOWXUDOKHULWDJHVLWHV archaeological archive) (Lorenz, 2013). Contamination leads to a decline in soil quality and biodiversity, which in turn reduces the soil’s capacity to perform ecosystem functions and provide these essential ecosystem services. The Thematic Strategy for Soil Protections of the European Union (EU) highlights soil contamination amongst the main soil degrading processes together ϭ http://eusoils.jrc.ec.europa.eu/library/themes/contamination/ Ϯ http://ec.europa.eu/environment/soil/index_en.htm 9 Introduction with erosion, losses in OM, compaction, salinisation, landslides, and soil sealing. The large volumes of industrial, municipal and agricultural wastes and widespread use of chemicals during past decades have left numerous contaminated sites across Europe. Local sources of soil contamination are mainly attributed to an inadequate or unauthorised waste disposal, or to the unsafe handling of dangerous substances within industrial or commercial processes and associated accidents (Panagos et al., 2013). The European Environmental Agency (EEA) estimated 2.5 million sites as potentially contaminated in EU countries and 342000 as contaminated sites, of which approximately 15% have already been remediated3. In general, production sectors (60%) contribute more to local soil contamination than service sectors (32%). The main anthropogenic sources of trace metals are associated with present 3 EEA, http://www.eea.europa.eu/data-and-maps/indicators/progress-in-management-ofcontaminated-sites-3/assessment Table 1.2. Main sources of metal(loid) pollution and their most common uses in industry (adapted from Adriano, 2001). Sources of pollution Uses in the industry Ni Mining and smelting, biosolid applications, fertilisers, coal Electroplating, alloy production and fabrication, Ni-Cd batteries and electronic components, and the preparation of catalysts for hydrogenation of fats and methanation. Stainless steel production, magnetic components, resistance to corrosion, heating elements Cr Metallurgical (ferroalloys and nonferrous allows), refractory and chemical industries. Electric furnaces, refractory brik, steel production, coal combustion, fertilisers, waste disposal. Manufacture of stainless steel, refractory industry, chemical industry (leather tanning, catalysts, pigments, textiles, wood preservatives and toner) Co High-grade steels, alloys, superalloys and magnetic alloys, drying agent, pigments glass decoloriser and catalysts in the petroleum industry Cu Electrical industry, metallurgical processing, iron and steel production and coal combustion, manure and sewage sludge addition Wire, fertilisers, bartericides and fungicides and in water putification, feed additive (antibiotics, drugs, chemical compounds, growth promoter agent for disease control) Cd Fertilisation, sewage sludge, incineration, fossil fuel, metallurgical, mining and smelter Alloys, electroplating, pigments, stabilisers for polyvinyl plastics, batteries Zn Automobile industry, fertilisation, sewage sludge, mining and smelting Protective coating to prevent corrosion and alloys such as building, transportation and appliance industries, water delivery industry. Manufacture, ingredient of several household items, fertilisers, textiles, adhesives or wood preservatives Pb Mining, smelting, refining, manufacturing, and recycling and disposal of Pbcontaining products Large rechargeable batteries, pigments, rolled and extruded products, cable sheathing, alloys, shot and ammunition and gasoline additives Chapter 1 10 and former mining activities, foundries, smelters (point sources) and diffuse sources such as piping, combustion of by-products, and traffic related to industrial and human activities (Panagos et al., 2013). In Spain, 71202 sites have been identified as potentially contaminated while 285 have been identified as contaminated (van Liedekerke et al., 2014). Most European countries have national or regional legislation to regulate local soil contamination but no legal framework has yet been established at the level of the EU. In general, legislation aims at preventing new contamination and setting targets for the remediation of sites where environmental standards have already been exceeded. At a Spanish level, soil contamination was regulated until 2011 by the Law 10/1998 of Residues and the Royal Decree 9/2005 of the 14th of January (BOE no. 15, 18 January 2005). Nowadays, soil contamination is regulated by the Law 5/2013 of Contaminated Soils of the 12th June (BOE no. 140, 12 June 2013) of Contaminated Soils and the Royal Decree 9/2005, both within the II Plan Nacional de Suelos Contaminados (2007-2015). These provide a list of potentially soil contaminating activities and the standard criteria on the basis of which a soil is declared as contaminated together with the definition of Generic Values of Reference (NGRs according to the Spanish spelling). NGR is the concentration of a contaminant in soils which does not entail a higher risk than that which is considered as the acceptable maximum for human health and ecosystems. The Royal Decree includes the NGRs for organic contaminants but not for TE and this responsibility is transferred to the regional governments, since metal(loid)s occur naturally in the soils and their background level and toxicity is strongly dependent of the geological substrate. In Galicia, the NGR values for TE are defined in the Soil Contamination Decree 60/2009 of the 6th of February (DOG no. 74, 24 March 2007). In terms of the addition of metal(loid)s to soils through soil amendments and/or fertilisers, the maximum permitted concentrations of TEs in sewage sludge and composts destined for agriculture use are regulated by the Royal Decree 1310/1990 (Ministry of Agriculture 1990) and by the Spanish Law on fertilisers, Royal Decree 824/2005 (Ministry of Agriculture 2005) respectively. However, the use of these waste products or residues as amendments in marginal or degraded lands (such as mine-soils) is not contemplated in current legislation. Mining activities are amongst the most polluting activities in the EU. The processing of the metallic ores involves several chemical and physical separation steps, such as the crushing and grinding of rocks, and the separation of waste from YDOXDEOH SURGXFWV 'XGND DQG $GULDQR  +RVNLQ et al., 2000). The ore 11 Introduction UHSUHVHQWVRQO\DVPDOOIUDFWLRQRIWKHWRWDOYROXPHRIPLQHGPDWHULDODVDUHVXOW the beneficiation processes generate large volumes of residues. These operations are restricted to relatively small areas, but tailings and rock deposits can cover extensive areas and are the main source of metal(loid) contamination (Salomons, 1995). Although mines are classified on the basis of their predominant product, they may produce large quantities of other toxic elements as co-products. As a result, metal(loid) ore processing usually leads to the multi-elemental contamination of the environment (Dudka and Adriano, 1997). In mine tailings, the main limitations for plant colonisation and growth are the lack of structure, water retenWLRQFDSDFLW\20DQGHVVHQWLDOQXWULHQWV:RQJ(UQVW,QDGGLWLRQ to the contamination of soils, mining areas are also sources of water and air pollution, arising from surface water runoff, acid mine drainage (which refers to the acidic waters produced by the dissolution of exposed sulphide minerals containing metal(loid)s in the presence of oxygen and water) and dust particles produced by grinding operations or wind erosion from bare soils (Hoskin et al., 2000). 1.2. Soil trace element availability and plant uptake and accumulation Metal(loid) (bio)availability can be defined as the metal(loid) fraction that can interact with a biological target (Geebelen et al., 2003). Plants acquire nutrients and trace metals from the soil which is in close contact with the roots (defined in 1904 by Hiltner as the rhizosphere). Metal(loid) availability to plants is governed by a pseudo-equilibrium between the elements in the soil solution and in the solid phase (Kidd et al., 2009) and is therefore determined not only by the concentration in the soil solution (from which plants mainly absorb soil elements) but also by the capacity of the solid phase to resupply the element (Bruemmer et al., 1986). In the soil solution elements are present as free uncomplexed ions, ion pairs, ions complexed with organic anions, and ions complexed with organic macromolecules and inorganic colloids. The most important metal(loid) pools in the solid phase include the exchange complex, metal(loid)s complexed by OM, sorbed onto or occluded within oxides and clay minerals, co-precipitated with secondary pedogenic minerals (e.g. Al, Fe, Mn oxides, carbonates and phosphates, sulphides) or as part of the crystal lattices of primary minerals (Adriano, 2001). When the rate of metal(loid) uptake by the plant exceeds its diffusional transport through the pore water its concentration in the soil solution is depleted and this induces resupply from the labile pools. Resupply occurs through the dissociation of complexes present in the solution phase but also via the release of labile Chapter 1 12 elements associated with the solid phase (e.g. on the exchange complex, complexed by OM) and is determined by the concentration of the element in the labile pool, its diffusional supply and the rate at which it is released from solid phase to solution (Lehto et al., 2006). Adsorbed metal(loid)s are exchangeable with the soil solution while the occluded fraction is (at least temporarily) unavailable for exchange with the soil solution. Finally, the fraction of trace metals which is structurally incorporated into soils minerals is not likely to become plant-available (Adriano, 2001). Plants mobilise soil nutrients and/or trace metals by actively changing the soil physicochemical properties in the rhizosphere. Plants differ in their capacity to convert non-available forms of (micro)nutrients to available forms according to differences in their root surface area (Sadana et al., 2002), in the composition and DPRXQWRIURRWH[XGDWHV5HQJHO-RQHV et al., 2004), or in their associated rhizospheric microbiota (Marschner et al., 2004). Between 10 to 20% of the photosynthetic carbon that roots receive is released as rhizodeposits in the rhizosphere (Singer et al., 2003). These exudates are believed to be implicated in the mobilisation of soil elements through the weathering of minerals (Hinsinger et al. -RQHVDQG(GZDUGV.LGG et al.%DUFHOyDQG3RVFKHQULHGHU Moreover, root-induced reactions such as chelation, precipitation or changes in pH and redox conditions also influence soil element availability, as well as the activity of plant-associated rhizosphere microorganisms (Uren and Reisenauer, 1988: Tao et al., 2004). Interactions between plant exudates and N, P or Fe nutrition are well known +LQVLQJHU-RQHV et al.+LQVLQJHU et al.'RWDQL\DDQG0HHQD 2015). Organic anions (such as citrate, oxalate or malate) are released into the rhizosphere in response to various nutritional stresses including P, Fe and micronutriHQWGHILFLHQF\EXWDOVRXQGHU$OWR[LFLW\VHHUHYLHZVE\+RFNLQJ5\DQ et al.1HXPDQQDQG5|PKHOG2UJDQLFDQLRQVDUHFRPPRQO\UHOHDVHG from roots in association with protons which results in an acidification of the rhizosphere (Dinkelaker et al.+RIIODQG et al.1HXPDQQ et al., 2002). Both, the decrease in pH and the complexing capacity of organic anions can also directly facilitate the mobilisation of sorbed P (Dinkelaker et al.+RIIODQG et al.6WU|P et al., 2005). Under Fe deficiency, Poaceae are known to secrete phytosiderophores, Fe(III)-solubilising molecules which are able to complex nonsoluble ferrous iron and make it available for plant uptake (Römheld and 0DUVFKQHU.RED\DVKLDQG1LVKL]DZD2WKHUPROHFXOHVVXFKDVIOD 13 Introduction vonoids, act as signal compounds in the establishment of the N2-fixing symbiosis with rhizobia under deficiency of external N supply (Cesco et al., 2010). The processes determining nutrient and metal(loid) speciation, distribution and availability are controlled by the combination of soil properties, plant characteristics and interactions with associated plant microorganisms (Jones et al., 2004). Metal(loid) availability in the rhizosphere of hyperaccumulators Metal(loid) hyperaccumulating plant species have an extraordinary capacity for trace metal accumulation, and furthering our understanding of the mechanisms controlling metal(loid) availability in the rhizosphere and their subsequent uptake and accumulation in shoots could lead to the development and improvement of remediation methods for metal(loid) contaminated soils using these plants Root system architecture plays an important role in nutrient uptake, since it determines the rhizosphere volume and plant access to soil elements (Comerford, 2005). Metal(loid) hyperaccumulators have generally been described with shallow root systems (< 0.5m) but they present a high proportion of fine roots which increases the surface area in contact with soil solution and greatly contributes to TE accumulation (Keller et al.  +LPPHOEDXHU et al., 2005). Direct root growth towards metal(loid)-rich patches in soil (chemotropism) has been reported in some hyperaccumulating species or ecotypes, for example, root foraging towards Zn and Cd metal(loid)-rich patches has been documented in N. caerulescens (Schwartz et al., 2003). However, non-hyperaccumulator plants are also known to alter their root growth to compensate for spatial variability in soil nutrients (which could be co-localised with trace metal(loid)-ULFK SDWFKHV -DFNVRQ DQG &DOGZHOO  5RELQVRQ%ORRP et al.'RXVVDQ et al., 2003). A higher metal(loid) accumulation in hyperaccumulators has been related to a higher expression of metal(loid) transporters at the root surface of these plants compared to non-hyperaccumulators, or to the presence of transporters with different kinetic properties (Pence et al.  /L et al., 2005). Specific transporters for essential elements are well documented (Grotz et al.  &OHPHQV3LWWPDQ%XUNKHDG et al.DQGQRQ-essential elements have been shown to enter roots through the same transporters by replacing these essential elements. For instance, As, Cd and Se can respectively replace P, Cu/Fe/ Mn/Zn or S (Cataldo et al.0DUVFKQHUGXHWRWKHLUVLPLODUYDOHQFH states and ionic diameter. In addition to membrane transporters, hyperaccumulators can promote metal(loid) accumulation through and increased release of metal Chapter 1 20 to contribute to the flux of the device, will be measured (Zhang and Davison, 2000). Figure 1.4 illustrates how the extent of depletion in the soil solution depends on the rate of release of the element from the solid phase. When the rate of release from the solid phase is fast, this supply effectively buffers the element concentration in soil solution and very little depletion occurs (Figure 1.4b). A small depletion in the solution would even occur if elements in the solid phase adjacent to the device are consumed. When the rate of release from the solid phase is very slow, there is little supply from this source, and element supply will depend on the diffusion of labile species in solution. In this case, the elements in solution adjacent to the DGT device are quickly depleted, causing a fast decline in the fluxes to the DGT device (Figure 1.4a). The intermediate case occurs when elements are released from the solid phase at a high rate, but insufficient to fully sustain the concentration in solution adjacent to the device (Figure 1.4c) (Zhang and Davison, 2006). The DGT technique can be used to determine the effective concentration (C E ) of elements in soil solution, and C E is based on both the soil solution concentration and its supply from the solid soil phase (Zhang et al., 2001). Plant element concentrations have been found to be linearly related and highly correlated with C E across a wide range of soil types, while nonlinear and more scattered relationships are obtained with free metal(loid) activity, total soil solution concentrations RUFKHPLFDOH[WUDFWLRQVHJ('7$'DYLVRQDQG=KDQJ1RZDFN et al., 7DQG\ et al.=KDQJ et al.=KDQJ et al., 2001). These results demonstrate that the kinetically labile solid phase pool of elements plays an important role in plant uptake and is included in the DGT measurement (Zhang et al., 2001). DIFS (DGT induced fluxes in sediments and soils) DGT measurements can be incorporated into numerical models to estimate the parameters quantifying the resupply of metal(loid)s from the solid phase to the soil solution, such as the equilibrium ratio between sorbed and dissolved concentrations (K D ) or the rate of release of sorbed fractions (T C ) (Ernstberger et al., (UQVWEHUJHU et al.+DUSHU et al., 1998). Harper et al. (1998) developed a one dimension numerical model, 1D-DIFS (DGT Induced Fluxes in Sediments and Soils), to quantify the processes influencing elemental flux into the DGT sink. The DIFS model describes the interaction of DGT with soil and estimates the 21 Introduction proportion of elements in the solid phase (KD) which includes all species adsorbed WRWKHVROLGSKDVHWKDWFDQGHVRUEHGDQGPHDVXUHGE\'*7DVZHOODVWKHUHVSRQVH times with which the equilibrium between the solid and solution phase concentrations takes place (Tc) (Ernstberger et al.  (UQVWEHUJHU et al., 2005). An improved version of the DIFS model, working in two dimensions (2D DIFS), was developed by Sochaczewski et al. (2007) permitting a more accurate determination of the KD and TC parameters. The DIFS model requires the input of several parameters which influence the rate of diffusion of elements such as the diffusion coefficient of the element in water (Do), tortuosity (ڧ2), soil particle concentration (Pc), mass of accumulated elements in the resin gel (M), time of deployment (T) and the R value. The R value is calculated as the ratio of CDGT to the initial measured soil solution concentration (Csol) and is governed by the quantity of elements available for supply from the solid phase and rate of this resupply and is dependent on the deployment time and the extent of depletion (Lehto et al., 2008). For example, R values equal to 1 indicate a continuous resupply from the solid phase to the soil solution while Figure 1.4. Schematic representation of a cross section through a DGT device in contact with a soil or sediment. Pseudo steady-state concentration gradients are illustrated for three cases: (a) unsustained, (b) sustained and (c) the general or partially sustained case. Ca is the interfacial pore water concentration between the sediment and DGT device. Chapter 1 22 values lower than 0.1 indicate that resupply is low and there is limited mobilisation from the solid phase after depletion in the soil solution (Conesa et al., 2010). The DIFS model quantifies the dependence of the experimental R (obtained from different deployment times) on resupply of trace metals by quantifying the distribution coefficient between the solid and solution phases (KD) and the adsorption/ desorption kinetics (TC) (Ernstberger et al./HKWR et al., 2008). Three major sets of conditions that determine the types of soil response to local depletion have been described (Figure 1.4) (Harper et al.  Sochaczewski et al., 2007): (b) sustained case: while dissolved species diffuse towards the DGT sink from the sediment/soil, there is fast resupply from the solid phase that has sufficient capacity to maintain this supply over the time of interest (large TC, KDFORVHWR]HURFSDUWLDOFDVHUHVXSSO\IURPWKHVROLGSKDVHH[LVWV but it is insufficient to sustain the demand by the DGT device at its maximum value (large KD and medium TC, or medium KDDQG7FDGLIIXVLYHFDVHWKH capacity of the media (KD) is very small or the response time of the sediment is very large (TC) so that there is no effective resupply to the dissolved phase (small TC and large KD) (Sochaczewski et al., 2007). 1.3. Gentle remediation options targeting trace element-contaminated land Remediation of contaminated soils is mainly based on the so-called “traditional” techniques, such as excavation and off-site disposal, which account for about one third of implemented management measures (van Liedekerke et al., 2014). However, in the last three decades an increase in the use of Gentle soil Remediation Options (GRO) has been documented, which include in-situ contaminant stabilisation and plant based options (phytoremediation). Phytoremediaton techniques are considered to be cost-effective options and less invasive than conventional civil engineering techniques (Mench et al.9DQJURQVYHOG et al.  0HQFK et al.  *UHHQODQG 7KHVH green techniques have been developed to target both organic contaminants and TEs and include the following processes (Wenzel et al.  6DOW et al.  3LORQ-6PLWV  Chaney et al..LGG et al., 2009): -Phytostabilisation uses metal(loid)-excluding plants for the in-situ stabilisation of metal(loid)s in soils usually in combination with soil amendments (aided phytostabilisation) thus reducing their bioavailability and transfer to the food chain. Phytostabilisation techniques have been suggested to be the more reasonable option for contaminated areas covering large areas and with a high soil metal 23 Introduction (loid) availability (Schwitzguébel, 2014). This technology does not lead to the actual clean-up of the soil, but reduces the potential toxic effects of metal(loid)s in the environment. -In situ stabilisation and phytoexclusion aims to reduce pollutant bioavailability immobilising or binding them to the soil matrix through the incorporation into the soil of organic or inorganic compounds, singly or in combination, to prevent the excessive uptake of essential elements and non-essential contaminants into the food chain. Phytoexclusion is the implementation of a stable vegetation cover using excluder crop plants which do not accumulate contaminants in the harvestable plant biomass can be combined with in situ immobilisation. -Phytoextraction involves the cultivation of tolerant plants that concentrate soil contaminants in their aboveground tissues. At the end of the growth period, plant biomass is harvested and can be burned to produce metal(loid)-enriched ash or “bio-ore” (the process is then known as phytomining). The energy gained from biomass combustion can support the profitability of the technology. Bio-ores can also be used for metal(loid) recovery (smelting, pytometallurgy). When aided by use of soil amendments, the technique is termed aided phytoextraction. -Phyto(rhizo)volatilisation employs plants and associated rhizosphere microorganisms to transform pollutants into volatile compounds that are then released into the atmosphere. -Phyto(rhizo)degradation refers to the use of the metabolic capabilities of plants and rhizosphere microorganisms to uptake, store and/or degrade organic pollutants. Phytostabilisation and phytoextraction are the most common techniques applied to TE-contaminated sites, and both aim to decrease the labile metal(loid) pool in soils which interacts directly with living organisms. When the phytoextracted metal(loid) can be recovered and commercialised with an economical gain, the technique receives the name of phytomining (Li et al.D&KDQH\ et al., 7DQJ et al., 2012). A major advantage is their lower cost as compared to the other more conventional remediation techniques (especially when these involve ex situ treatment). Moreover, GRO can also provide a range of additional economic (e.g. biomass generation), social (e.g. leisure and recreation) and environmental (e.g. C sequestration, water filtration and drainage management, restoration of plant, microbial and animal communities) benefits (which are encompassed in the generic term “ecosystem services”). On the other hand, the major limiting factor is the time taken to reduce the “total” TE concentrations to below regulatory limits. Chapter 1 24 This has led to intensive discussions, in particular, about the viability of phytoextraction (Robinson et al., 2009). The factor determining the duration of phytoextraction is the mass of TEs removed by the crop per unit of time (years) compared to the mass of TE in the soil. A decline in plant uptake as soil TE concentrations decline leads to further increases in the time needed for clean-up (Robinson et al., 2009). Phytoextraction is only considered applicable in low to moderately contaminated soils, and its effectiveness can also be reduced by the limited rooting depth of plants and the presence of contaminant mixtures. The success of GRO is highly dependent on the solubility and speciation of TE in soil, as well as their tolerance and uptake by plants. Adequate plant selection for each technique will be vital. In phytoextraction-related technologies a high uptake of specific TE into the shoots is desirable, while successful phytostabilisation requires that TE be excluded from the aerial portions. Plant-microbial-soil interactions in the rhizosphere will affect TE mobility and availability, and these can be manipulated to optimise the processes operating in GRO. For example, the development of methods for modifying TE bioavailability could have a positive impact on phytoremediation efficiency. Increasing contaminant bioavailability would improve processes such as phytoextraction, whereas the use of amendments which reduce contaminant bioavailability could substantially improve phytostabilisation (Wenzel, 2009). During the last decade, agronomic and biotechnological methods have been developed to optimise the effectiveness and success of GRO. Some of these methods which are suitable for the remediation of trace metalcontaminated sites are described below. Finally, the number of bench-scale or greenhouse-based studies by far exceeds the number of field-scale evaluations of GRO and there is a clear need for field trials to evaluate their performance under realistic conditions (Mench et al.9DQJURQVYHOG et al.0HQFK et al., .LGG et al., 2015). 1. 4. Improving efficiency of Gentle Remediation Options Selection of adequate plant species for implementing gentle remediation options Metal(loid) toxicity can severely limit the performance and establishment of the plants, and the selection of TE-tolerant plant species is therefore vital for the successful implementation of GRO in metalliferous sites. However, plants must also tolerate numerous additional abiotic and biotic stress factors, such as water and nutrient deficiency, soil acidity or salinity, soil erosion or compaction, flooding, herbivory or pests. 25 Introduction Natural metal(loid)-enriched areas (such as ultramafic regions) or metal (loid)-contaminated sites (such as mine tailings and spoils) are important sources of metal(loid)-tolerant plant genotypes. Despite the unfavourable growth conditions at these sites, several metallophyte plants have evolved mechanisms to tolerate these conditions and are able to colonise these types of substrates (Batty,  0HQGH] DQG 0DLHU  :KLWLQJ et al., 2004). Within the same plant species various ecotypes, cultivars, varieties or clones can differ greatly in their response to the presence of contaminants (Vyslouzilova et al.0DUPLUROL et al.5XWWHQV et al., 2011). The appropriate plant species will depend on the remediation option to be used. For phytoextraction plants must be able to tolerate and accumulate high concentrations of TEs in their harvestable parts and have a reasonably high biomass production. One relevant option is to use TE-hyperaccumulators, and one argument in favour of their use is the possible recuperation of TE from metal(loid)-rich biomass. Due to the extreme ability of hyperaccumulators to accumulate and concentrate metal(loid)s in their aerial parts (harvestable biomass), these plants have been described as good candidates for phytoextraction techniques. At present, Ni and Co present the highest prices on the market (average in $USD kg -1 : 5.5 for Cu, 2.0 for Zn, 11.4 for Ni, 1.8 for Pb and 31.1 for Co at the London Metal Exchange /0(-XO\KRZHYHUGXHWRWKHKLJKQXPEHUDQGDPSOHGLVWULEXWLRQRI1Lhyperaccumulators, Ni seems to be the most appropriate element for phytomining (Chaney et al., 2014). The cultivation of hyperaccumulator plants as part of phytomining strategies on mineral wastes (metal(loid)-contaminated soils or residues) or natural metal(loid)-rich soils (sub-economic ores such as ultramafic soils) can generate Ni bio-ores (i.e. plant ashes) with up to 25% Ni, which is significantly higher than that present in current lateritic ores (< 1.5%) or normal Ni -ores (approx. 3%) where conventional mining processes are considered economiFDOO\XQYLDEOH&KDQH\&KDQH\ et al., 1998). Moreover, this bio-ore is free of Fe, Mn and Mg silicates present in the soil matrix which are known to increase the costs of metal(loid) extraction (Chaney et al..RSSROX et al.YDQ der Ent et al.=KDQJ et al., 2014). Phytomining of Ni has been successfully demonstrated in several studies carried out at a field scale (Figure 1.5). Li et al. (2003b) reported a biomass production of A. murale and A. corsicum in field trials of up to 20 t ha -1 . With a shoot Ni concentration of more than 20,000 mg kg -1 the total amount of phytoextracted Ni was 400 kg ha -1 , resulting in a net annual benefit of $1749 ha -1 (with a Ni price Chapter 1 26 of $7.33 kg-1 Ni and assuming costs of production of $1000 ha-1). Chaney et al., (2007) reported economic returns of $16000 ha−1 from phytomining processes using Alyssum species (at the time of this study the value of Ni was more than $40 kg-1). Successful phytomining trials have also been reported in Europe, for example, Bani et al. (2015) showed that native populations of A. murale in Albania could phytoextract an economically-viable crop of Ni after incorporating inorganic fertilisation. These authors achieved a biomass production of 9 t ha-1 with a Ni concentration of 11.5 g kg-1 which resulted in a total amount of phytoextracted Ni of 105 kg Ni ha-1 per year. The estimated costs of the phytomining cycle were around $1000 ha-1 yr-1 and, based on the price of Ni at the time of the study ($10 kg-1/RQGRQ0HWDO([FKDQJHRQ-XO\ WKHQHW UHWXUQZDVFDOFXODWHGDV 1055 ha-1 yr-1. These calculations were based on prices of gross Ni metal, however, other Ni products (such as Ni salts, Ni(NH4)2(SO4)2.6H20) with a high purity and higher value and which can be easily obtained from the biomass of A. murale, would substantially increase the profit of the crop (Barbaroux et al., 2012). Other Ni-hyperaccumulating genera such as Leptoplax or Bornmuellera have also been suggested as potential candidates for Ni recovery (Li et al.D Zhang et al., 2014). Studies evaluating the use of metal(loid) hyperaccumulators to phytoextract elements other than Ni, such as, Zn, Cd or As, from contaminated sites can be found in the literature. A field experiment evaluating the use of N. caerulescens (Viviez) to phytoextract Cd and Zn from an agricultural soil amended with TErich-sewage sludge addition found two successive crops simultaneously extracted about 9% of the total Cd and 7% of the total Zn (Schwartz et al., 2003). Similarly, the same hyperaccumulator extracted up to 5% of total Cd and 11% of total Zn from an industrial soil affected by a former coking plant, and the available Zn fraction (NH4NO3-extractable) was reduced by up to 70% (Schwartz et al., 2003). The As-hyperaccumulator Pteris vittata decreased total As concentrations in contaminated paddy soils by up to 11.4% after 9 months of growth, while As pore water concentrations decreased up to 77%. Shoot As concentrations in P. vitatta ranged from 45 to 206 mg kg-1, which are close to or above the threshold concentrations given for As hyperaccumulation (Ye et al., 2011). Recently, van der Ent et al. (2013) suggested the potential for phytomining of rare elements, such as gold (Au) or rare earth elements (REE) such as cerium (Ce) or lanthanum (La)). 27 Introduction Figure 1.5. Phytomining projects carried out at (a) Refinery field at Port Colborne, Ontatio; Alyssum corsicum (b-c) Prrenjas / Pojskë (Albania) Alyssum murale (d) harvesting (e) combustion to recover biomass energy and Ni-rich ash and (e) recovery of Ni from ash. Photos courtesy of Aida Bani and Rufus L. Chaney Chapter 1 28 However, to date only Ni phytomining has been proven to be economically feasible (Chaney et al., 2007). The price of metals, such as Zn, on the world market is at present too low to warrant Zn recovery from contaminated soil. On the other hand, Zn enriched biomass obtained during phytoextraction can be used to produce Lewis acid catalysts for organic chemistry, thus providing an economic incentive to the process. Nonetheless, the practical application of hyperaccumulators continues to be limited by the low biomass production of most of these species (e.g. N. caerulescens) and the high number of cropping cycles required for clean-up (if the objective is to reduce total TE concentrations in soils). Additional limiting factors include the absence of commercially available seeds, their sensibility to the presence of metal(loid)s other than the hyperaccumulated, lack of knowledge related to their cultivation, climate needs or competition with other metal(loid)-tolerant plants (Kidd et al., 2015). High-biomass crops and woody plants are recognised as viable alternatives to hyperaccumulators for phytoextraction of TEs (particularly Cd, Se and Zn) if they show moderate-high bioconcentration factors (BCF) and high shoot yields. Both high biomass annual plants and fast growing trees can be easily cultivated XVLQJZHOOHVWDEOLVKHGDJURQRPLFSUDFWLFHV*KRVKDQG6LQJK0HHUV et al., 6ROKL et al., 2005). Extensive work has been done on the trace metal(loid) uptake capacity of high biomass crop plants, such as Indian mustard (Brassica juncea), sunflower (Helianthus annuus), and maize (Zea mays) (Cui et al. Turgut et al.6]DEyDQG)RGRU+LJKELRPDVVWUHHVVXFKDVZLOORZV (Salix spp.) and poplars (Populus spp.) have great potential in furthering efforts to develop phytoextraction (Liphadzi et al.  9HUYDHNH et al., 2003). Fast growing trees are ideal plant species due to their extensive root systems, rapid growth, large biomass production and easy harvesting with subsequent re-sprouting (Peuke and Rennenberg, 2005). Zhuang et al. (2007) compared the phytoextraction efficiency of two hyperaccumulator species with six high-biomass plant (sub)species. Although metal(loid) accumulation was higher in the hyperaccumulators (Viola baoshanensis and Sedum alfredii), the high biomass production of the non-hyperaccumulator Rumex ciprus (20 t ha -1 ) compensated for its moderate metal(loid) uptake and metal(loid) removal from soils was similar for both plant types. In vitro breeding (cell and callus tissue culturing on metal(loid) spiked media) and chemical mutagenesis can significantly improve the metal(loid) tolerance and phytostabilisation or phytoextraction capacity of high yielding DQQXDOFURSVVXFKDVWREDFFRRUVXQIORZHU*XDGDJQLQL+HU]LJ et al. 29 Introduction Nehnevajova et al.  1HKQHYDMRYD et al., 2009). These non-genetically modified plants can be directly tested for their metal(loid) extraction potential under real field conditions without any legal restrictions (Herzig et al., 2014). Sunflower accumulates moderate amounts of metal(loid)s from contaminated soil and water, but their high biomass production makes them interesting for phytoextraction (Lombi et al.0DGHMyQ et al.1HKQHYDMRYD et al., 2005). Furthermore, a negligible metal(loid) concentration in sunflower seeds and oil limits the risk of food chain contamination and should allow the use of RLOIRUELRIXHOV.DVKHPDQG6LQJK0DGHMyQ et al., 2003). Herzig et al. (2014) estimated a remediation time of between 10 to 46 years for a smelteraffected site with high Zn contamination (16 mg kg NaNO 3 -extractable Zn) using rotations of improved cultivars of sunflower (accumulating up to 346 mg Zn kg -1 and biomass production of up to 26.8 t ha -1 ) and tobacco (accumulating up to 617 mg Zn kg -1 and biomass production of up to 37.5 t ha -1 ). Delplanque et al. (2013) calculated a remediation time 41 years to decrease the total soil Cd concentration from 2.39 mg kg -1 to 2 mg kg -1 (threshold concentration according to the French regulation) using Salix schwerinii x Salix viminalis “Tora” (a wellknown Cd accumulator) if both stems and leaves were harvested. The remediation time increased to up to 62 years if only stems were harvested. The long times needed for soil clean-up indicate that the time-scale for Cd phytoextraction could only be acceptable if the remediation process is combined with biomass production for bioenergy or feedstock. In the latter study, the authors also calculated the removal efficiency of Cd according to the extractable fraction, in this case, clean-up time ranged from 1 to 2 years. This process is known as bioavailable metal(loid) stripping and several authors have argued in favour of this rather than targeting reductions in the total metal(loid) contamination (Herzig et al., 2014). However, current legislation in most of European countries and initial regulations at a European level for the identification of pollutants thresholds are still based on total metal(loid) concentrations (Bachmann et al.(&)LJXUHVKRZVH[DPSOHVRIILHOGWULDOV implementing phytoextraction techniques in Europe. Phytostabilisation often combines re-vegetation with metal(loid)- excluding plants and the incorporation of inorganic and/or organic amendments to improve soil physical, chemical and biological properties (Vangronsveld et al.9DQJURQVYHOGDQG&XQQLQJKDP%DUUXWLD et al., 2011). Plants Chapter 1 36 ( Continued ) Site description Plant species Soil management Type of remediation technique Pollutants Experiment duration References Metal(loid)-contaminated wastes derived from pyrite ore roasting for sulphur extraction were discharged adjacent to a chemical plant at Torviscosa (Italy) P. alba, P. nigra, P. tremula and S. alba Fertilised with NPK and irrigation Phytostabilisation As, Co, Cu, Pb and Zn 6 months Vamerali et al. (2009) Diffusely contaminated region by smelter activities (Campine region in Flanders, Belgium) S. alba var. Alba (‘Belders’), S. x rubens var. basfordiana (‘Belgisch Rood’), S. viminalis (‘Christina’), S. triandra x viminalis (‘Inger’), S. viminalis (‘Jorr’), S. dasyclados (‘Loden’), S. schwerinii x viminalis (‘Tora’) and S. triandra (‘Zwarte Driebast’) Limed one month before planting and mechanical weeding Phytoextraction Pb, Cd, Zn 1 growing season Van Slycken et al. (2013) Dredged sediment disposal sites in Menen (Belgium) S. viminalis ‘Orm’ Phytoextraction Cd, Pb, Zn, Cu 1.5 year Vervaeke et al. (2003) Cd, Pb and Zn-contaminated agricultural soil (Podlesí, Czech Republic) ([S. schwerinii × viminalis] × viminalis); S. × smithiana; P. aximowiczii × nigra; P. nigra clone Wolterson Phytoextraction As, Cd, Pb and Zn. 4 years Zárubová et al. (2015)  37 Introduction Fig 1.7. Field–scale phytostabilisation trials in (a-b) the Piekary site in Poland where smelteraffected soils were amended with biosolids and lime and planted with a grass cover, (c-d) Biogeco site in France planted with ectomycorrhizal poplar and willows on the left and unmycorrhizal poplar and willows on the right and (e-f) Phytosed Ec site in France where soils were amended with Thomas Basic Slag and planted with Deschampsia cespitosa (vegetation cover) and Salix SRC Chapter 1 38 The practice of liming (chalk or limestone (CaCO3), quicklime (CaO), or hydrated lime (Ca(OH)2) to increase soil pH has been commonly used and associated effects on metal(loid) mobility are well known (reviewed by Vangronsveld et al., 2009). Liming can also be ineffective in metal(loid) immobilisation but this is usually metal(loid)-dependant. McLaughlin et al. (2000) reported an increase in Cd accumulation in potato tubers after lime addition. Addition of lime can also potentially mobilise metalloids such as As (Jones et al., 1997). The long-term effects and sustainability of liming are also questionable (Gray et al.5XWWHQV et al., 2010). For this reason, other alkaline amendments, and different combinations of amendments, have been evaluated. Amendments rich in metal(loid) oxides combined with compost, fertilisers, EHULQJLWHF\FORQLFDVKHVRUOLPHHQKDQFHGSODQWJURZWK%HVDQG0HQFK Vangronsveld et al., 2009). The combination of iron grit with lime and compost was more effective in reducing Cu concentrations in soil pore water than individual amendments (Bes and Mench, 2008). Application of cyclonic ashes (formerly known as beringite) which are rich in clay minerals induced a strong decrease in plant-available Pb, Zn and Cd concentrations, and restored vegetation, in the Zn-smelter affected area of Lommel (Belgium). Moreover, this effect was still observed after 12 years (Vangronsveld et al. D 9DQJURQVYHOG et al., 2000b). The application of red muds (an alkaline and Fe/Al oxide-rich by-product of the aluminium industry) led to a decrease of up to 63%, 425% and 50% in soil Zn, Cd and Ni extractability in a 15-month pot study. However, similar effects of these amendments were not observed when applied at a field scale (Friesl-Hanl et al., 2006). Many of these amendments are by-products of industrial activities and are therefore inexpensive and available in large amounts (Mench et al., 1998). Moreover, their use can provide an environmentally sustainable means for the recycling of such residues (instead of incineration). Organic matter addition can improve soil physical properties, water infiltration and WHC, provide microand macro-nutrients for plant growth and stimulate microbial activity. In addition, the formation of stable organo-metal(loid) complexes can reduce the availability of metal(loid)s to plants for uptake (Clemente and Bernal, 2006). Authors have recommended the use of organic amendments with a low mineralisation rate within a neutral pH range in order to avoid the release of metal(loid)s and reduce their uptake and bioaccumulation by plants (Blake and Goulding, 2002). Although the most widely described effect of OM additions on metal(loid) mobility is one of immobilisation, some studies have 39 Introduction also shown an enhanced metal(loid) solubilisation (Almås et al.&OHPHQWH DQG %HUQDO  7DQG\ et al., 2009). The effects of organic amendments on metal(loid) availability depend on the nature of the OM, and on the particular soil type and elements concerned (Clemente et al.5XWWHQV et al.D*RHFNH et al.  .XPSLHQH et al.  /DJRPDUVLQR et al., 2011). Other organic amendments, such as municipal wastewater, sewage sludge, landfill leachate or wood ash, have also been applied in GRO (Marchiol et al.$GOHU et al., 2008). A key element in the sustainability and (self)-maintenance of phytostabilisation techniques is the monitoring of amendment-induced effects and the need for new applications (especially in the case of liming or OM addition), since soil reacidification after liming or OM mineralisation may result in the re-mobilisation of sorbed metal(loid)s (Tisch et al., 2000). Nutrient deficiency is a major limiting factor for plant growth in TEcontaminated areas, and is especially limiting in environments such as mine tailings (Ye et al., 2002). Fertilisation of tailings soil can facilitate plant survival, alleviate growth inhibition and ensure a better plant development. Pioneer studies already demonstrated the importance of fertilisation during the re-vegetation of TE -contaminated mining areas in the USA (Pancholy et al.&XQGHOO Fertilisation regimes can be designed with the aim of improving plant growth and establishment but they can also be designed to increase plant uptake of TEs. Inorganic fertiliser application can affect TE bioavailability, and much research has been carried out towards a better management of fertilisers and reduction in food-chain transfer of TE in agriculture (McLaughlin et al.0HQFK McLaughlin et al.0F/DXJKOLQDQG6LQJK6LQJK et al., 2011 Chaney, 2012). The type and amount of fertiliser used and interactions between TE and major nutrients (N, P and S) and among TE themselves (e.g. Zn–Cd and Fe–Cd) are key-players in TE uptake by crops (Tiller et al.0HQFK3ODQW availability of Cd can be affected directly through the addition of Cd as a contaminant in P fertiliser, or indirectly through ion exchange reactions in the soil solution 0F/DXJKOLQDQG6LQJK:nQJVWUDQG et al., 2007). Furthermore, fertilisers can influence Cd speciation and complexation. Chloride can increase crop Cd uptake more than other soil factors due to the formation of chloride complexes (Smolders et al., 1998), and this is particularly pronounced in alkaline soils (Hattori et al., 2006). Ammonium fertilisers frequently cause higher Cd concentrations in crops than nitrate fertilisers as a result of the pH decrease caused by nitrification or plant uptake of NH4+ (Grant et al., 1999). The chemical form of Chapter 1 40 fertilisers can therefore be selected according to the GRO technique to be implemented: in the case of phytoextraction, ammonium-, sulfateor chloride-based fertilisers may be managed so as to enhance the phytoavailability and hence uptake of metallic cations, while in the case of GRO where TE stabilisation is the aim avoiding certain fertiliser types can minimise plant TE uptake. Ammoniumsulphate enhanced metal uptake and accumulation (Cd, Zn and Pb) by tobacco mutants in phytoextraction field trials, and this treatment was shown to be more efficient than the addition of chelating agents such as citric acid or nitrilotriacetic acid (NTA) (Herzig et al., 2005). The application of elemental S may increase the metal(loid) solubility for plant uptake (Wenger et al.:DQJ et al. Iqbal et al., 2012). Microbial oxidisation of elemental S leads to the production of sulphuric acid and consequent decrease of soil pH (depending on the soil pH buffering capacity). Sulphur oxidation along with microbial and root respiration might also lead to partly anoxic conditions in the rhizosphere, inducing reductive dissolution of manganese oxides, which further enhances the Zn and Cd solubility (Iqbal et al., 2012). However, a site-specific balance between increasing metal (loid) solubility and potential negative effects due to decreasing pH needs to be found. A decrease in soil pH to values below 4.5-5.0 would lead to irreversible soil degradation and massive leaching of macronutrients such as Ca, Mg and K and to an increase in potential Al phytotoxicity (Fässler et al., 2012). Fässler et al. (2012) increased soluble Cd and Zn after elemental S application by decreasing soil pH. Hoefer et al. (2015) reported an increase in the Cd and Zn availability of a calcareous soil after S0 addition and at the same time an increase in shoot metal concentration by willows was found. These authors attributed the effects to soil acidification after sulphur application but also to the co-solubilisation of metals presents in Mn-rich minerals. Improvements in the biomass production of hyperaccumulators using inorganic fertilisers have been also reported (Robinson et al., 6FKZDUW] et al.%DQL et al., 2007). Plant cropping patterns to improve gentle remediation options performance Crops patterns can be designed so as to improve plant biomass and nutrition, and/or enhance or mitigate TE availability and uptake and accumulation (Kidd et al., 2015). Most plant species selected as relevant candidates for GRO have been studied as monocultures. Monocultures can lead to a decline in biomass yields due to the depletion of nutrients, occurrence of diseases, pests, and weeds, DQGKDYHDQHJDWLYHHIIHFWRQVRLOIHUWLOLW\)DFNQDWKDQG/DOOMHH  /DVDW 41 Introduction 0HQFK et al., 2010). Reductions in biomass yield can reduce the plants’ phytoextraction capacity. While alternative cropping patterns, such as rotations or intercropping, can improve plant productivity and nutrition, enhance biodiversity and habitat, or aid pest control. For both improved sunflowers and tobacco mutant lines a crop rotation scheme is obligatory to avoid plant diseases over a longer period of time (Herzig et al., 2014). Crop rotation with non-hosts, such as rapeseed, corn, and winter fodder pea is suggested. Although a minimum rotation period of four years is recommended between successive sunflower crops, good results have been obtained with three year crop rotations of sunflower – tobacco – corn, or a two year rotation of sunflower – tobacco followed by winter fodder pea in Zn-contaminated agricultural soils which were polluted with deposits of a IRUPHU KRW GLS =Q IDFWRU\ %HWWZLHVHQ 6ZLW]HUODQG )LJXUH +HU]LJ et al., 1HKQHYDMRYD et al.+HU]LJ et al., 2014). Winter fodder pea (green fertiliser) can still grow during winter and avoids periods of bare soil, thus reducing the risks of transfer of the contaminated soil material by wind or water erosion processes and the leaching of TE’s during this time. After five years of the sunflower-tobacco rotation scheme, bioavailable (labile) Zn concentrations were reduced by up to 70% in top soils (Herzig et al., 2014). Sunflower-tobacco rotations and tobacco–sunflower–vetiver rotation were also successfully implemented in Cu-contaminated soils at a former wood preservation site (Kolbas, 2012). As discussed above, SRC of fast-growing tree species (such as Salix and Populus spp.) is a common technique for production of biomass for conversion to energy or as a feedstock for other chemical processes /DQGEHUJDQG*UHJHU5LGGHOOBlack et al.  .D\VHU et al.  +DPPHU et al., 2003) and has been successfully implemented within GRO on metal(loid) contaminated soils (Dimitriou et al., 2012). These woody species show the ability to re-sprout from the stumps after harvests which are performed at short time intervals (i.e. 2 – 6 years) (Dimitriou et al., 2012). Management practices for SRC e.g. weed control, planting, fertilisation, and harvesting, are in between agriculture and forestry and the equipment used resembles more that of common agriculture, although they are specially designed for SRC. Harvest of SRC is generally carried out in winter or early spring (Dimitriou et al., 2012). The use of polyclonal stands of Salix and Populus clones rather than monocultures has also been recommended when implementing SRC in phytostabilisation options as a means of reducing disease incidence, particularly from Melampsora rust 0F&UDFNHQ DQG 'DZVRQ  French et al., 2006). Chapter 1 42 Intercropping of plants is often used in agriculture and aims to stimulate interspecific below-ground interactions, which may result in improved nutrient availability and increased yield of crops (Wieshammer et al., 2007). Nutrient deficiency is a common characteristic of contaminated soils, but intercropping with leguminous plants can supply nutrients (through N2 fixation, transfer of fixed N and mobilisation of P due to rhizosphere acidification) to phytoremediating crops. Intercropping can also alter conditions in the shared rhizosphere and thereby affect the availability of selected TE to neighbouring plants (Tang et al., 2012). For phytoextraction purposes, combining hyperaccumulators (with relatively restricted root proliferation) with deep-rooting species (such as willows) could favour a more complete and deeper use of soil resources, or even target deeper contaminants (Keller and Hammer, 2005). Combining low-metal(loid) accumulating crops with hyperaccumulators or other appropriate metal(loid)-accumulating plants could permit simultaneous remediation with agricultural production on TE-contaminated soils (Tang et al., 2012). Co-cropping of the hyperaccumulator N. caerulescens with a non-accumulator Thlaspi arvense increased the growth of the nonaccumulator and reduced its Zn uptake, while at the same time Zn uptake was enhanced in the hyperaccumulator (Whiting et al., 2001b). The co-cropping of metal(loid) hyperaccumulators with edible crops, such as maize, soybean or ryegrass, has also been shown to significantly improve the phytoextraction capacity of Zn/Cd hyperaccumulators (Jiang et al.-LDQJ et al., 2010). Moreover, multi-cropping of various hyperaccumulators resulted in a higher total biomass production and metal(loid) accumulation (Lucisine et al.5XH et al., 2015). Plant diversity and composition induces a wider variety of released rhizodeposits (Zak et al.%HQL]ULDQG$PLDXGWKXVJHQHUDWLQJDQLQFUHDVHLQVRLO bacterial community diversity and improvements in soil functionality (Benizri and $PLDXG*DR et al., 2010). Intercropping of agricultural crops with legumes has nutritional advantages compared to monocultures due to the ability of legumes to establish symbiotic interactions with N2 fixing bacteria (Bloem et al./LX et al., 2013). Legumes include economically important grain crops, oilseed crops, forage crops, and agroforestry species. Maize has traditionally been intercropped with legumes within conservation agriculture to inhibit soil degradation and control pests (Sekamatte et al.  1JZLUD et al., 2012). More recently, the legume-rhizobia symbiosis has been applied in metal(loid)-contaminated soils (Carrasco et al.'DU\ et al.+DR et al., 2013). Native leguminous species such as Medicago, Trifolium, Vicia,Lotus or Lupinus can be found growing in 43 Introduction metal(loid)-FRQWDPLQDWHGVLWHV3UDVDGDQG'H2OLYHLUD)UHLWDV&DUUDVFR et al.  +DR et al., 2013). Cluster roots of species such as Lupinus are also known to be involved in inorganic P solubilisation by organic acids exudation in response to P deficiency (Keerthisinghe et al.  1HXPDQQ DQG 0DUWLQRLD  9DQFH et al., 2003). Woody trees such as alders (Alnus sp.) also form symbioses with N2-fixing microorganisms (such as the genera Frankia sp.), ectomycorrhizae and arbuscular mycorrhizal fungi 7DUUDQW DQG 0LOOHU  Taleshi et al., 2009). Keleberda (1978) studied spoil heaps of a Mn quarry in Ukraine planted with Scotch pine (Pinus sylvestris) and black alder (Alnus glutinosa). Tree growth was healthier in mixed rather than in monoculture stands, and the presence of alder much improved the soil fertility. Nitrogen fixation can decrease soil pH due to nitric acid accumulation in the rhizosphere (Van Miegroet and Cole, 1984), which for phytoextraction purposes can in turn induce an increase in TE bioavailability to the co-cropped TE-accumulators. The role of plant associated-bacteria in gentle remediation options effectiveness Plant establishment and growth in TE-contaminated soils can be improved using plant growth-promoting bacteria (PGPB) and fungi (including mycorrhizae and endophytic fungi). This introduction will only discuss the beneficial effects of plant-associated bacteria on phytoremediation processes in TE-contaminated soils, however similar interactions also occur in organic-contaminated sites (see reviews by Vangronsveld et al.ÈOYDUH] et al..KDQ et al.6HJXUDDQG Ramos, 2013) as well as between plants and their associated fungi (see reviews by *|KUHDQG3DV]NRZVNL/HEHDX et al., 2008). Plant-associated bacteria include several groups: endophytic bacteria, which colonise the internal tissues of SODQWV ZLWKRXW FDXVLQJ QHJDWLYH HIIHFWV RQ SODQW KHDOWK SK\OORVSKHULF EDFWHULD ZKLFKLQKDELWWKHH[WHUQDOVXUIDFHVRISODQWSDUWVDQGUKL]REDFWHULDZKLFKDUHSUH sent in the rhizosphere soil in direct contact with plant roots (Sessitsch and 3XVFKHQUHLWHU:H\HQV et al., 2009). Using soil bacteria as bioferilisers is a common approach in agriculture to avoid excessive input of fertilisers (which can be a further source of pollution). The use of bacteria, which are able to fix atmospheric N2 or to solubilise unavailable P and Fe, have been identified as good candidates for sustainable biomass production in TE-contaminated lands (Weyens et al., 2009). Higher plants present associations with diazotrophic bacteria which can reduce N2 to the NH4+ form (that is then available to plants for uptake), occurring in the form of nodules (symbionts) or as free living cells. Inorganic P-solubilising Chapter 1 44 bacteria are able to mineralise insoluble phosphate and make it accessible for plant XSWDNH5RGULJXH]DQG)UDJD9HVVH\6LGHURSKRUHFRPSRXQGVDUH Fe(III)-specific chelating agents, these compounds are produced by bacteria and can bind insoluble Fe(III) making it available for plants (Crowley and Kraemer, 2007). Bacteria can also produce phytohormones such as auxins, cytokinins and JLEEHUHOOLQV ZKLFK VWLPXODWH SODQW JURZWK DQG GHYHORSPHQW 7DQLPRWR  Taghavi et al., 2009). Phytohormones are involved in root growth and root hair proliferation, the stimulation of cell division and tissue expansion, and in stomatal opening or modifying plant morphology and tissue extension (Hare et al. 9HVVH\  7DJKDYL et al., 2009). Plant associated-bacteria can also reduce plant stress through the synthesis of ACC deaminase which reduces the high levels of ethylene by consuming its immediate precursor, the 1-aminocyclopropane-1carboxylic acid (ACC) (Glick, 2014). Other well-known mechanisms of PGPBs include increasing water uptake, alteration of root morphology, production of antibiotics and the induction of plant defence mechanisms (Lin et al.YDQ/RRQ DQG%DNNHU.LGG et al., 2009). Plant associated-bacteria can also influence TE mobility, and consequently plant uptake and accumulation (see reviews by *DGG +DIHUEXUJDQG.RWKH:H\HQV et al.6HVVLWVFK et al., 2013). Plant-microbial interactions greatly influence TE mobility and availability to plants for uptake (Anderson et al.  0DUVFKQHU et al., 2004). Three specific mechanisms of microbial-induced mobilisation of TE and increased plant metal(loid) accumulation are: (i) increasing root surface area and hair production, (ii) increasing element solubility (iii) increasing soluble element transfer from the rhizosphere to the plant (Whiting et al., 2001a). An increase in soil metal(loid) mobility and availability can result in increased metal(loid) uptake and therefore phytoextracted metal(loid) (Kidd et al.:H\HQV et al.6HVVLWVFK et al., 2013). In contrast, microorganisms can immobilise metal(loid)s through sorption to cell components or exopolymers, transport and intracellular sequestration, release of metal(loid) binding compounds or precipitation as insoluble organic or inorganic molecules (Gadd, 2004). Mineralisation of dissolved metal(loid)-organic complexes may be another cause of microbial mediated immobilisation (Gadd, 2004). In contrast, precipitated metal(loid)s can be solubilised by acidification, chelation and ligand-induced dissolution (Gadd, 2004). Oxalate, malate and citrate are some of the most important organic acids identified in root and microbial H[XGDWHV(KUOLFK-RQHV The last two decades has seen an increase in the search for candidate bacte- 45 Introduction rial strains with potential application in GRO. Natural metal(loid)-enriched areas (such as serpentine soils) or TE-contaminated sites are not only a source of interesting plant species for application in phytoremediation but also of microorganisms (Schippers et al.  %DWW\  0HQJRQL et al., 2010). Numerous studies have focused on the isolation and characterisation of cultivable microorganisms from this type of substrate (Lodewyckx et al.  $ERXShanab et al.  ,GULV et al.  3DO et al.  %DU]DQWL et al.  Grandlic et al.:H\HQV et al., 2009). To be useful in phytoremediation processes the selected bacterial inoculants must be able to tolerate the potentially toxic concentrations of metal(loid)s present in the substrate but should also be adapted to the specific conditions of the site (such as nutrient deficiency). In serpentine areas, these abiotic factors have often been suggested to be the cause of the low number of microorganisms in WKHVH VRLOV /LSPDQ  $FHD DQG &DUEDOODV  3DO et al., 2005). Nonetheless, bacterial communities in metal(loid)-enriched environments present a high genetic diversity. Moreover, highly metal(loid)-tolerant bacterial strains have been isolated from serpentine soils (Mengoni et al.$ERX-Shanab et al.  $PLU DQG 3LQHDX  3DO et al., 2005). TE-contaminated mine tailings are also an important source of highly metal(loid)-tolerant bacteria. Numerous metal(loid)-tolerant bacterial strains belonging to diverse taxonomic groups have been isolated from mining sites, as well as domestic and industrial ZDVWHV6WRSSHODQG6FKOHJHO6FKLSSHUV et al.%DWW\6SURFDWL et al., 2006). Bacterial densities are higher in the rhizosphere than in non-vegetated soil, GXH WR WKH UHOHDVH RI HDVLO\ GHJUDGDEOH ODELOH & FRPSRXQGV 5RXDWW  Grayston et al., 1998). The rhizosphere of (pseudo)metallophyte plants provides a metal(loid)-rich niche where populations of metal(loid)-resistant bacteria can establish (Mengoni et al.  /RGHZ\FN[ et al.  $ERX-Shanab et al., $ERXGUDU et al..XIIQHU et al.%HFHUUD-Castro et al. Janssen et al., 2015). Schlegel et al. (1991) detected higher numbers of TEresistant bacteria in the rhizosphere of Ni accumulating plants. Similarly, Idris et al. (2004) identified a high number of different Ni-resistant bacteria in the rhizosphere of the Ni-hyperaccumulator Noccaea goesingensis. Cultivation of bacteria on Ni-containing medium resulted mostly in the isolation of Methylobacterium spp., an alphaproteobacterial genus, as well as Rhodococcus spp. and Okibacterium spp., belonging to the Actinobacteria (or Gram-positive Chapter 1 52 (Continued) Inoculation method Bacterial inoculant Plant Host Soil / substrate type Amount of soil / substrate Experiment duration Plant Seed Substrate Re-inoculation (weeks) Cellular density Reference Pseudomonas sp., Deltia lacustris, Bacillus sp., Variovorax boronicumulans, Pseudoxanthomonas mexicana Pteris vitatta As-contaminated landfill site 3000 g 120 d X applied at the beginning and after 2 months 108CFU g−1 soil Lampis et al. (2015) 8 Pseudomonas sp Zea mays Artificially Cu contaminated soil using CuCl2solution 200 g 30 d X X Applied together 7.5 x 108CFUs mL-1 Li and Ramakrishna (2011) Helianthus annuus Ralstonia eutropha, Chryseobacterium humi Helianthus annuus Agricultural soil artificially contaminated with Zn or Cd 300 g 140 d X 10 mL 1 x 108 CFUs mL-1 Marques et al. (2013) 4 Pseudomonas fluorescens Mirabilis jalapa Contaminated soil with Cd, Cr,Cu, Ni, Pb and Zn by a smelter dump 75 d X 1 - 2 x 109 CFUs mL-1 Petriccione et al. (2013) Micrococcus sp. Zea mays. cv. CPDK 888 Garden soil Cd-spiked soil 3000 g 112 d X repeated at 2, 4 and 6 weeks 106CFU g−1 soil Sangthong et al. (2015) Root endophyte isolated from N. caerulescens:Arthrobacter sp., Kocuria rhizophila Bacillus sp. Microbacterium oxydans, Bacillus amyloliquefaciens Noccaea caerulescens, Arabidopsis thaliana Ni-rich serpentine soil 750 g 60 d X 108CFU mL−1 (5 × 108cells mL−1) Visioli et al. (2015) Bacillus subtilis Brassica juncea var. 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This study aimed to characterise the culturable rhizobacterial community associated with two Ni-hyperaccumulators and to obtain a collection of isolates for application in Ni phytomining. Non-vegetated and rhizosphere soil samples were collected from the Ni-hyperaccumulator Alyssum serpyllifolium ssp. lusitanicum and Alyssum serpyllifolium ssp. malacitanum, as well as from non-hyperaccumulating plants (Dactylis glomerata, Santolina semidentata and Alyssum serpyllifolium ssp. serpyllifolium). Rhizobacteria were isolated and characterised genotypically and phenotypically. Hyperaccumulating Alyssum subspecies hosted higher densities of bacteria compared to either non-hyperaccumulators or non-vegetated soil. In some cases hyperaccumulators showed selective enrichment of Ni-tolerant bacteria. Most bacterial strains belonged to the Actinobacteria phylum and presented Ni resistance. Phosphorus-solubilisers were mostly associated with the hyperaccumulators, siderophore-producers with D. glomerata, and IAAproducers with both these species. Taxonomic diversity and phenotypic characteristics were soil-, plant speciesand plant population-specific. Several strains presented PGP characteristics which could be useful when selecting microorganisms for bioaugmentation trials. &KDSWHU 5KL]REDFWHULDO FRPPXQLWLHV DVVRFLDWHG ZLWK WKH IORUD RI WKUHH VHUSHQWLQH RXWFURSV RIWKH,EHULDQ3HQLQVXOD This study formed part of the following publication: Álvarez-López, V., Prieto-Fernández, A., Becerra-Castro, C., Monterroso, C., Kidd, P.S. (2016).Rhizobacterial communities associated with the flora of three serpentine outcrops of the Iberian Peninsula.Plant Soil 403:233-252. 87 Rhizobacterial communities from serpentine soils 3.1. Introduction Serpentine soils are derived from ultramafic rocks, where the term ultramafic refers to igneous or metamorphic rocks containing more than 70% of ferromagnesium minerals and a low content in silicon (< 45% SiO2) (Brooks, 1987). Ultramafic rocks are patchily distributed throughout the world (occupying approximately 1% of the earth’s surface area (Proctor, 1999)) and are well-known for their physical and chemical anomalies that present a hostile environment for plant growth. Some common traits include an elevated concentration of Mg and Fe, a low availability of Ca relative to Mg (unfavourable for Ca absorption), a deficiency in essential nutrients such as N, P and K, and high concentrations of potentially phytotoxic trace metals such as Ni, Co and Cr. As a result the plant communities in these areas often present a high number of endemic species (serpentinophytes), and have evolved both morphological and physiological adaptations differentiating them from the flora of adjacent geological substrates (Brooks, 1987). Serpentine flora includes an unusual plant group, the so-called hyperaccumulators, which are able to accumulate extremely high concentrations of Ni in their aerial biomass (>1000 mg kg-1 dry weight (DW) matter) (Baker and %URRNV&KDQH\et al., 2007). The Iberian Peninsula hosts two subspecies of Alyssum serpyllifolium Desf. which are both serpentine-endemic and hyperaccumulators of Ni: Alyssum serpyllifolium ssp. lusitanicum from Galicia (NW Spain) and Trás-os-Montes (NE Portugal), and Alyssum serpyllifolium ssp. malacitanum from Andalucía (S Spain). Nickel-hyperaccumulating plants are considered ideal candidates for application in phytomining, a non-destructive approach for the recovery of high value metals (e.g. Ni) from metal-enriched soils and ores. Plants are cultivated to accumulate trace metals from soils and transport them to the shoots which can then be harvested (Chaney et al., 2007). Bioaugmentation using bacteria associated with hyperaccumulators can improve the plant’s capacity to phytoextract metals from soils (Mengoni et al.,  %HFHUUD-Castro et al.,  6HVVLWVFK et al., 2013). Plant-associated bacteria can enhance plant growth, reduce stress and/or modify soil metal bioavailability (Lebeau et al., .LGGet al., %HFHUUDCastro et al.,  6HVVLWVFK et al.,  &DEHOOR-Conejo et al.,  *OLFN 2014). Serpentine soils are a potential source of metal-tolerant (Co, Cr and Ni) bacteria (Pal et al., 2005). Schlegel et al. (1991) found bacterial strains isolated from serpentine soils tolerated up to 10-20 mM Ni (in the culture medium), while Chapter 3 88 strains from other soil types tolerated only 1 mM Ni. Pal et al. (2007) found that bacterial strains isolated from the rhizosphere of the Ni-hyperaccumulators Rinorea bengalensis and Dichapetalum gelonoides tolerated up to 28.9 mM of Ni in the culture medium, while Turgay et al. (2012) found that, bacterial strains isolated from Turkish serpentine soils, tolerated up to 34 mM Ni in the growth medium. Furthermore, the rhizosphere bacterial communities associated with Nihyperaccumulating plants have been shown to differ from those of nonaccumulating plants growing at the same site or of non-vegetated soil, and are also characterised by a higher number of Ni-tolerant bacteria (Schlegel et al.,  Mengoni et al., $ERX-Shanab et al., ,GULVet al., 2004). This selective enrichment in Ni-tolerant bacteria in the rhizosphere has been correlated with an increase in soil Ni availability (Becerra-Castro et al., 2009). Amongst the culturable Ni-tolerant bacterial strains isolated from serpentine soils and/or associated with Ni-hyperaccumulating plants members of the Actinobacteria, Acidobacteria, Chlorobi,Firmicutes,Verrucomicrobia and Proteobacteria have been described (Mengoni et al., 2OLQH3DOet al., 7XUJD\et al., 2012). In this study, the culturable bacterial community associated with different populations of the two Ni-hyperaccumulating subspecies of Alyssum serpyllifolium (subsp. lusitanicum and subsp. malacitanum) of the Iberian Peninsula were characterised. The rhizosphere bacterial communities associated with these plants were compared with those of non-hyperaccumulating plant species (the Ni-excluder Dactylis glomerata and the facultative serpentinophyte Santolina semidentata) growing at the same sites, as well as with the nonhyperaccumulating Alyssum serpyllifolium subsp. serpyllifolium growing in calcareous soils (developed over limestone and dolomite) in Sierra Nevada (S Spain). A collection of rhizobacterial isolates was obtained and the strains were characterised both genotypically (BOX-PCR and 16S rDNA partial sequencing) and phenotypically (for their resistance to Ni and plant growth promoting (PGP) traits). We focused on the culturable bacterial community since the global aim is to obtain potentially useful isolates which can be used to improve the Ni phytoextraction capacity of Ni-(hyper)accumulating plant species. 3.2. Materials and methods Study areas and collection of samples The study was carried out in five areas of the Iberian Peninsula in which 89 Rhizobacterial communities from serpentine soils three subspecies of Alyssum serpyllifolium Desf. (Brassicaceae) are found growing. Four of these areas are serpentine outcrops: two sites in Trás-os-Montes 6DPLO6′48²1 ′47²:DQG0RUDLV 0′21²1′20²W), NW Portugal), one site in Barazón (L) (42°51′09²1′15²W, NW Spain) and one in Sierra Bermeja (SB) (36°28′48²1′52²W, S Spain). These serpentinitic areas host the two Ni-hyperaccumulating subspecies of A. serpyllifolium which are endemic to the Iberian Peninsula: A. serpyllifolium subsp. lusitanicum Dudley and P. Silva (hereafter referred to as A. pintodasilvae) in S, M and L, and A. serpyllifolium ssp. malacitanum Rivas Goday (hereafter referred to as A. malacitanum) in SB. The fifth sampling site was the calcareous dolomitic area of 6LHUUD 1HYDGD 61 ´1 ´: 6 6SDLQ ZKHUH WKH QRQhyperaccumulator A. serpyllifolium subsp. serpyllifolium Desfontaines grows (hereafter referred to as A. serpyllifolium). Both Ni-hyperaccumulating subspecies presented leaf Ni concentrations above 10 g kg-1 and the highest Ni accumulation was found in the Spanish population of L (15.5 g kg-17DEOH)RUFRPSDUD tive purposes non-hyperaccumulating plants were also collected in some study sites: Dactylis glomerata L. was sampled at sites M, S and L, Santolina semidentata Hoffmanns. & Link. was sampled at sites M and S. Trás-os-Montes has a Mediterranean climate, with a mean annual temperature of 12.4ºC and annual precipitation of 720 mm (Carballeira et al., 1983) (Menezes de Sequeira and Pinto da Silva, 1992). Barazón has a European humidtemperate climate with a mean annual temperature of 12.9ºC and mean annual precipitation of 1381 mm (Carballeira et al., 1983). Sierra Bermeja and Sierra Nevada have a Mediterranean oceanic climate (Rivas-Martinez and Rivas-Saenz, 1996-2009), Sierra Bermeja has a mean annual precipitation between 800-1600 mm and mean annual temperature between 14-16ºC (Gómez-Zotano et al., 2014) and Sierra Nevada has a mean temperature of 12ºC and mean annual precipitation between 450 and 1000 mm (Castillo-Martín, 2000) . At each site, 5 to 7 samples of non-vegetated soil (0-15 cm) or of rhizosphere soil were collected for the isolation of bacterial strains. The whole plant and root system (including root-adhering soil) of each species were collected at late flowering stage, and after gently crushing the root ball the tightly held soil (< 3 mm from the root surface) was considered as rhizosphere soil. Samples were sieved (< 4 mm) and kept at 4ºC until processing. Soil samples are named as follows: L, M, S, SB and SN according to the sampling site (Barazón, Morais, 6DPLO 6LHUUD %HUPHMD DQG 6LHUUD 1HYDGD UHVSHFWLYHO\ DQG DV $*DQG6 Chapter 3 90 according to the plant species (Alyssum,Dactylis and Santolina, respectively). Rhizosphere soil is indicated by an R and the non-vegetated soil by NV. Elemental analysis of field-collected soils and plant material Non-vegetated (NV) and rhizosphere samples were air-dried and sieved through a 2-mm stainless steel sieve. Soil pH was measured in H2O and KCl using a 1:2.5 soil:solution ratio. Total C and N were analysed by combustion with a CHN analyser (Model CHN-1000, LECO Corp., St Joseph, MI). Exchangeable cations (Ca, Mg, Al, Na and K) were extracted with 1 M NH4Cl and determined by inductively coupled plasma optical emission spectrometry (ICP-OES, model Vista-PRO, Varian). Soils were digested in a 3:1 mixture of concentrated HNO3:HCl and the total concentrations of Co, Cr and Ni were analysed by ICPOES. Soil metal availability was evaluated after extraction with water after 24 h shaking using a 1:2.5 soil:H2O ratio. Plant material collected in the field was separated in shoots and roots, washed with pressurised tap water followed by deionised water, oven–dried at 45ºC, weighed and ground. Plant shoots (approximately 0.1 g) were digested in a 2:1 concentrated HNO3:HCl mixture on a hot plate at 120ºC, and the concentration of metals was measured by ICP-OES. Microbiological analyses Five grams of fresh rhizosphere soil were suspended in 45 mL sterile sodium hexametaphosphate solution (1%) and shaken for 30 min in an end-overend shaker. Soil suspensions were diluted in 10-fold series and plated in duplicate onto 284 agar medium (Schlegel et al., 1991) supplemented with 100 ȝg mL−1 of the fungicide cycloheximide. The 284 medium contains (per litre): 6.06 g TrisHCl, 4.68 g NaCl, 1.49 g KCl, 1.07 g NH4Cl, 0.43 g Na2SO4, 0.2g MgCl2.6H2O, 0.03 g CaCl2.2H2O, 0.04 g Na2HPO4.2H2O, 10 mL Fe(III)NH4 citrate solution (containing 48 mg/100mL) plus oligoelements (1.5 mg FeSO4.7H2O, 0.3 mg H3BO4, 0.19 mg CoCl2.H2O, 0.1 mg MnCl2.4H2O, 0.08 mg ZnSO4.7H2O, 0.02 mg CuSO4.5H2O, 0.036 mg Na2MoO4.2H2O) adjusted to pH 7. The medium was supplemented with a mixture of different carbon sources: lactate (0.7g L−1), glucose (0.5g L−1), gluconate (0.7g L−1), fructose (0.5g L−1), and succinate (0.8 g L−1). Densities of metal tolerant bacteria were determined in 284 agar media supplemented with an increasing concentration of Ni (0 mM, 0.5 mM, 1.0 mM, P0P0DGGHGDV1L624.6H2O). After 7 days incubation at 28ºC, colony 91 Rhizobacterial communities from serpentine soils forming units (CFUs) were counted and calculated per gram DW soil. Distinct colony morphotypes from Ni-enriched media associated with each plant species and from each site were sub-cultured at least three times to ensure purity and cryopreserved at −70ºC in culture medium supplemented with 15% (v/v) glycerol. Phenotypic characterisation of bacterial isolates Rhizobacterial strains were screened for Ni resistance, the ability to produce biosurfactants, and for various plant growth promoting characteristics: phosphate solubilisation capacity, siderophore production, organic acid production, and indoleacetic acid (IAA) production. Nickel resistance of the strains was tested using 284 agar medium (see above) supplemented with an increasing concentraWLRQRI1LP0P0P0P0P0DGGHGDV1L624.6H2O) and incubated at 28ºC for 7 days. The Maximal Tolerable Concentration (MTC) of Ni was recorded for each isolate, as the highest Ni concentration tested where the isolate was able to grow. The ability to solubilise inorganic phosphate was assessed in a modified NBRIP agar medium (1.8%) supplied with 5 g L−1 of hydroxyapatite and incubated at 28ºC for 5 days (10.0 g glucose, 5.0 g MgCl2.6H2O, 0.25 g MgSO4.7H2O, 0.2 g KCl, 0.1 g (NH4)2SO4, 0.1 g yeast H[WUDFW LQ  / GHLRQL]HG ZDWHU DGMXVWHG WR S+  PRGLILHG IURP 1DXWL\DO 1999)). A clear halo around the bacterial colony indicated solubilisation of mineral phosphate. Yeast extract was added since some strains were unable to grow in yeast-free NBRIP medium. Siderophore production was detected in a modified 284 liquid medium (without Fe) using the Chrome Azurol S (CAS) method described by Schwyn and Neilands (1987). All glassware used in this assay was previously cleaned with 30% HNO3 followed by washing in distilled water (Cox, 1994). Each isolate was screened for acid production. Single colonies were plated on agar medium containing 0.002% bromocresol purple (per litre medium): 10.0 g glucose, 1.0 g tryptone, 0.5 g yeast extract, 0.5 g NaCl, 0.03 g CaCl2.2H2O. Colonies forming a yellow halo after 1 day of growth at 28ºC indicated a pH change in the medium and were considered acid producers. The ability of isolated strains to produce IAA was evaluated in liquid medium (5.0 g glucose, 1.0 g (NH4)2SO4, 2.0 g K2HPO4, 0.5 g CaCO3, 0.5 g MgSO4.7H2O, 0.1 g NaCl, 0.1 g \HDVWH[WUDFWDGMXVWHGWRS+PRGLILHGIURP6KHQJet al. (2008) supplemented with 0.5 mg mL−1 of tryptophan). After 5 days incubation at 28ºC, cultures were centrifuged and the supernatant was incubated with Salkowski reagent for 25 min. The production of IAA was recognised by the presence of red colouring and iso- Chapter 3 92 lates were considered IAA-producers when the concentration of IAA determined was more than 4 mg L-1 culture. Strains were screened for potential biosurfactant production using the qualitative method of Chen et al. (2007). The strains were inoculated in 284 liquid medium and cultured overnight at 28ºC, at 150 rpm on a rotary shaker. A 100 ȝL sample was taken from the supernatant of each strain and added to a microwell of a 96-microwell plate. The plate was then viewed using a backing sheet of paper with a black and white grid. The optical distortion of the grid provided a qualitative assay for the presence of surfactants. Isolates were classified in different phenotype groups, according to their metal resistance (low Ni tolerance (LT): 2.5 mM Ni MTC, or high Ni tolerance (HT): 5-P01L07&DQG3*3WUDLWV1QRQ3*3WUDLWVSURGXFHUV$RUJDQLF acid-SURGXFHUV 3 3-VROXELOLVHUV 6 VLGHURSKRUH-SURGXFHUV 7 ELRVXUIDFWDQWVSURGXFHUV+LQGROHDFHWLFDFLG-producers and their possible combinations). Eighty four out of the 550 rhizobacterial isolates obtained were already phenotypically characterised in a prior study and used to evaluate the potential of the cell-free cultures to mobilise soil Ni under in vitro conditions (Becerra-Castro et al., 2011). This previous study targeted only rhizobacterial strains associated with Alyssum sp. and their capacity for soil Ni mobilisation. Genotypic characterisation of bacterial isolates BOX-PCR genomic DNA profiling BOX-PCR fingerprinting was used to genotype and group bacterial strains within each isolate collection (L, S, M, SB, SN). Crude cell lysates (colonies suspended in 100 ȝL and heated to 100ºC for 5 min) were used as DNA templates for BOX-PCR reactions. Box reactions were performed in a total volume of 20 ȝL containing: 1x Taq buffer (Invitrogen), 1.5 mM MgCl2, 0.1 mM of each dNTP, 0.5U Taq polymerase (Invitrogen), 2 ȝM of BOX A1R primer (5’- CTACGGCAAGGCGACGCTGACG-3’) (Versalovic et al., 1994), and 2 ȝL of cell lysate. 7KHUPRF\FOLQJFRQGLWLRQVZHUHF\FOHRI&IRUPLQF\FOHVRIPLQDW &PLQDW&DQGPLQDW&DQGF\FOHRIPLQDW&7KHRE tained PCR products were separated by gel electrophoresis in 1.8% agarose run for 3 h at 3.3 V cm−1 gel. Gel images were analysed, using the Pearson correlation coefficient and UPGMA clustering algorithm of the Gel Compar Bionumerics program (Bionumerics Version 6.6, Applied Maths, Belgium). Rhizobacterial isolates were grouped according to their BOX-PCR profiles at a similarity level of 93 Rhizobacterial communities from serpentine soils 82%. BOX-PCR fingerprints were analysed for each sampling site (L, S, M, SB and SN) and moreover, bacterial strains associated with the Ni-hyperaccumulator Alyssum subspecies were analysed separately (LA, SA, MA and SBA). Genetic diversity of isolates was assessed for each plant species and population using the Shannon diversity index (H’ = -(x i /x 0 )ln(x i /x 0 )) where x i is the number of strains per BOX-PCR group for each level of metal tolerance (LT or HT) ), and x 0 is the total number of strains in either LT or HT. Isolates within each BOX-PCR group were also classified according to the phenotypes (PGP traits and Ni resistance described above). DNA extraction and 16S rRNA gene amplification For DNA extraction, purified strains were grown in 1/10 strength 869 liquid medium (1.0 g tryptone, 0.5g yeast extract, 0.5 g NaCl, 0.1 g glucose, 0.035 g CaCl 2 .2H 2 O in 1L deionised water (Mergeay et al., 1985)) and genomic DNA was extracted from bacterial cell pellets. Briefly, the method consists of alkaline cell lysis followed by phenol/chloroform/isopropanol alcohol purification. DNA quality was checked by gel electrophoresis on a 0.8% agarose gel. PCR amplification targeting the 16S rRNA gene was carried out using the primers 16S-27F (5’-AGAGTTTGATCMTGGCTCAG-3’) and 16S-1492R (5’-TACGGYTACCTTGTTA CGACTT-3’) (Lane, 1991). PCR reactions were performed in a total volume of 50 ȝL containing: 1x Taq buffer (Invitrogen), 2.5 mM MgCl 2 , 0.1 mM of each dNTP, 1.75U Taq polymerase (Invitrogen), 0.4 ȝM of each primer, and 1 ȝL of extracted DNA. Thermocycling conditions were: 2 PLQDW&F\FOHVRIPLQDW&PLQDW&DQGPLQDW&DQG cycle of 10 min at 72ºC. PCR products were partially sequenced (between 750 and 1000 bases) using the primer 16S-27F (Lane, 1991). Sequence data were checked using the Chromas v. 1.45 software (Technelysium Pty. Ltd., Australia), uploaded DQG DOLJQHG LQ WKH 5LERVRPDO 'DWDEDVH 3URMHFW 5'3 &ROH et al. (2009)) and assessed for similarity with sequences in the RDP. The sequences used for identification of the culturable strains are available in the EMBL database (www.ebi.ac.uk) under accession numbers HG941722 - HG942010, HE646570 - HE646571, and FN908759 - FN908797 (the latter group were identified and described for Ni mobilisation capacity by Becerra-Castro et al. 2011). Genetic diversity of isolates was assessed for each plant species and population using the Shannon diversity index (H’ = -(x i /x 0 )ln(x i /x 0 )) where x i is the number of strains per genera, and x 0 is the total number of strains. Chapter 3 100 Siderophore-producing strains were found in the rhizosphere of all the Alyssum populations studied except in A. pintodasilvae from M. The frequency of isolates showing this characteristic varied between 6.8% in L and 33.3% in SN. The siderophore-producing ability was abundant in isolates from the rhizosphere of D. glomerata collected in M (51.2%) and S (21.7%). In contrast, in the rhizosphere of S. semidentata these were much less frequent (only 2.3% of isolates in S). The abilities to solubilise mineral phosphate or to produce biosurfactants were relatively rare traits amongst the bacterial isolates, representing 4.5 and 0.5% of the total number of the isolates, respectively. P solubilisers were not found in isolates from D. glomerata but were isolated from the rhizosphere of the Alyssum and S. semidentata populations (except in L). In Alyssum between 4.8 and 11.6% of the isolates were able to solubilise P while in S. semidentata this characteristic represented 2.2 and 4.6% of the isolates. Biosurfactant production was only detected amongst 3 rhizobacteria isolated from A. malacitanum (SB), S. semidentata (of M) and D. glomerata (of S), and the percentage was never greater than 3%. Figure 3.1b shows the frequency of isolates (as a % of the total number of isolates in each group) for each MTC. All of the serpentine plant species harboured rhizobacteria capable of growing in all the tested Ni concentrations. However, the level of Ni tolerance depended on both the soil type and the plant species. About 90% of rhizobacterial isolates associated with the nonhyperaccumulator (SN) were only able to grow in medium without Ni (MTC < 1 mM), while for the plant species growing in the serpentine sites from 4.2 to 18.5% of rhizobacteria presented a MTC of 10 mM. Of the rhizobacterial isolates associated with non-hyperaccumulators species growing on serpentine sites the majority presented MTC values < than 1 mM Ni. For instance, most of the isolates associated with the Ni-excluder D. glomerata (66-73%) did not tolerate Ni concentrations >2.5 mM, while in the rhizosphere of the Ni-hyperaccumulators this percentage ranged from 36 to 51%. Within the Ni-hyperaccumulators, only 9.4% of strains associated with the S population of A. pintodasilvae were unable to grow in the presence of Ni in the growth medium. The same population presented the highest percentage (45.3%) of associated microorganisms with a MTC of 2.5 mM Ni. While the L population presented the highest number of isolates with an MTC of 5 mM (30%) and the M population of 10 mM (18.5%). Similarly, it was the Ni-hyperaccumulators which presented the highest numbers of associated bacteria with an MTC of 10 mM Ni (6.8, 10.9 and 18.5% in the L, S and M sites, respectively). Nonetheless, a high 101 Rhizobacterial communities from serpentine soils percentage (at times similar or even higher than the Ni-hyperaccumulators) of isolates associated with the non-hyperaccumulator S. semidentata also presented elevated Ni MTC values (up to 5 or 10 mM). Table 3.3 summarises the frequency of each phenotype within the collection of isolates. The most frequent phenotype was LT-N (47% in L, 25% in M, 28% in S, 31% in SB and 38% in SN), which represents low Ni tolerance (up to 2.5 mM Ni MTC) and no PGP trait. This was followed in the serpentine sites, by HT-N (high Ni tolerance but no PGP trait) (18% in L, 16% in M, 12% in S, 23% in SB) and by LT-A in the SN site (17%). Phenotypes A and H (which correspond with organic acidand IAA-producers respectively), were also frequently found in LT strains (up to 18% in L) but were less often represented by HT strains (up to 10% in S). Phenotypes presenting three PGP traits (phenotypes APH, ASH and PSH) were rarely found (at most 2% of strains). In general, all phenotypes were represented in both ranges of metal tolerance (LT and HT), except phenotypes T and APH which were exclusive to HT isolates and phenotypes AP, PH, ST and PSH which were only present in LT isolates. Strains isolated from L showed a maximum of two PGP traits (HT-SH, siderophoreand IAA-SURGXFHUV 7DEOH  Whereas in S, M and SB several strains with more than two PGP traits were found SKHQRW\SHV$3+$6+DQG36+DQGWKHVHZHUHPDLQO\DVVRFLDWHGZLWKWKH1Lhyperaccumulator. Rhizobacterial strains associated with Alyssum ssp. generally presented a higher diversity in phenotypes than the other plant species. For example, in the L population A. pintodasilvae strains were allocated into 12 different phenotypes, while only five phenotypes were found in rhizobacteria from the gramineae growing in the same site, and none of these were specific to the grass species. In the Portuguese (M and S) populations, phenotype distribution depended on the plant species. In M, phenotypes LT-AP and HT-APH were associated with A. pintodasilvae, while HT-S, LT-AS and LT-SH (all of which include siderophore-producers) with D. glomerata, and finally, HT-T with S. semidentata. In S, the 18 phenotypes were distributed among the plant species as follows: A. pintodasilvae (15 groups) > D. glomerata (10) > S. semidentata  four of the groups were specific to the hyperaccumulator (LT-PH, LT-PSH, HTAS, and HT-ASH), two to D. glomerata (LT-ST and LT-ASH) and only one (HTAPH) was specific to S. semidentata. In SB and SN, strains associated with Alyssum sp. were represented by ASH and PH phenotypes, respectively. Chapter 3 102 Table 3.3. Frequency of bacterial phenotypes associated with the different plant species (A, Alyssum * Dactylis 6D Santolina) of each site: L (Barazón), S (Samil), M (Morais), SB (S. Bermeja) and SN (S. Nevada). L S M SB SN Phenot y pic g roup % A (%) G(%) %A (%) G (%) Sa (%) %A (%) G (%) Sa (%) %A (%) %A (%) LT N 47 21 26 28 14 7 8 25 7 13 4 31 31 38 38 HT N 18 11 7 12 3 4 4 16 8 2 6 23 23 LT A 7 3 3 11 5 1 4 18 4 2 11 12 12 17 17 HT A 5 3 2 10 2 8 4 1 3 2 2 LT P 3 2 1 1 1 1 4 4 5 5 HT P 4 4 LT S 1 1 8 1 7 1 4 1 3 5 5 17 17 HT S 2 2 1 1 1 1 1 2 2 LT T HT T 1 1 1 1 LT H 9 9 5 3 1 1 14 9 5 3 3 2 2 HT H 4 3 2 4 4 1 LT AP 1 1 1 1 HT AP LT AS 1 1 6 5 1 1 1 6 6 7 7 HT AS 1 1 LT AH 5 5 5 3 2 4 2 1 3 3 5 5 HT AH 2 2 4 1 4 LT PH 1 1 HT PH LT ST 1 1 HT ST LT SH 5 1 4 3 3 5 5 HT SH 1 1 LT APH HT APH 1 1 1 1 LT ASH 1 1 2 2 2 2 HT ASH 1 1 LT PSH 1 1 1 1 HT PSH Number of strains 74 48 64 41 43 54 44 45 95 42  LT (low Ni tolerance  2.5 mM Ni MTC), HT (high Ni tolerance 5-P01L07&1QR3*3WUDLWV$RUJDQLFDFLG-SURGXFHUV33-sROXELOLVHUV6 Siderophore-SURGXFHUV7ELRVXUIDFWDQWV-SURGXFHUV+LQGROHDFHWLFDFLG-producers and combinations. 103 Rhizobacterial communities from serpentine soils Identification and genotypic diversity of culturable rhizobacteria BOX-PCR profiles A BOX-PCR profile was obtained for 538 out of the 550 isolated bacterial strains. Profiles were analysed on a site-by-site basis (for each plant species present) and for all the Ni-hyperaccumulating Alyssum species together (Table 3.4 and Table 3.5). Table 3.6 presents a summary of the total number of BOX-PCR groups obtained in each of these analyses and the distribution of bacterial isolates amongst these groups for each site (Table 3.6a) and plant species (Table 3.6b). In the case of L the BOX-PCR profiles of isolates were distributed amongst 19 groups, strains associated with the Ni-hyperaccumulator A. pintodasilvae were distributed amongst 13 groups and those of D. glomerata in 13 groups. Twelve of these BOX-PCR groups are represented exclusively by strains of either A. pintodasilvae or D. glomerata (groups L01-L02, L04, L07-L08, L11-L12, L14 and L16-L19), while seven groups are made-up of isolates from both species (Table 3.4). In the case of S, the BOX-PCR profiles of isolates from A. pintodasilvae and S. semidentata were distributed amongst 19 and 9 groups each, while for D. glomerata isolates were distributed amongst only 4 groups. Twelve groups were exclusively made up of isolates associated with A. pintodasilvae (S01, S02, S07, S09, S11, S12, S15, S16, S17, S18, S20 and S21), three groups with isolates of S. semidentata (S10, S19 and S22), and some groups (S04, S08 and S13) were composed of isolates from both the Ni-hyperaccumulator and S. semidentata. There were no groups solely represented by isolates of D. glomerata (Table 3.5). In the case of plant populations of M, isolates were distributed in 24 BOX-3&5JURXSV A. pintodasilvae isolates were distributed amongst 17 groups, S. semidentata amongst 17 groups and D. glomerata amongst 8 groups. The groups are generally represented by isolates from all three plant species, with some exceptions being group M01 (solely represented by isolates of D. glomerata), M02, M10, M19, M23-M24 (each represented by a single isolate of A. pintodasilvae), M03, M14, M16-18 and M22 (isolates of S. semidentata) (Table 3.4). Rhizobacterial strains isolated from A. malacitanum (from SB) and from A. serpyllifolium (from SN) were allocated into 18 and 10 BOX-PCR groups, respectively (Table 3.4). Chapter 3 104 Table 3.4. BOX-PCR groups and phenotypic characteristics of rhizobacterial strains of Alyssum sp. (A), Dactylis glomerata (G) isolated from Barazón (L), Samil (S), Morais (M), S. Bermeja (SB) and S. Nevada (SN). OA, organic acid-SURGXFHU 3 SKRVSKDWH VROXELOLVDWLRQ FDSDFLW\ 6I ELRVXUIDFWDQW-SURGXFHU 6G VLGHURSKRUH-SURGXFHU ,$$ LQGROHDFHWLF DFLGproducer. In Barazón, BOX PCR groups were based on profiles obtained for 119 bacterial strains (72 correspond with LAR, 47 with LGR), In Samil, BOX PCR groups were based on profiles obtained for 144 bacterial strains (61 associated with SAR, 41 with SGR, 42 with SSR). In Morais, BOX PCR groups were based on profiles obtained for 139 bacterial strains (53 associated with MAR, 41 with MGR and 45 with MSR). In S. Bermeja, BOX PCR groups were based on profiles obtained for 94 bacterial strains and in S. Nevada groups were based on profiles obtained for 42 bacterial strains. The number of strains for each phenotype and plant are given for each BOX group. Box Group OA P Sd Sf IAA Phenotype A (nº) G (nº) BARAZÓN L01 - - - - - LT N 1 + - - - - LT A 1 L02 - - - - - HT N 2 - - - - + HT H 1 L03 - - - - - LT N 5 10 + - - - - LT A 1 - - + - - LT S 1 - - - - + LT H 2 + - - - + LT AH 1 - - - - - HT N 1 3 + - - - - HT A 1 - - + - - HT S 1 - - - - + HT H 1 + - - - + HT AH 1 L04 - - - - + LT H 2 + - - - + LT AH 1 L05 - - - - - HT N 1 - - + - + HT SH 1 L06 - - - - + LT H 1 - - - - - HT N 1 1 L07 + - - - - HT A 1 L08 - - - - - LT N 1 L09 - - - - - LT N 11 7 + - - - - LT A 1 - - - - + LT H 3 + - - - + LT AH 2 - - - - - HT N 4 1 + - - - - HT A 1 2 + - - - + HT AH 1 L10 - - - - - LT N 5 3 + - - - - LT A 3 1 - - - - + LT H 1 + - + - - LT AS 1 + - - - + LT AH 1 - - - - - HT N 2 L11 - - - - - LT N 1 + - - - - LT A 1 L12 - - - - - LT N 1 105 Rhizobacterial communities from serpentine soils Box Group OA P Sd Sf IAA Phenotype A (nº) G (nº) L13 - - - - - LT N 3 5 - - - - + LT H 1 - - - - - HT N 4 - - + - - HT S 1 - - - - + HT H 1 L14 + - - - + LT AH 1 - - - - + HT H 1 L15 - - - - - LT N 1 - - - - - HT N 1 L16 - - - - - LT N 1 L17 - - - - - LT N 1 L18 + - - - - HT A 1 L19 - - - - + LT H 1 Box Group OA P Sd Sf IAA Phenotype A (nº) G (nº) Sa (nº) SAMIL S01 + - - - - HT A 1 S02 - - - - + LT H 1 S03 + - - - - LT A 1 - + + - + LT PSH 1 S04 - - - - - LT N 1 + - + - - LT AS 2 S05 - - - - - LT N 4 3 + - - - - LT A 3 1 - + - - - LT P 1 - - + - - LT S 1 - - - - + LT H 2 + - + - - LT AS 4 + - - - + LT AH 2 - - + - + LT SH 2 2 - - - - - HT N 1 2 + - - - - HT A 1 4 + - + - + HT ASH 1 S06 - - - - - LT N 5 7 7 + - - - - LT A 2 2 1 - + - - - LT P 1 - - + - - LT S 10 1 - - - - + LT H 2 + - + - - LT AS 2 + - - - + LT AH 1 2 - - + + - LT ST 1 - - + - + LT SH 3 + - + - + LT ASH 1 - - - - - HT N 3 5 2 + - - - - HT A 1 6 - - + - - HT S 1 1 + + - - + HT APH 1 S07 - - - - + LT H 1 S08 - - - - - LT N 1 1 S09 - - - - - LT N 1 S10 - + - - - LT P 1 S11 + - - - - LT A 1 + - + - - LT AS 1  Chapter 3 106 (continued) Box Group OA P Sd Sf IAA Phenotype A (nº) G (nº) Sa (nº) S12 - - - - - LT N 1 + - - - - LT A 1 S13 - - - - - LT N 1 + - - - + LT AH 1 - + - - + LT PH 1 - - - - - HT N 1 + - - - - HT A 1 S14 - - - - - LT N 2 1 + - - - - LT A 2 + - - - + LT AH 1 - - + - + LT SH 1 - - - - - HT N 1 + - + - - HT AS 1 S15 - - - - - LT N 1 - + - - + LT PH 1 S16 - - - - - LT N 1 S17 - - - - - LT N 1 - + - - - LT P 1 S18 + - - - - LT A 1 S19 + - - - - LT A 1 - - - - + LT H 1 - - - - - HT N 1 S20 - - - - - HT N 1 S21 - - - - - LT N 1 S22 + - - - - HT A 1 Box Group OA P Sd Sf IAA Phenotype A (nº) G (nº) Sa (nº) MORAIS M01 - - - - - LT N 1 - - - - + LT H 1 + - + - - LT AS 1 M02 + - - - - LT A 1 M03 - - - - - LT N 2 + - - - - LT A 1 M04 - - - - - LT N 6 8 2 + - - - - LT A 2 1 - - + - - LT S 1 1 - - - - + LT H 8 1 + + - - - LT AP 1 + - - - + LT AH 1 - - + - + LT SH 1 - - - - - HT N 2 1 1 - - + - - HT S 1 - - - - + HT H 1 + - - - + HT AH 1 M05 + - - - - LT A 1 + - - - + LT AH 1 M06 - - - - - LT N 1 - - - - - HT N 1 107 Rhizobacterial communities from serpentine soils Box Group OA PSd Sf IAA Phenotype A (nº) G (nº) Sa (nº) M07 - - - - - LT N 3 + - - - - LT A 1 - - + - - LT S 1 - - + - + LT SH 1 - - - - - HT N 2 + - - - - HT A 2 M08 - - - - - LT N 1 4 1 + - - - - LT A 1 2 5 - - + - - LT S 2 - - - - + LT H 1 2 - - + -+ LT SH 2 - - - - - HT N 1 1 4 + - - - - HT A 1 - - - - + HT H 1 1 + - - - + HT AH 2 + + - - + HT APH 1 M09 + - - - - LT A 1 - - - - + LT H 1 - - - + -HT T 1 - - - - + HT H 1 M10 - - - - + HT H 1 M11 + - - - - LT A 1 1 - - - - + LT H 1 2 - - - - - HT N 1 + - - - - HT A 1 - - - - + HT H 1 + - - - + HT AH 1 M12 - - - - - LT N 1 + - - - - LT A 1 2 - - - - + LT H 2 + - - - + LT AH 1 - - - - - HT N 2 1 2 M13 + - - - - LT A 1 - - - - - HT N 1 M14 - - - - - LT N 1 M15 + - - - - LT A 1 1 + - - - + LT AH 1 + - - - + HT AH 1 M16 + - - - + HT AH 1 M17 - + - - - LT P 1 M18 + - - - - LT A 1 M19 - - - - - LT N 1 M20 -+ - - - LT P 1 + - - - + LT AH 1 + - - - - HT A 1 M21 - - - - - LT N 1 - - - - + LT H 1 M22 - - - - - HT N 1 M23 - - - - - HT N 1 M24 - - - - - HT N 1  Chapter 3 108 (continued) Box Group OA P Sd Sf IAA Phenotype A (nº) S. BERMEJA SB01 - - - - - LT N 1 - - + - - LT S 1 - - - - - HT N 1 SB02 + - + - - LT AS 1 SB03 - - - - - LT N 1 - - - - - HT N 1 SB04 + - - - - LT A 3 - - - - + LT H 1 - - - - - HT N 5 - + - - - HT P 1 SB05 + + - - - LT AP 1 SB06 - - - - - LT N 12 + - - - - LT A 2 - + - - - LT P 1 - - + - - LT S 1 - - - - + LT H 1 + - + - + LT ASH 2 - - - - - HT N 5 + - - - - HT A 1 - + - - - HT P 1 - - - + - HT T 1 SB07 - - + - - LT S 1 SB08 - - - - - LT N 2 + - - - - LT A 3 + - + - - LT AS 2 + - - - + LT AH 2 - - - - - HT N 3 - + - - - HT P 1 SB09 - - - - - LT N 10 + - - - - LT A 2 - + - - - LT P 3 - - + - - LT S 1 + - + - - LT AS 1 + - - - + LT AH 1 - + + - + LT PSH 1 - - - - - HT N 6 - + - - - HT P 1 SB10 - - - - - LT N 1 - - + - - LT S 1 SB11 - - - - + LT H 1 SB12 + - + - - LT AS 1 SB13 + - - - - LT A 1 SB14 - - - - - LT N 1 SB15 + - + - - LT AS 1 SB16 + - - - - HT A 1 SB17 - - - - - HT N 1 SB18 - - - - - LT N 1 109 Rhizobacterial communities from serpentine soils Box Group OA P Sd Sf IAA Phenotype A (nº) S. NEVADA SN01 - - - - - LT N 3 + - - - - LT A 2 - - + - - LT S 1 SN02 - - - - - LT N 1 SN03 - - - - - LT N 1 + - - - - LT A 1 SN04 + - + - - LT AS 1 SN05 - - - - - LT N 4 + - - - - LT A 3 - + - - - LT P 1 - - + - - LT S 1 + - + - - LT AS 2 + - - - + LT AH 1 - - + - - HT S 1 SN06 - + - - - LT P 1 - - + - - LT S 1 SN07 - - - - - LT N 5 - - + - - LT S 1 - - - - + LT H 1 + - - - + LT AH 1 + - + - + LT ASH 1 SN08 - - + - - LT S 1 SN10 - - - - - LT N 2 + - - - - LT A 1 - - + - - LT S 1 - - + - + LT SH 2 SN11 - - + - - LT S 1  LT (low Ni tolerance  2.5 mM Ni MTC), HT (high Ni tolerance 5-10 mM Ni MTC).N: no PGP WUDLWV $ RUJDQLF DFLG-SURGXFHUV 3 3-VROXELOLVHUV 6 6LGHURSKRUH-SURGXFHUV 7 ELRVXUIDFWDQWVSURGXFHUV+LQGROHDFHWLFDFLG-producers 212 Chapter 6 (continued) Element (g kg -1 ) Plant species Soil treatment Bacterial inoculant NI P87 P29 P30 P64 P75 Fe N. tabacum Unt 0.59 ± 0.08 0.42 ± 0.06 0.49 ± 0.06 0.43 ± 0.02 0.42 ± 0.05 0.41 ± 0.03 * Comp 0.27 ± 0.03 0.37 ± 0.04 0.40 ± 0.09 0.30 ± 0.05 0.35 ± 0.04 0.43 ± 0.07 S. caprea Leaves Unt 0.21 ± 0.03 0.57 ± 0.15 * 0.55 ± 0.09 * 0.39 ± 0.05 * 0.38 ± 0.02 * 0.33 ± 0.04 * Comp 0.18 ± 0.02 0.12 ± 0.01 0.17 ± 0.03 0.19 ± 0.04 0.11 ± 0.01 * 0.13 ± 0.02 Stem Unt 0.06 ± 0.01 0.15 ± 0.07 * 0.08 ± 0.01 0.09 ± 0.02 0.08 ± 0.01 0.23 ± 0.07 * Comp 0.06 ± 0.00 0.07 ± 0.01 0.12 ± 0.02 * 0.07 ± 0.00 0.05 ± 0.01 0.09 ± 0.01 P N. tabacum Unt 1.96 ± 0.03 1.76 ± 0.09 1.30 ± 0.09 * 1.46 ± 0.07 * 1.76 ± 0.16 1.51 ± 0.05 * Comp 2.40 ± 0.13 2.11 ± 0.04 2.26 ± 0.10 2.36 ± 0.12 2.44 ± 0.13 2.51 ± 0.09 S. caprea Leaves Unt 1.05 ± 0.17 0.97 ± 0.20 1.17 ± 0.07 0.75 ± 0.10 1.83 ± 0.68 0.77 ± 0.12 Comp 1.17 ± 0.23 0.84 ± 0.09 0.94 ± 0.11 0.84 ± 0.08 0.72 ± 0.13 0.83 ± 0.06 Stem Unt 1.01 ± 0.07 1.20 ± 0.07 1.24 ± 0.08 1.12 ± 0.07 1.10 ± 0.02 1.26 ± 0.06 * Comp 1.54 ± 0.05 1.47 ± 0.13 1.49 ± 0.11 1.16 ± 0.05 1.42 ± 0.13 1.68 ± 0.18  8QW8QWUHDWHGVRLO&RPS&RPSRVW-amended soil 213 Evaluating compost and rhizobacterial inoculants for improving phytoextraction of up to 2.7-fold with strain P87 were observed (p < 0.05). Bioaugmentation decreased P content in tobacco grown in untreated soil, particularly with strains P29, P30 and P75 (p < 0.05). The same tendency was observed in leaves of S. caprea grown in compost-amended soil and, albeit less markedly, also in untreated soils. Curiously the opposite effect was found in the stems of S. caprea grown in untreated soil, where in the case of P75 a significant increase in stem P concentration was observed (p < 0.05). Strains P64 and P87 induced an increase in K content of tobacco growing in untreated soils (significant for P64, p < 0.05). On the contrary, strain P30 significantly reduced K content in tobacco grown in both untreated and compost-amended soils. No clear effects of inoculants on K concentrations were observed for S. caprea. Effects of bacterial inoculants on soil metal availability and fractionation Metal soil availability was assessed at the end of the experiment after extraction with NH4NO3. Some bacterial-induced changes in metal availability were observed, and these were more pronounced in the untreated soil. After growth of N. tabacum all the bacterial strains significantly increased available Zn and Cd in this soil (p < 0.05) (Figure 6.3). Strain P75 showed the highest ability to increase Zn availability (NH4NO3-extractable concentrations increased from 30.7 mg kg-1 in NI soil to 39.8 mg kg-1 after inoculation), while strain P29 caused the highest increase in Cd availability (increasing from 0.16 mg kg-1 to 0.26 mg kg-1). No bacterial-induced changes were detected for Zn in the compost-amended soils, while for Cd, strains P87, P30 and P75 significantly decreased the availability of this element (p < 0.05). In the case of Salix, strain P87 significantly increased soil Zn availability (from 32.1 mg kg-1 to 36.9 mg kg-1 (p < 0.05)), and all strains (except P87) significantly increased Cd availability in the untreated soils (p < 0.05) (Figure 6.3). The most pronounced increase was found for P29: NH4NO3extractable concentrations increased from 0.16 mg Cd kg-1 in NI soil to 0.24 mg Cd kg-1 after inoculation. In the compost-amended soils strains P64 and P29 significantly increased soil Zn and Cd availability, respectively (from 10.1 mg kg-1 to 13.8 mg kg-1 and from0.14 mg kg-1 to 0.18 mg kg-1 in the case of Zn and Cd, respectively). 214 Chapter 6 Figure 6.4 shows the soil metal fractionation (each fraction is represented as the % of the total) after plant growth in the NI treatment (Figure 6.4a), as well as the bacterial-induced depletion / increase for each fraction relative to the NI treatment (Figure 6.4b). Only those cases where a bacterial-induced effect on metal fractionation was observed are presented. In the untreated soils and after plant growth, Zn was mainly present in the acid-extractable fraction (2600 mg kg-1 and 2480 mg kg-1 in N. tabacum and S. caprea planted soils, respectively) and residual phases (2590 mg kg-1 and 2360 mg kg-1 in N. tabacum and S. caprea planted soils, respectively). In the compostamended soils a reduction of about 50% was observed in the most labile Zn fraction (F1) and an increase in the residual fraction. In these soils the F2 fraction increased and F4 decreased after the growth of S. caprea FRQYHUVHO\ WKH ) fraction decreased and F4 increased after growth of N. tabacum. Nevertheless, the dominant Zn pools remained the same. In untreated soils after the growth of either plant species about 50% of the total Cd was soluble and exchangeable (2.9 mg kg-1 and 4.4 mg kg-1 in N. tabacum and S. caprea, respectively) or acid-extractable (5.4 mg kg-1 and 3.2 mg kg-1 in N. tabacum and S. caprea, respectively). While in compost-amended soils these two fractions decreased and the reducible and N. tabacum S. caprea Figure 6.3. Effects of bacterial inoculants on metal availability (NH 4 NO 3 –extractable metals) in N. tabacum and S. caprea planted soils. Asterisks indicate significant differences from the NI plants (p < 0.05) 215 Evaluating compost and rhizobacterial inoculants for improving phytoextraction residual Cd fraction increased. Moreover, a 1.7-fold increase in the oxidisable Cd fraction (F4) was also detected in the case of tobacco. Bacterial inoculation significantly influenced soil metal fractionation, and these changes depended on both the bacterial strain and the soil treatment. In untreated soils where N. tabacum was grown, strains P29 and P30 significantly increased the Zn concentration in the most plant-available F1 fraction (from 191 mg kg-1 in NI soils to 232 mg kg-1 and to 218 mg kg-1, respectively). (a) 216 Chapter 6 Figure 6.4. Metal fractionation in the NI soil (a) and mean relative (percent) bacterially induced depletion / increase of each fraction relative to the NI treatment (n = 6) in N. tabacum and S. caprea planted soils (b). Different letters in metal fractionation indicate significant differences between different fractions and asterisks indicate a significant effect of the compost on each fraction and plant species (p < 0.05). The depletion / increase was calculated by subtraction of the amounts of metals in each fraction at bacterial treatment from the amount in the non-inoculated treatment. A significant depletion / increase in the metal concentration is denoted with an asterisk (p < 0.05) Nicotiana tabacum Untreated soil Compost-amended Salix caprea Untreated soil Compost-amended (b) 217 Evaluating compost and rhizobacterial inoculants for improving phytoextraction Similarly, strain P75 significantly increased the concentration of Cd in the F1 fraction (from 2.9 mg kg-1 to 3.4 mg kg-1) (Figure 6.4b). In the same soil, strain P75 significantly increased the oxidisable (F4) fraction of Zn by up to 100%. The acid-extractable (F2) Cd was the only fraction where a bacterial-induced reduction was found after inoculation with all the strains (although this effect was not significant). In untreated soils where S. caprea was grown, similar effects of strains P87 and P30 were found on soil Zn and Cd fractionation (Figure 6.4b). Both strains reduced the F2-F4 fractions, and this decrease was significant for both metals in the case of F3 (reducible phase). These metals were redistributed amongst the soluble (F1) and residual phases (F5). In the compost-amended soils, bacterial effects were less pronounced. In general, the bacterial inocula increased water soluble and exchangeable metal pools and decreased the acid-extractable, reducible and residual fractions (this was more pronounced for S. caprea where many of these alterations were statistically significant). Effects of rhizobacterial inoculants on shoot metal concentrations and metal yield of Nicotiana tabacum and Salix caprea Shoot metal concentrations of harvested tobacco are presented in Figure 6.5. In untreated soils, strains P29, P30 and P75 decreased the shoot metal concentration of the three major contaminating metals (Zn, Cd and Pb), and this effect was significant for Pb after inoculation with strains P30 and P75 (p < 0.05)). The effects of strains P87 and P64 were metal-dependent but not significant. For example, these inoculants increased shoot Zn concentrations, but they did not influence the shoot Cd concentration and slightly decreased shoot Pb concentration. In the compost-amended soils, bacterial inoculation did not significantly affect shoot Zn concentration, while strains P29 and P30 strains significantly decreased the Cd concentration (p < 0.05). Metal concentrations in S. caprea were determined for leaves and stems (Figure 6.6). Higher metal concentrations, and more pronounced bacterial-induced effects, were found in the leaves compared to the stems. Strain P87 significantly increased leaf Cd concentration (from 7.5 mg kg-1 to 13.1 mg kg-1SLQSODQWVJURZQLQXQWUHDWHGVRLOV2QWKH other hand, as observed with tobacco plants, strains P29 and P30 decreased both these elements in the leaves of S. caprea. In the compost-amended soil none of the bacterial strains led to an increase in leaf metal concentrations. Leaf Zn and Cd concentrations of NI plants were up to 5and 1.3-fold higher than stem 218 Chapter 6 Figure 6.5. Effects of bacterial inoculants on shoot metal concentration of N. tabacum growing in untreated and compost-amended soils. Asterisks indicate significant differences from the NI plants (p < 0.05). White bars show shoot Zn, Cd and Pb concentrations in the untreated soils and grey bars indicate metal concentrations in plants growing in the compost-amended soils. 219 Evaluating compost and rhizobacterial inoculants for improving phytoextraction concentrations in untreated soils, and up to 10and 2-fold higher in compostamended soils, respectively. In untreated soils, strains P87 and P75 increased the stem Zn concentration, and only one strain (P75) led to an increase in Cd concentration. In contrast to what was observed in leaves, strains P29 and P75 led to an increase in the stem Cd and Zn concentration in willows growing in compostamended soil (p < 0.05). Effects of bacterial inoculation on N. tabacum metal yields were strongly influenced by the soil treatment (Table 6.5). In untreated soil, metal yield was negatively affected by bacterial inoculation and even a significant reduction was found for Pb after inoculation with P87 and P64 compared to non-inoculated plants (by up to 50% and 42%, respectively (p < 0.05)). However, in plants growing in the compost-amended soils, all five inoculants improved metal yields and this was the case for all three metals (although not significant). Strain P87 had the highest impact on Zn yields (increasing from 2559 μg plant-1 to 2994 μg plant-1 after inoculation), P64 on Cd yield (increasing from 69.4 μg plant-1 to 77.8 μg plant-1) and P75 on Pb yield (increasing from 31.7 μg plant-1 to 58.9 μg plant-1). For S. caprea, the beneficial effects of inoculation on metal yields were more pronounced in the untreated soil: strains P87, P29 and P75 induced a slight Figure 6.6. Effects of bacterial inoculants on leaves and stem metal concentration of S. caprea growing in untreated and compost-amended soils. Asterisks indicate significant differences from the NI plants (p < 0.05) 220 Chapter 6 increase in the Zn yield compared to NI plants (from 4119 mg plant-1 in NI plants to 4429, 4429 and 4241 mg plant-1 UHVSHFWLYHO\ p > 0,05), while strain P30 increased the Cd yield (from 61.4 μg plant-1 to 65.2 μg plant-1 S !  However, in compost-amended soils bacterial inoculants generally reduced the Zn and Cd yield, with the exception of strains P29 and P75. 6.4. Discussion This study evaluated the benefits of two strategies, compost amendment and/or bioaugmentation, on the metal phytoextraction efficiency of N. tabacum and S. caprea. Tobacco grew poorly in the untreated soil, and this was most likely related to nutrient deficiency and metal phytotoxicity. The addition of compost had a strong promoting effect on tobacco growth, probably due to improvements in mine-soil fertility and structure and possibly a reduction in metal phytotoxicity. The increase in some of the evaluated indices related to the plant physiological Table 6.5. Effect of the five selected bacterial strains (P87, P29, P30, P64 and P75) compared to non-inoculated plants (NI) on the metal yield (μg) of N. tabacum and S. caprea grown in an untreated soil and compost amended soil. Asterisks indicate significant differences from the NI (p < 0.05) N. tabacum S. caprea Yield (μg) Yield (μg) Cd Zn Pb Cd Zn Non-amended soil NI 21.0 ± 1.2 1241 ± 102 11.6 ± 1.3 61.4 ± 12.4 4120 ± 451 P87 13.5 ± 0.9 961 ± 57 5.9 ± 0.7 * 56.1 ± 4.2 4429 ± 267 P29 13.6 ± 1.9 925 ± 120 8.4 ± 1.7 57.6 ± 6.0 4429 ± 138 P30 20.1 ± 2.8 1217 ± 142 8.3 ± 1.0 65.2 ± 12.4 3471 ± 286 P64 15.3 ± 2.0 1068 ± 125 6.7 ± 0.8 * 61.5 ± 2.4 3635 ± 360 P75 20.3 ± 2.6 1133 ± 92 8.0 ± 1.1 52.1 ± 5.2 4241 ± 312 Compost-amended soil NI 69.4 ± 3.0 2560 ± 192 31.7 ± 5.3 47.3 ± 6.9 8511 ± 1217 P87 74.1 ± 3.5 2994 ± 246 49.3 ± 8.4 28.6 ± 6.2 4454 ± 945 P29 74.2 ± 2.0 2960 ± 209 49.3 ± 10.2 54.1 ± 3.4 8121 ± 593 P30 66.7 ± 3.9 2750 ± 128 41.2 ± 5.7 31.0 ± 2.0 4951 ± 660 P64 77.8 ± 2.8 2840 ± 107 39.8 ± 6.9 42.2 ± 7.8 6665 ± 1528 P75 73.9 ± 6.5 2901 ± 226 58.9 ± 8.4 49.9 ± 8.5 6953 ± 1357 221 Evaluating compost and rhizobacterial inoculants for improving phytoextraction status confirmed this reduction in plant stress. A positive response in growth and biomass production of the same tobacco clone was also observed after the application of inorganic (mainly N) fertilisers in smelterand sewage sludgeaffected TE-contaminated agricultural soils (PHYTAC, 2005). Improvements in soil properties and promotion of plant biomass by compost addition are widely described (van Herwijnen et al., $OYDUHQJDet al., D$OYDUHQJDet al., E$OEXUTXHUTXHet al., 3DUUDJD-Aguado et al., 2015). The incorporation of these organic wastes into phytomanagement processes can contribute towards the recycling of residues, thus supporting current European policies aimed at reducing wastes (EU’s Zero Waste Policy, 2013) and may overcome the need for inorganic fertilisation. Our results suggest that, provided soil amendments (such as compost) are incorporated, tobacco could be cultivated in contaminated mine-soils. However, field and economic assessments would be required to determine its full potential for the phytomanagement of this type of site. The aerial biomass of S. caprea was almost 5-fold higher than tobacco plants in the untreated soil, and the compost addition only led to a slight but not significant increase in biomass production (although plants grown in compostamended soils generally showed higher photosynthetic efficiency). In contrast, French et al. (2006) found a large stimulation in the yield of several woody crops (including Salix) after amending a TE-contaminated landfill site with sewage. However, Salix caprea is well known to colonise soils with edaphically extreme properties, probably due to their low nutritional requirements (Dickinson et al., 1994). Several authors have reported a better survival rate and growth of S. caprea in contaminated soils compared to other Salix species (Dos Santos 8WPD]LDQ DQG :HQ]HO  'RV 6DQWRV 8WPD]LDQ et al., 2007b). Our results confirm the metal tolerance of this species and its potential for re-vegetation of mine-soils. Compost addition significantly decreased CEC and this reduction was possibly due to the occlusion of exchanges sites by added organic matter. Moreover, compost had a significant effect on soil metal availability and fractionation. Mineralisation of organic matter after its incorporation into the soil can frequently lead to a reduction in soil pH but this was not observed here. Alburquerque et al. (2011) also found no changes in the soil pH of neutral soils after compost addition. The effects of organic matter additions on metal availability and fractionation are generally described as being pH dependent. The formation of soluble organo-metal complexes both at high or low pH can increase 228 Chapter 6 Dos Santos Utmazian MN, Wenzel WW (2007) Cadmium and zinc accumulation in willow and poplar species grown on polluted soils. J Plant Nutr Soil Sci 170:265-272 Dos Santos Utmazian MN, Wieshammer G, Vega R, Wenzel WW (2007b) Hydroponic screening for metal resistance and accumulation of cadmium and zinc in twenty clones of willows and poplars. Environ Pollut 148:155-165 EU’s Zero Waste Policy (2013) Waste: A Resource to Recycle, Reuse and Recover Raw Materials Towards a near-zero waste society. (H2020-WASTE-2014/2015). French CJ, Dickinson NM, Putwain PD (2006) Woody biomass phytoremediation of contaminated brownfield land. Environ Pollut 141:387-395 Glick BR (2010) Using soil bacteria to facilitate phytoremediation. Biotechnol Adv 28:367374 Grandlic CJ, Mendez MO, Chorover J, Machado B, Maier RM (2008) Plant growthpromoting bacteria for phytostabilization of mine tailings. Environ Sci Technol 42:2079-2084 Gregson S, Alloway BJ (1984) Gel permeation chromatography studies on the speciation of lead in solutions of heavily polluted soils. J Soil Sci 35:55-61 Guadagnini M (2000) In vitro-breeding for metal-accumulation in two tobacco (Nicotiana tabacum) cultivars Innagural-Dissertation No 1288 der Mathematisch Naturwissenschftlichen Fakultät der Universität reiburg in der Schweiz: p. 109 Herzig R, Guadagnini M, Rehnert A, Erismann K (2003) Phytoextraction efficiency of in vitro-bred tobacco variants using a non-GMO approach. In: Vanek T, Schwitzguébel J (eds) Phytoremediation Inventory – COST Action 837 View., Prague (Czech Republic). 2003. UOCHB AVCR., p 73 Herzig R, Nehnevajova E, Pfistner C, Schwitzguebel JP, Ricci A, Keller C (2014) Feasibility of labile zn phytoextraction using enhanced tobacco and sunflower: Results of fiveand one-year field-scale experiments in Switzerland. Int J Phytorem 16:735-754 Isermann K (1983) Cd-content of the edible plant parts of different varieties of several crop species grown on highly Cd-contaminated neutral loam soil. Landwirtsch Forsch 36:283 Jahiruddin M, Livesey NT, Cresser MS (1985) Observations on the effect of soil pH upon zinc absorption by soils. Commun Soil Sci Plant Anal 16:909-922 Janssen J, Weyens N, Croes S, Beckers B, Meiresonne L, Van Peteghem P, Carleer R, Vangronsveld J (2015) Phytoremediation of metal contaminated soil using willow: Exploiting plant-associated bacteria to improve biomass production and metal uptake. Int J Phytorem 17:1123-1136 Kidd P Barceló J, Bernal MP, Navari-Izzo F, Poschenrieder C, Shilev S, Clemente R, Monterroso C (2009) Trace element behaviour at the root–soil interface: Implications in phytoremediation. Environ Exp Bot 67:243-259 Kidd P Mench M, Álvarez-López V, Bert V, Dimitriou I, Friesl-Hanl W, Herzig R, Janssen JO, Kolbas A, Müller I, Neu S, Renella G, Ruttens A, Vangronsveld J, Puschenreiter M (2015) Agronomic practices for improving gentle remediation of trace element-contaminated soils. Int J Phytorem 17:1005-1037 Kiikkilä O, Pennanen T, Perkiömäki J, Derome J, Fritze H (2002) Organic material as a copper immobilising agent: a microcosm study on remediation. Basic Appl Ecol 3:245-253 Kuffner M, Puschenreiter M, Wieshammer G, Gorfer M, Sessitsch A (2008) Rhizosphere bacteria affect growth and metal uptake of heavy metal accumulating willows. Plant Soil 304:35-44 229 Evaluating compost and rhizobacterial inoculants for improving phytoextraction Kumpiene J, Fitts JP, Mench M (2012) Arsenic fractionation in mine spoils 10 years after aided phytostabilization. Environ Pollut 166:82-88 Kumpiene J, Lagerkvist A, Maurice C (2008) Stabilization of As, Cr, Cu, Pb and Zn in soil using amendments--a review. Waste Manag 28:215-225 Kumpiene J, Mench M, Bes CM, Fitts JP (2011) Assessment of aided phytostabilization of copper-contaminated soil by X-ray absorption spectroscopy and chemical extractions. Environ Pollut 159:1536-1542 Langella F, Grawunder A, Stark R, Weist A, Merten D, Haferburg G, Büchel G, Kothe E (2014) Microbially assisted phytoremediation approaches for two multi-element contaminated sites. Environ Sci Pollut Res 21:6845-6858 Lichtenthaler HK, Gitelson A, Lang M (1996) Non-destructive determination of chlorophyll content of leaves of a green and an aurea mutant of tobacco by reflectance measurements. J Plant Physiol 148:483-493 Lyubenova L, Nehnevajova E, Herzig R, Schroder P (2009) Response of antioxidant enzymes in Nicotiana tabacum clones during phytoextraction of heavy metals. Environ Sci Pollut Res Int 16:573-581 Martínez-Fernández D, Arco-Lázaro E, Bernal MP, Clemente R (2014) Comparison of compost and humic fertiliser effects on growth and trace elements accumulation of native plant species in a mine soil phytorestoration experiment. Ecological Engineering 73:588-597 Mench M, Lepp N, Bert V, Schwitzguébel J-P, Gawronski S, Schröder P, Vangronsveld J (2010) Successes and limitations of phytotechnologies at field scale: outcomes, assessment and outlook from COST Action 859. J Soils Sediments 10:1039-1070 Mench M, Vangronsveld J, Clijsters H, Lepp N, Edwards R (2000) In situ metal immobilization and phytostabilization of contaminated soils. In: Terry N, Bañuelos G (eds) Phytoremediation of contaminated soil and water. Lewis Publishers, Florida, USA, pp 323-362 Mergeay M, Nies D, Schlegel H, Gerits J, Charles P, Van Gijsegem F (1985) Alcaligenes eutrophus CH34 is a facultative chemolithotroph with plasmid-bound resistance to heavy metals. J Bacteriol 162:328-334 Monterroso C, Rodríguez F, Chaves R, Diez J, Becerra-Castro C, Kidd PS, Macías F (2014) Heavy metal distribution in mine-soils and plants growing in a Pb/Znmining area in NW Spain. Appl Geochem 44:3-11 Panagos P, Van Liedekerke M, Yigini Y, Montanarella L (2013) Contaminated sites in Europe: Review of the current situation based on data collected through a European network. J Environ Public Health 2013 Parraga-Aguado I, González-Alcaraz MN, Schulin R, Conesa HM (2015) The potential use of Piptatherum miliaceum for the phytomanagement of mine tailings in semiarid areas: Role of soil fertility and plant competition. J Environ Manage 158:74-84 Peñuelas J, Filella I, Gamon JA (1995) Assessment of photosynthetic radiation-use efficiency with spectral reflectance. New Phytol 131:291-296 PHYTAC (2005) Development of systems to improve phytoremediation of metal contaminated soils through improved phtoaccumulation. Pulford ID, Watson C (2003) Phytoremediation of heavy metal-contaminated land by trees: a review. Environ Int 29:529-540 Rauret G, Lopez-Sanchez JF, Sahuquillo A, Barahona E, Lachica M, Ure AM, Davidson CM, Gomez A, Luck D, Bacon J, Yli-Halla M, Muntau H, Quevauviller PH, (2000) Application of a modified BCR sequential extraction (three-step) 230 Chapter 6 procedure for the determination of extractable trace metal contents in a sewage sludge amended soil reference material (CRM 483), complemented by a threeyear stability study of acetic acid and EDTA extractable metal content. J Environ Monit 2:228-233 Richardson AD, Duigan SP, Berlyn GP (2002) An evaluation of noninvasive methods to estimate foliar chlorophyll content. New Phytol 153:185-194 Riddell-Black D, Pulford I, Stewart C (1997) Clonal variation in heavy metal uptake by willow. Aspects of Applied Biology 49:327-334 Schreiber U, Bilger W, Hormann H, Neubauer C (1998) Chlorophyll fluorescence as a diagnostic tool: basics and some aspects of practical relevance. In: Raghavendra AS (ed) Photosynthesis: a comprehensive treatise. Cambridge, United Kingdom, pp 320-336 Sessitsch A, Kuffner M, Kidd P, Vangronsveld J, Wenzel WW, Fallmann K, Puschenreiter M (2013) The role of plant-associated bacteria in the mobilization and phytoextraction of trace elements in contaminated soils. Soil Biol Biochem 60:182 -194 Shuman LM (1999) Organic waste amendments effect on zinc fractions of two soils. J Environ Qua 28:1442-1447 Sims DA, Gamon JA (2002) Relationships between leaf pigment content and spectral reflectance across a wide range of species, leaf structures and developmental stages. Remote Sensing of Environment 81:337-354 Tandy S, Healey JR, Nason MA, Williamson JC, Jones DL (2009) Remediation of metal polluted mine soil with compost: Co-composting versus incorporation. Environ Pollut 157:690-697 Unterbrunner R, Puschenreiter M, Sommer P, Wieshammer G, Tlustos P, Zupan M, Wenzel WW (2007) Heavy metal accumulation in trees growing on contaminated sites in Central Europe. Environ Pollut 148:107-114 van Herwijnen R, Laverye T, Poole J, Hodson ME, Hutchings TR (2007) The effect of organic materials on the mobility and toxicity of metals in contaminated soils. Appl Geochem 22:2422-2434 Vangronsveld J, Herzig R, Weyens N, Boulet J, Adriaensen K, Ruttens A, Thewys T, Vassilev A, Meers E, Nehnevajova E, van der Lelie D, Mench M (2009) Phytoremediation of contaminated soils and groundwater: lessons from the field. Environ Sci Pollut Res Int 16:765-794 Vyslouzilova M, Tlustos P, Száková J (2003) Cadmium and zinc phytoextraction potential of seven clones of Salix spp. planted on heavy metal contaminated soils. Plant Soil Environ 49:542-547 Wieshammer G, Unterbrunner R, García TB, Zivkovic MF, Puschenreiter M, Wenzel WW (2007) Phytoextraction of Cd and Zn from agricultural soils by Salix ssp. and intercropping of Salix caprea and Arabidopsis halleri. Plant Soil 298:255-264 Wong MH (2003) Ecological restoration of mine degraded soils, with emphasis on metal contaminated soils. Chemosphere 50:775-780 $%675$&7 In this study different bacterial inoculation methods were tested for tobacco plants growing in a mine-soil contaminated with Pb, Zn and Cd. The inoculation methods evaluated were: seed inoculation, soil inoculation, dual soil inoculation event and seed+soil inoculation. Each inoculum was added at two bacterial densities (10 6 CFUs mL -1 and 10 8 CFUs mL -1 ). The objectives were to evaluate whether or not the mode of inoculation or the number of applied microorganisms influences plant response. The most pronounced bacterialinduced effect was found for biomass production, and the soil inoculation treatment (using 10 6 CFUs mL -1 ) led to the highest increase in shoot dry weight yield (up to 45%). Bacterial-induced effects on shoot metal concentrations were OHVVSURQRXQFHGDOWKRXJKDSRVLWLYHHIIHFWZDVIRXQGRQVKRRW3EFRQFHQWUDWLRQ when using 10 8 CFUs mL -1 in the soil inoculation (29% increase) and in the seed+soil inoculation (34% increase). Also shoot Zn concentration increased by 24% after seed inoculation with 10 6 CFUs mL -1 . The best effects on the total metal yield were not correlated with an increasing number of inoculated bacteria. In fact the best results were found after a single soil inoculation using the lower cellular density of 10 6 CFUs mL -1 . &KDSWHU ,QRFXODWLRQ PHWKRGV XVLQJ 5KRGRFRFFXV HU\WKURSROLV  VWUDLQ 3 DIIHFWV EDFWHULDO DVVLWHG SK\WRH[WUDFWLRQ FDSDFLW\ RI 1LFRWLDQDWDEDFXP This study formed part of the following publication: Álvarez-López, V., Prieto-Fernández, A., Janssen, J., Herzig, R., Vangronsveld, J., Kidd, P.S ;2016). Inoculation methods using Rhodococcus erythropolis strain P30 affects bacterial assisted phytoextraction capacity of Nicotiana tabacum.Int J Phytorem 18:406-415. 235 Inoculation method influences plant responses of Nicotiana tabacum 7.1. Introduction Phytomanagement of trace metal-contaminated soils describes a series of plant-based remediation technologies which aim to reduce associated environmental risks and restore soil functions and quality (Mench et al.5RELQVRQ et al., 9DQJURQVYHOG et al., 2009). Phytoextraction is based on the use of plants that accumulate large amounts of metals into their above-ground biomass which can then be harvested. Metal removal depends on the metal concentration in harvestable plant parts and biomass production (Kidd et al.5RELQVRQ et al.,  9DQJURQVYHOG et al., 2009). High-biomass crops and woody plants are considered good candidates for phytoextraction of trace metals (particularly Cd, Se and Zn) if they also show relevant shoot metal removals and a high shoot yield. Moreover, metal extraction can be combined with bioenergy production (Schroder et al.1VDQJDQZLPDQD et al..LGG et al., 2015). Phytoextraction efficiency can be limited by reduced plant biomass production or low soil metal bioavailability (Baker et al.:HQ]HO6HYHUDO strategies have been proposed to optimise the process, such as using improved plant cultivars, appropriate agronomic techniques, or beneficial plant-associated microbial inoculants (Nehnevajova et al.9DQJURQVYHOG et al.*OLFN .LGG et al., 2015). Microbial-assisted phytoextraction uses native microorganisms which enhance plant growth and/or alter soil metal availability (Glick,  0D et al.  -DQVVHQ et al., 2015). Numerous authors have isolated bacterial strains from metal-rich or metal-contaminated areas, screened these for their plant growth promoting (PGP) traits and metal resistance, and used them as plant inoculants (Haslmayr et al., 2014). In many cases a bacterial-induced enhancement in plant biomass production and hence metal removal has been shown, and several strains have been reported to be useful for phytoextraction (see review by Sessitsch et al. (2013)). Phytoextraction efficiency is enhanced through a bacterial-induced increase in soil metal mobility (Sessitsch et al., 2013) or by LQFUHDVLQJ SODQW JURZWK SODQW JURZWK SURPRWLQJ EDFWHULD 3*3% *OLFN  Visioli et al., 2015). It is generally accepted that the experimental conditions, in particular the host plant species, soil type and properties, and nature and type of contamination (monoor poly-PHWDOOLF FRQWDPLQDWLRQ FR-contamination with RUJDQLFDQGLQRUJDQLFSROOXWDQWVSROOXWDQWFRQFHQWUDWLRQDQWKURSRJHQLFJHRJHQLF or spiked contamination), are important factors determining whether or not a beneficial bacterial-induced effect is observed (Weyens et al.%HFHUUD-Castro et Chapter 7 236 al.  6HVVLWVFK et al.  &DEHOOR-Conejo et al., 2014). However, few studies have evaluated the effects of the inoculation method on plant growth and metal accumulation. Inoculation methods differ greatly between studies (inoculation of seed/plant/soil, number/timing of inoculation events, bacterial cellular densities) and these aspects are likely to influence the final effects of such plant-microbial associations. The most common methods used are seed and soil inoculation, while plant inoculation and experiments including more than one inoculation event are rare (Abou-Shanab et al.,  'HOO¶$PLFR &DYDOFD DQG $QGUHRQL  %HFHUUD-Castro et al.,  &DEHOOR-Conejo et al., 2014). Bacterial densities of the inoculum generally fluctuate between 107108 CFUs mL-1, however the total density of inoculated bacteria per pot (CFUs per pot) is highly variable due to the different amounts of soil used (Kuffner et al.  Marques et al.3HWULFFLRQH et al./DQJHOOD et al., 2014). According to Malusá et al. (2012), the commercial use of microbial inoculants has been limited due to the inconsistency of the results obtained. There is therefore a need to test the effects of different inoculation methods on plant growth promotion and metal removal under standardised conditions. The objectives of this study were to evaluate if the effects of a bacterial strain with PGP traits (Rhodococcus erythropolis P30) on metal phytoextraction capacity of Nicotiana tabacum are dependent on the inoculation method. For this, we assessed the effects on plant biomass production and metal accumulation of (i) seed versus soil inoculation, (ii) inocula with different bacterial densities: 106 versus 108 CFUs mL-1, and (iii) the number of inoculation events (single versus repeated inoculation). 7.2. Materials and Methods Soil and plant species selected for the pot experiment Mine soil (< 20 cm depth) was collected from the tailings of the abandoned Pb-Zn mine of Rubiais in Lugo, NW Spain (UTM 29T 660781/4726800). The physicochemical characteristics of the tailings have been described in detail previously (Becerra-Castro et al.0RQWHUURVR et al., 2014). The mine soil was air-dried and sieved to < 8 mm and mixed with 5% compost (w/w). Compost addition of 5% (w/w) was necessary to achieve optimal growth of tobacco (Álvarez-López et al., 2011). The compost was based on municipal sewage sludge and wood chippings. The compost presented a slightly acid pH (6.2), high C/N 237 Inoculation method influences plant responses of Nicotiana tabacum ratio (13.5), high CEC (57.6 cmolc kg-1), and a low to moderate concentration of trace metals. The mine-soil had a slightly alkaline pH (8.23), low cationic exchangeable capacity (18.0 cmolc kg-1) dominated by Ca, low P availability (25.9 mg kg-1) and high concentrations of total Cd, Zn and Pb (7, 5328 and 893 mg kg-1, respectively). The NH4Cl-extractable Cd and Zn concentrations were 1.84 and 201 mg kg-1, respectively. The NH4NO3-extractable Cd and Zn concentrations were 0.23 and 33.7 mg kg-1, respectively. Pb concentrations were below the quantification limit in both NH4NO3and NH4Cl-extractable fractions (0.15 and 2 mg kg-1, respectively). Tobacco can be used for metal phytoextraction and at the same time can provide biomass for bioenergy, green fine chemistry, and other ecotechnologies (Schroder et al., 2008). In this study, the NBCu 10-8 variant of Nicotiana tabacum (F2 generation) was used. This variant was selected from 106 somaclonal variants of tobacco by in vitro breeding techniques and showed an improved Cd DQG =Q H[WUDFWLRQ FDSDFLW\ FRPSDUHG WR WKH PRWKHU OLQH *XDGDJQLQL  Herzig et al., 2014). Plants were grown in 2.5 L pots (with 2 kg of soil). Rhizobacterial strain and inoculation treatments The strain P30 (Rhodococcus erythropolis $FFHVVLRQ QXPEHU +( available at www.ebi.ac.uk) was previously isolated from the rhizosphere soil of Festuca rubra which is a natural coloniser of the tailings in the Rubiais mine (Becerra-Castro et al., 2012). It is Cdand Zn-tolerant, and both an organic acidand indoleacetic acid (IAA)-producer (5 mg IAA mL-1). Metal resistance was determined using a minimal salt medium (with sugars and organic acids as C sources) enriched with Cd (added as CdSO4.8/3H2O from 0 to 6 mM) or Zn (added as Zn(SO4)2·7H2O from 0 to 25 mM). The Maximal Tolerable Concentration (MTC) was 4 mM Cd and 2.5 mM Zn. Organic acid production was assessed using the qualitative bromocresol method. The ability produce IAA were evaluated in liquid medium supplemented with tryptophan, and following the Salkowski method. The P30 strain did not produce siderophores and was not able to solubilise inorganic phosphate. A full description of the methods used for the phenotypic characterisation is given by Becerra-Castro et al. (2011). It has been shown to mobilise Zn from metal-rich minerals and soils under in vitro conditions (unpublished data). The strain has been previously characterised as a plant growth promoting bacteria (PGPB) due to its beneficial effect on the biomass production of a range of plant species, including Festuca sp., Salix spp. and N. tabacum, Chapter 7 244 The z scores calculated for each nutrient indicated that the shoot element concentrations grouped differently according to treatments (Figure 7.3). Although not consistent, a general positive effect on shoot nutrient concentrations was found in those treatments in which soil was inoculated once (soil and seed+soil treatments). However, the addition of MgSO4 (non-inoculated treatments) also had an effect on the shoot metal concentrations: the respective controls of the seed and dual soil inoculation treatments presented the highest concentrations of K and Fe or of Ca, K and P in tobacco shoots, respectively. The inoculation mode, but not the inoculum density, significantly influenced shoot metal concentration (p < 0.05) compared to corresponding NI SODQWV S   7DEOH  DQG )LJXUH  6KRRW 3E FRQFHQWUDWLRQ LQFUHDVHG (albeit not significantly) compared to non-inoculated plants after addition of 106 CFUs mL-1 in the seed, soil, and seed+soil inoculation modes (by 13%, 15% and 17%, respectively). Addition of 108 CFUs mL-1, led to a significant increase in the shoot Pb concentration (p < 0.05) in the soil inoculation (from 12.6 mg kg-1 in NI plants to 16.2 mg kg-1 in inoculated plants) and seed+soil inoculation modes (from 15.6 mg kg-1 in NI plants to 20.9 mg kg-1 in inoculated plants). In contrast to what was observed for biomass, successive inoculations (seed+soil inoculation) increased the mean shoot Pb concentration and this was observed with both cellular densities. In the case of the shoot Zn concentration, a significant increase (p < Figure 7.3. Mean values of z-scores performed in relation to concentrations of Ca, Fe, K, Mg, and P in shoots of N. tabacum growing in the different treatments. The z score value is given by z = (x - μ) / ı, ZKHUH[LVWKHHOHPHQWFRQFHQWUDWLRQRIHDFKVDPSOHDQGı are respectively the mean and the standard deviation of each element. 245 Inoculation method influences plant responses of Nicotiana tabacum 0.05) was observed after seed inoculation with the lower cellular density (from 522.1 mg kg-1 in NI to 645.1 mg kg-1 in inoculated plants) and also a tendency to increase shoot Zn concentration was found with the higher cellular density (by 10%). A slight but not significant increase in shoot Zn concentration was observed after inoculation with the bacterial density of 108 CFUs mL-1 for the soil inoculation mode (7%) and after inoculation with the bacterial density of 106 CFUs mL-1 for the seed+soil inoculation mode (6%). No significant differences in shoot Cd concentration were observed, although a tendency towards a higher concentration was observed with both bacterial densities after seed (10% and 9%) and seed + plant inoculation (8% and 7%), and with the density of 108 CFUs mL-1 after soil inoculation (13%). Effect of bacterial inoculation on metal availability in soils The effect of plant growth and bacterial inoculation on soil labile metal pools were evaluated using NH4NO3 and NH4Cl extractants (Figure 7.4). After 16 weeks of growth the labile (NH4Cland NH4NO3-extractable) Zn pool was reduced by up to 67% and 51% (in non-inoculated N. tabacum) which was likely due to plant uptake and accumulation (Figure 7.4). NI plants also decreased available Cd by up to 58% and 74%, (NH4Cland NH4NO3-extractable, respectively) (Figure 7.4). Inoculation treatments also affected the soil metal availability (albeit not always significantly). Moreover, the same effects on metal availability Figure 7.4. Effect of the different inoculation treatments on the metal availability (mean ± SE) assessed using NH 4 NO 3 and NH 4 Cl soil extractions. Different letters above bars indicate significant differences (p < 0.05) within each treatment (NI, 10 6 CFUs mL -1 and 10 8 CFUs mL -1 ) (n=6). Capital letters within bars indicate differences between NI treatments (p < 0.05). A A A B Chapter 7 246 were not always observed with both soil extractants, and the main divergences between the two extractants were observed in the seed+soil treatment. The concentrations of NH4NO3-extractable Zn and Cd in pots after seed inoculation with a bacterial density of 106CFUs mL-1, were significantly higher than in NI plants (from 16.5 to 18.3 mg Zn kg-1 and from 0.06 to 0.08 mg Cd kg-1 UHVSHFWLYHO\S 0.05). Similarly, a higher Zn and Cd availability measured with NH4NO3 (only significant for Zn) was found after inoculation with the higher bacterial density (from 16.5 to 17.6 mg Zn kg-1S7KHVLQJOHDQGGXDOVRLOLQRFXODWLRQZLWK the higher bacterial density of 108CFUs mL-1 tended to increase NH4NO3extractable Zn and Cd but the difference detected was only significant for Cd after dual soil inoculation (0.07 mg kg-1 to 0.10 mg kg-1, p < 0.05). Similar tendencies were observed using the NH4Cl extractant but the differences were not significant. Conversely, in the case of seed+soil inoculations, changes in Zn or Cd availability were not detected with NH4NO3EXWWKH1+4Cl-extractable Zn and Cd concentrations were significantly higher in pots receiving inocula of 106 CFUs mL-1 (113.9 mg Zn kg-1 and 1.53 mg Cd kg-1) than in NIpots (73.8 mg Zn kg-1 and 1.07 mg Cd kg-1). Effect of bacterial inoculation on plant metal yield The efficiency and success of metal phytoextraction depends on the total amount of extracted metal (or metal yield) which is dependent on the total amount of biomass produced and the accumulated metal in harvestable tissues. Figure 7.5 presents the percentage increase or decrease in plant metal yields for the four LQRFXODWLRQPRGHVDQGFHOOXODUGHQVLWLHVLQUHODWLRQWR1,SODQWVPHWDO\LHOGZDV only significantly infuenced by the inoculation mode (Table 7.1) (p < 0.05). ,QWHUDFWLRQHIIHFWVGLGQRWUHDFKVWDWLVWLFDOVLJQLILFDQFHIRUPHWDO7DEOHS! 0.05). Non-inoculated plants extracted between 0.18 to 0.30 mg Pb plant-1, 7.08 to 9.18 mg Zn plant-1 and 0.43 to 0.57 mg Cd plant-1. In general, all the treatments led to an improvement in the phytoextraction of the evaluated metals and the most pronounced effects were found for Pb. However, the single soil inoculation mode was the only case where a significant increase in the yield of all three metals was found using both celullar densities. The best results in this inoculation mode were achieved using the lower bacterial density, which caused increases in Pb, Zn and &G\LHOGVRIXSWRDQGUHVSHFWLYHO\S)LJXUH)LQDOO\ the dual soil inoculation using 108 CFUs mL-1 tended to reduce the amount of 247 Inoculation method influences plant responses of Nicotiana tabacum Figure 7.5. Percentage of increase / depletion in total metal extracted (metal yield) relative to the NI treatment (n = 6). The depletion / increase was calculated by substracting the amount of accumulated metals in the NI plants from the corresponding amounts in each inoculation treatment (mean ± SE). A significant depletion / increase is denoted by an asterisk (p < 0.05). Chapter 7 248 extracted metal for all three elements (Pb, Zn and Cd), and Cd extraction was also reduced in this treatment when using 106 CFUs mL-1. 7.4. Discussion The PGP effect of strain P30 was found to be dependent on the inoculation method and cellular density of the inoculum. The selected inoculation methods used (seed and soil inoculation) represent the most common practices used for the application of microbes to soil (Afzal et al., 2011). Seed inoculation has been integrated within phytoremediation techniques targeting organic- (Germaine et al.,  *XUVND et al., 2009) or metal-contaminated sites (Kolbas et al., 2015). Bashan (1986) tested the effect of inoculation time (seed and soil inoculation) on bacterial induced effects on plant growth, and found that seed inoculation showed the best results in terms of increasing shoot DW. In contrast, in this study, soil inoculation (inoculation after plant emergence) achieved the best results in terms of plant growth promotion. Similarly, Afzal et al. (2011) found that bacterial strains applied by soil inoculation significantly enhanced hydrocarbon degradation and rhizosphere colonisation to a higher degree than when the strains were applied to seeds. These authors observed a poorer colonisation after seed inoculation and attributed this to the differences in the exudates released which favoured the growth of other microorganisms already established on the seed coat. It has been shown that a poor microbial survival in the rhizosphere can led to a reduction in the desirable positive effects (Gilbertson et al.*XQGHUVRQ et al., 2007). Both bacterial densities used for the inoculations in this study are within the ranges of those commonly used in the literature (Kuffner et al.0DUTXHV et al.3HWULFFLRQH et al./DQJHOOD et al., 2014). Increasing the number of inoculated bacterial cells (from 106 CFUs mL-1 to 108 CFUs mL-1) did not lead to a further improvement in biomass. Moreover, increasing the number of inoculation events could be considered a crucial factor influencing the success of establishment and colonisation of the desired bacterial strain in the rhizosphere, but this aspect has rarely been studied in phytoremediation-orientated studies. Here we did not find an improvement in biomass production after successive bacterial inoculation events (dual soil or seed+soil inoculations) and none of these increased plant growth by more than 10%. Bashan (1986) also showed that multiple inoculations have a marginal effect on plant growth, and only after four successive inoculations was an increase in plant response observed. These authors attributed this effect to poor root colonisation. The less pronounced effect after a second soil 249 Inoculation method influences plant responses of Nicotiana tabacum inoculation could be related to the presence of a more established bacterial community in the rhizosphere of tobacco plants which impedes the possibilities of the inoculum at the later plant growth stage to colonise efficiently. Whereas at the first inoculation event newly transplanted and rapidly growing plants allow for the inoculum to establish and proliferate more effectively. Tobacco plants have been previously described as efficient metal phytoextractors (Fässler et al.*XDGDJQLQL,QWKHSUHVHQWVWXG\WKHVKRRW Cd and Zn concentrations were close to the upper threshold values of the phytotoxic concentrations for these elements (5-30 mg kg-1 &G-400 mg kg-1 Zn) as proposed by Kabata-Pendias (2010). On the other hand, shoot Pb concentrations were well below the 30-300 mg kg-1 phytotoxic concentration threshold given for this element. A similar range in shoot concentrations of Cd (2.5-25 mg kg-1), Zn (150-850 mg kg-1) and Pb (15-30 mg kg-1) were observed in tobacco variants assessed in experimental metal-contaminated field sites in Balen (Belgium) and Rafz (Switzerland) (PHYTAC, 2005). Both these sites are agricultural soils which were affected by fall-out from a Zn/Pb smelter (Balen) or by sewage sludge application (Rafz). The present study suggests that tobacco is also a potential candidate for cultivation in contaminated mine-soils, although field assessments would be required to determine its full potential. The efficiency and success of metal phytoextraction depends on the total amount of extracted metal (or metal yield) which is dependent on the total amount of biomass produced and the accumulated metal in harvestable tissues. Depending on the inoculation method and cellular density, strain P30 induced an increase in the biomass production of N. tabacum. This strain is an IAA-producer and the beneficial effects of this phytohormone on plant growth are well known 'HOO¶$PLFRHWDO'XFD et al., 2014). Inoculation with strain P30 resulted in a more successful growth promotion (up to 45% increase) than metal accumulation enhancement (up to 34% increase) compared to non-inoculated plants. This result coincides with other studies in which the bacterial inoculants used tended to be more successful in promoting plant growth, rather than increasing shoot metal concentration. For example, Janssen et al. (2015) observed an increase in Zn/Cd phytoextraction efficiency of S. viminalis by increasing twigs growth after inoculation with Rahnella sp. an IAA-, ACC deaminase-, organic acidsand siderophoreproducer. Lampis et al. (2015) also found than a bacterial inoculum combination of siderophoreand IAAproducers (Bacillus sp. MPV12, Variovorax sp. P4III4, and Pseudoxanthomonas sp. P4V6) increased biomass of Chapter 7 250 Pteris vitatta by up to 30% increasing total As removal. Kuffner et al. (2010) observed an increase in growth of willows after inoculation with the strain Agromyces AR33. Sessitsch et al. (2013) carried out a meta-analysis of phytoremediation-orientated inoculation studies (in which over 70 publications and 738 individual cases or treatments were included) and found that the most frequent effects observed of plant inoculation were an increase in shoot biomass (30% of cases). Although there was no significant increase in the shoot Cd concentration of N. tabacum after soil inoculation with strain P30, the increase in shoot DW biomass led to a significant increase in the Cd removal. This strain could therefore be a potential candidate for improving the metal extraction capacity of phytoextracting plants. Similarly, Becerra-Castro et al. (2012) observed that rhizosphere bacterial strains did not affect Cd and Zn accumulation in the leaves of Salix viminalis but the microbial-induced increase in biomass production resulted in an increase in the amount of trace elements phytoextracted. In the present study, positive effects of bacterial inoculation on metal uptake by tobacco were also observed (especially Pb). An increase in metal uptake by tobacco after bacterial inoculation was also reported by Mastretta et al. (2009), who found that the inoculation of Nicotiana tabacum seeds with the endophyte Sanguibacter sp. Sd2 increased shoot Cd concentrations compared to non-inoculated plants. Ma et al. (2015) also found an increase in Cd and Zn uptake by Sedum plumbizicola after inoculation with B. pumilus E2S2 and Bacillus sp E1S2 and Sangthong et al. (2015) observed that bacterial inoculation enhanced Cd shoot concentration in maize plants. After 12 weeks of growth the labile (NH4Cl-extractable) Zn pool was reduced by up to 33% (in non-inoculated N. tabacum) which was likely due to plant uptake and accumulation. Likewise, Herzig et al. (2014) found a decrease in the labile (NaNO3-extractable) Zn pool of up to 69% in the soils at Rafz planted with the same variant (NBCu 10-8, 1st generation). Overall, no significant correlation was found between the abundance of inoculated bacteria and shoot DW, metal concentration or removal. These results are in agreement with other studies in which the authors evaluated the effects of LQFUHDVLQJWKHEDFWHULDOGHQVLW\RILQRFXOD%DVKDQ5DL et al.%UDXG et al.5DPtUH]DQG.ORHSSHU6LPLODUILQGLQJVKDYHEHHQREWDLQHG when using mycorrhizal inocula, since results indicate that plant growth is not correlated with the rate of mycorrhizal colonisation (Baum et al., 2006). Our results therefore indicate that the activity of the inoculated bacterial strain in the soil could be more or equally as important as the cellular density of the inoculum. 251 Inoculation method influences plant responses of Nicotiana tabacum Schwartz et al. (2000) attributed the more beneficial effects of a lower bacterial density in plant growth promotion to the depletion of limited soil resources necessary for growth and bacterial survival when a higher density is inoculated. In our case, bacterial inoculation using the lower cellular density (106 CFUs mL-1) actually had a more pronounced growth promoting effect than the inoculum with a higher cellular density (108CFUs mL-1). A high dose of inoculum might also have a deleterious effect on other soil microorganisms which are beneficial for tobacco growth. Similar results were found by Kapulnik et al. (1985) who showed that plant response to high levels of Azospirillum was negative and that the optimal bacterial level of inoculation in hydroponic systems was 105 - 106 CFUs mL-1. 7.5. Conclusions Several studies can be found on the development of PGPB inoculation methods (principally for N2-fixing bacteria) for agricultural food crops. In contrast, the optimisation of inoculation methods for plants growing in contaminated soils has rarely been assessed, and even less so in phytoextractionorientated studies. In general, inoculation with the IAA-producing bacterial P30 strain positively affected the phytoextraction process of the three problematic metals (Pb, Zn and Cd) in the mine soil. The most pronounced effect was observed in plant biomass production, and to a lesser extent in the shoot metal accumulation and plant nutritive status. Our results indicated that both the inoculation mode and the cellular density of the inoculum are important factors to take into account when designing inoculation strategies, since they can modify the obtained results in terms of plant performance and soil metal removal. In contrast to what was expected, a single soil inoculation event was found to give the best results together with the lower bacterial density (106 CFUs mL-1). Further studies at a field scale will be required to corroborate the beneficial effects of the best inoculation method selected in this study. 7.6. References Abou-Shanab R, Angle J, Chaney R (2006) Bacterial inoculants affecting nickel uptake by Alyssum murale from low, moderate and high Ni soils. Soil Biol Biochem 38: 2882-2889 Afzal M, Yousaf S, Reichenauer TG, Sessitsch A (2011) The inoculation method affects colonization and performance of bacterial inoculant strains in the phytoremediation of soil contaminated with diesel oil. Int J Phytorem 14: 35-47 Álvarez-López V, Prieto-Fernández A, Becerra-Castro C, Rodríguez-Garrido B, Herzig R, Puschenreiter M, Kidd P (2011) Improving phytoextraction efficiency of high- Chapter 7 252 biomass crops using microbial inoculants and organic amendments. Paper presented at the 9th International Phytotechnology Society (IPS). +DVVHOW%HOJLXP 6HSWHPEHU Baker A, McGrath S, Reeves R, Smith J (2000) Metal hyperaccumulator plants: A review of the ecology and physiology of a biological resource for phytoremediation of metal-polluted soils. In: Terry N, Bañuelos G, eds. Phytoremediation of contaminated soil and water. Florida, USA. Lewis Publishers, Boca Raton. p. 85– 107 Bashan Y (1986) Significance of timing and level of inoculation with rhizosphere bacteria on wheat plants. Soil Biol Biochem 18: 297-301 Baum C, Hrynkiewicz K, Leinweber P, Meißner R (2006) Heavy-metal mobilization and uptake by mycorrhizal and non-mycorrhizal willows (Salix × dasyclados). J Plant Nutr Soil Sci 169: 516-522 Becerra-Castro C, Prieto-Fernández Á, Álvarez-López V, Cabello-Conejo MI, Acea MJ, Kidd PS, Monterroso C (2011) Nickel solubilizing capacity and characterization of rhizobacteria isolated from hyperaccumulating and non-hyperaccumulating subspecies of Alyssum serpyllifolium. Int J Phytorem 13: 229–244 Becerra-Castro C, Monterroso C, Prieto-Fernández A, Rodríguez-Lamas L, LoureiroViñas M, Acea MJ, Kidd PS (2012) Pseudometallophytes colonising Pb/Zn mine tailings: a description of the plant-microorganism-rhizosphere soil system and isolation of metal-tolerant bacteria. J Hazard Mater 217-218: 350-359 Belimov AA, Kunakova AM, Safronova VI, Stepanok VV, Yudkin LY, Alekseev YV, Kozhemyakov AP (2004) Employment of rhizobacteria for the inoculation of barley plants cultivated in soil contaminated with lead and cadmium. Microbiology 73: 99-106 Braud A, Jézéquel K, Vieille E, Tritter A, Lebeau T (2006) Changes in extractability of Cr and Pb in a polycontaminated soil after bioaugmentation with microbial producers of biosurfactants, organic acids and siderophores. Water Air Soil Pollut: Focus 6: 261-279 Cabello-Conejo M I, Becerra-Castro C, Prieto-Fernández A, Monterroso C, SaavedraFerro A, Mench M, Kidd PS (2014) Rhizobacterial inoculants can improve nickel phytoextraction by the hyperaccumulator Alyssum pintodasilvae. Plant Soil 379: 35-50 Dell’Amico E, Cavalca L, Andreoni V (2008) Improvement of Brassica napus growth under cadmium stress by cadmium-resistant rhizobacteria. Soil Biol Biochem 40: 74-84. Duca D, Lorv J, Patten CL, Rose D, Glick BR (2014) Indole-3-acetic acid in plantmicrobe interactions. Antonie van Leeuwenhoek 106: 85-125 Fässler E, Robinson BH, Stauffer W, Gupta SK, Papritz A, Schulin R (2010) Phytomanagement of metal-contaminated agricultural land using sunflower, maize and tobacco. Agric Ecosyst Environ 136: 49-58 Germaine KJ, Keogh E, Ryan D, Dowling DN (2009) Bacterial endophyte-mediated naphthalene phytoprotection and phytoremediation. FEMS Microbiol Lett 296: 226-234 Gilbertson AW, Fitch MW, Burken JG, Wood TK (2007) Transport and survival of GFPtagged root-colonizing microbes: Implications for rhizodegradation. Eur J Soil Biol 43: 224-232 Glick BR (2010) Using soil bacteria to facilitate phytoremediation. Biotechnol Adv 28: 367 -374 Glick BR (2014) Bacteria with ACC deaminase can promote plant growth and help to feed the world. Microbiol Res 169: 30-39 Guadagnini M (2000) In vitro-breeding for metal-accumulation in two tobacco (Nicotiana tabacum) cultivars. Innagural-Dissertation No. 1288 der Mathematisch Naturwissenschftlichen Fakultät der Universität reiburg in der Schweiz. p. 109. Gunderson JJ, Knight JD, Van Rees KCJ (2007) Impact of ectomycorrhizal colonization of hybrid poplar on the remediation of diesel-contaminated soil. J Environ Qual 36: 927-934 Gurska J, Wang W, Gerhardt KE, Khalid AM, Isherwood DM, Huang XD, Glick BR, Greenberg BM (2009) Three year field test of a plant growth promoting rhizobacteria enhanced phytoremediation system at a land farm for treatment of hydrocarbon waste. Environ Sci Technol 43: 4472-4479 Haslmayr HP, Meißner S, Langella F, Baumgarten A, Geletneky J (2014) Establishing best practice for microbially aided phytoremediation. Environ Sci Pollut Res 21: 6765-6774 Herzig R, Nehnevajova E, Pfistner C, Schwitzguebel JP, Ricci A, Keller C (2014) Feasibility of labile zn phytoextraction using enhanced tobacco and sunflower: Results of fiveand one-year field-scale experiments in Switzerland. Int J Phytorem 16: 735-754 Janssen J, Weyens N, Croes S, Beckers B, Meiresonne L, Van Peteghem P, Carleer R, Vangronsveld J (2015) Phytoremediation of metal contaminated soil using willow: exploiting plant-associated bacteria to improve biomass production and metal uptake. Int J Phytorem 17: 1123-1136 Kabata-Pendias A (2010) Trace Elements in Soils and Plants (4th ed.). Florida: CRC Press LLC, Boca Raton Kapulnik Y, Okon Y, Henis Y (1985) Changes in root morphology of wheat caused by Azospirillum inoculation. Can J Microbiol 31: 881-887 Kennedy VH, Sanchez AL, Oughton DH, Rowland AP (1997) Use of single and sequential chemical extractants to assess radionuclide and heavy metal availability from soils for root uptake. Analyst 122: 89-100 Kidd P, Barceló J, Bernal MP, Navari-Izzo F, Poschenrieder C, Shilev S, Clemente R, Monterroso C (2009) Trace element behaviour at the root–soil interface: Implications in phytoremediation. Environ Exp Bot 67: 243-259 Kidd P, Mench M, Álvarez-López V, Bert V, Dimitriou I, Friesl-Hanl W, Herzig R, Janssen JO, Kolbas A, Müller I, Neu S, Renella G, Ruttens A, Vangronsveld J, Puschenreiter M (2015) Agronomic practices for improving gentle remediation of trace element-contaminated soils. Int J Phytorem 17: 1005-1037 Kolbas A, Kidd P, Guinberteau J, Jaunatre R, Herzig R, Mench M (2015) Endophytic bacteria take the challenge to improve Cu phytoextraction by sunflower. Environ Sci Pollut Res 22: 5370-5382 Kuffner M, De Maria S, Puschenreiter M, Fallmann K, Wieshammer G, Gorfer M, Strauss J, Rivelli AR, Sessitsch A (2010) Culturable bacteria from Znand Cdaccumulating Salix caprea with differential effects on plant growth and heavy metal availability. J Appl Microbiol 108: 1471-1484 Lampis S, Santi C, Ciurli A, Andreolli M, Vallini G (2015) Promotion of arsenic phytoextraction efficiency in the fern Pteris vittata by the inoculation of Asresistant bacteria: a soil bioremediation perspective. Front Plant Sci 6:80 25 Inoculation method influences plant responses of Nicotiana tabacum Chapter 8 260 Clemente et al., 5XWWHQVet al., 2006). Incorporation of organic residues (in particular composts, manure, sewage sludge or biosolids) improves soil physical properties, water infiltration and WHC. They further contain microand macronutrients for plant growth, and also decrease bulk density (Vangronsveld et al.,  6RULDQR-Disla et al., 2010). Addition of this type of amendment is a common practice to facilitate re-vegetation of contaminated soils and at the same time provides a viable manner of recycling waste products. Zanuzzi et al. (2009) amended acidic mine tailings with pig manure or sewage sludge in combination with blanket application of marble wastes. Two years after soil amendment soil pH increased (from pH 2.7 to 7.4) alongside a build-up in total organic carbon. The authors proposed that the formation of stable soil organic matter–calcite complexes provided water and nutrients to support initial seedling establishment in the mine tailings. After 5 years, microbial biomass and soil enzymatic activities, as well as vegetation cover and richness, were higher in the amended soils than in the untreated mine tailings (Zornoza et al., 2012). The effects of organic amendments on TE bioavailability depend on the OM nature, and on the particular soil type and elements concerned (Clemente et al.,  5XWWHQV et al.,  .XPSLHQH et al.,  /DJRPDUVLQR et al., 2011). Some forms of unstabilised organic matter such as manure or biosolids may increase TE mobility due to their high dissolved organic matter (DOM) content that can complex metals and result in their leaching through soil (Kiikkilä et al., D7DQG\et al., 2009). Composting stabilises organic wastes and can reduce their DOM content and the potential for metal(loid) leaching (Kiikkilä et al., 2002b). Composted organic wastes can successfully improve the quality and reduce the mobility of contaminants in metal-contaminated mine-soils (Pardo et al., 6ROtV-Dominguez et al., 1RYRet al., 2013). Only a limited number of studies evaluating the feasibility of phytostabilisation techniques under field conditions can be found, and there is a clear need for long-term field experiments monitoring the efficiency of these processes if they are to reach full-scale deployment (Mench et al., 9DQJURQVYHOGet al., .LGGet al., 2015). Spain has a long history of mining activities as a result of the large extensions of metalliferous ore deposits and is one of the primary producers of Cu, Fe, Zn and Pb in the European Union (Instituto Geológico y Minero de España, 2014). Here, a field trial of aided phytostabilisation was implemented in spring 2011 in Cu-rich mine tailings in the NW of Spain. The mine tailings at this site were amended with composted municipal solid wastes and planted with 261 Monitoring the effectiveness of (aided) phytostabilisation in a field trial Salix spp., Populus nigra L. or Agrostis capillaris L. cv. Highland. Plant growth, nutritive status and metal accumulation, and soil physico-chemical and biological properties, were monitored over four years. The objective was to establish a healthy plant cover (either a SRC system or a grass cover) by reducing Cu mobility and phytotoxicity and improving soil fertility and health. 8.2. Materials and Methods Study site Experimental field plots were established in spring 2011 in tailings at the 7RXURFRSSHUPLQH$&RUXxD1:6SDLQ¶´1¶´:7KHPLQH was active from 1974 until 1988 and the area is now confined to extraction of material for road construction. The geological substrate and characteristics of the tailings were previously described by Calvo de Anta et al. (1991), Vega et al. (2006), Álvarez et al. (2011) and Asensio et al. (2013). The mine tailings cover an area of around 550 ha and are practically bare of vegetation. The geological substrate is amphibolite, with significant quantities of metal sulphides (pyrite, pyrrhotite, and chalcopyrite). The mine soils are shallow and extremely gravelly, and are classified as Spolic Technosols (Episkeletic) (IUSS Working Group WRB, 2006). The uncontrolled oxidation of sulphides generates hyperacidic (pH 2-3) and hyperoxidising (>500 mvol) soils and waters, with a high electrical conductivity (EC) and high concentrations of sulphates, Fe, Al, and potentially toxic trace metal(loid)s. The area selected for the experimental field plots was completely barren of vegetation. The climate of the region is Atlantic (oceanic), with a mean annual precipitation of 1900 mm and mean annual temperature of 12.6ºC (Asensio et al., 2013). Experimental design and set-up The experimental field plot (surface area of approx. 0.1 ha) was amended with mature organic compost in spring 2011 at an addition rate of 250 kg Mg-1 dry tailings. The compost was elaborated from municipal solid wastes and bark chippings by the local company Tratamientos Ecológicos de Noroeste (TEN s.l.) which is dedicated to the recycling of waste products and manufacturing of organic fertilisers and soil improvers. The physicochemical properties of the compost are given in Table 8.1. The tailings (the first 50 cm of surface soil) were broken-up prior to compost addition using a bull dozer (equipped with ripper), Chapter 8 262 and mechanical soil tillage was carried out after compost addition to a depth of 0-30 cm. The compost was mechanically applied. Three replicate sub-plots (5 x 5 m) were planted four weeks after amendment incorporation with each plant species. A. capillaris cv. Highland was seeded at a rate equivalent to 250 kg ha-1. Rooted cuttings of two Salix spp. (Salix caprea Mauerbach (BOKU 01 AT-004) and S. viminalis) and P. nigra were planted with a 1 m spacing to yield a planting density of 10000 cuttings ha–1. The cuttings of S. viminalis were obtained from the Touro mine itself where they can be found growing sporadically. In hydroponic solutions they showed relevant metal tolerance and reduced transport of metals (including Cu) to aerial tissues (unpublished results). Salix caprea Mauerbach was shown in previous experiments to be metal tolerant and exclude Cu from its aerial shoots (Dos Santos Utmazian et al., 2007). Cuttings of P. nigra were obtained from the Biogeco experimental site at a former wood preservative facility (Gironde, France) where phytostabilisation trials were implemented in 2006 to reduce Cu mobility and phytotoxicity (Lagomarsimo et al., 2011). Unplanted compost-amended sub-plots (5 x 5 m) were also established to evaluate the effects of plant growth and activity. Non-amended plots were not established since nothing grows in the mine tailings unless some type of amendment is added. Weed control was carried out mechanically throughout each growth season over the period 2011 to 2015. The set-up of the experimental plots, incorporation of the compost amendment and evolution in plant growth can be seen in Figure 8.1. Soil sampling and analyses Soil samples (five samples per sub-plot from the upper 20 cm layer) were taken with an unpainted stainless steel spade to determine the baseline trace metal content and general physicochemical properties across the whole experimental plot before amendment addition/planting (T=0). To monitor changes in the physicochemical, biochemical and biological properties soils were then sampled 1, 2 and 3 years after incorporating the compost. Each soil sample was divided into two parts, one part for physicochemical analyses and one part for biochemical/ microbiological analyses. For physicochemical analyses, soil samples were airdried and sieved to <2 mm. Fresh soil samples were sieved to <2 mm and stored at 4ºC for biochemical analyses or at -60ºC until DNA extraction for microbiological analyses. Soil pH was measured in H2O and 0.1 M KCl using a 1:2.5 (w/v) ratio. Total C and N were analysed by combustion with a CHN analyser (Model CHN-1000, 263 Monitoring the effectiveness of (aided) phytostabilisation in a field trial LECO Corp., St Joseph, MI). Soils were extracted with Olsen´s reagent (0.5 M NaHCO3 adjusted to pH 8.2, 1:20 w/v), and available P determined colorimetrically using the molybdenum blue method (Murphy and Riley, 1962). Soils (0.5 g) were digested in a 3:1 mixture of concentrated HNO3:HCl after the addition of 3 mL H2O2 in Teflon PFA vessels in a microwave accelerated reaction system (MarsXpress; CEM Corp., USA) and the total concentrations of metals were analysed by inductively coupled plasma optical emission spectrometry (ICP-OES, model Vista-PRO, Varian). A certified soil standard (ISE 979 Rendzina Soil) was used to ensure quality of metal quantification. Exchangeable cations (Ca, Mg, Al, Na and K) were extracted with 1M NH4Cl (1:20 w/v 16h equilibration) and measured by ICP-OES. Metal availability was evaluated by extraction with 0.1 M NaNO3 (1:2.5 w/v, 2h shaking) and elements were analysed by ICP-OES. Soil biochemical properties were monitored in collaboration with the Soil Enzymology Group of the IIAG (CSIC). Soil enzymatic analyses were determined in the non-amended tailings soil (before incorporation of compost), nonplanted compost-amended soil and the compost-amended soil with either S. viminalis or A. capillaris plant cover after 1, 2 and 3 years. Dehydrogenase activity was determined as described by Camiña et al. (1998), and the activity was exFigure 8.1. Implementation of the aided phytostabilisation field trial in the mine tailings. (a) Non-amended mine tailings, (b, c) addition and incorporation of compost amendment, (d) planting of woody tree cuttings, (e) growth of woody species after 1 year, (f) growth of Agrostis capillaris after 2 years, (g, h) growth of Salix and Populus after 2 years and (i) after 3 years, and (j-m) re-sprouting of trees after the first harvest. Chapter 8 264 expressed as µmol INTF g-1 h-1. Catalase activity was determined according to Trasar-Cepeda et al. (1999) and expressed in mmol H2O2 consumed g-1 h-1. Urease activity was determined as described by Nannipieri et al. (1980), except that the released NH4+ was measured with an ammonia electrode. The activity was expressed in µmol NH3 g-1 h-1. Invertase activity was determined following the method described by Schinner and von Mersi (1990). Acid and alkaline phosphomonoesterase activity were determined at pH 5.0 and 10.0, respectively, with 16 mM p-nitrophenyl phosphate as a substrate following the method of Tabatabai and Bremner (1969), but with some modifications. Modified Universal Buffer was used to maintain the pH of the reaction mixture as described by Trasar-Cepeda et al. (1985). After the incubation (30 min) 2 M CaCl2 was added (to prevent dispersion of soil colloids and to avoid the brown coloration caused by organic matter) and the liberated p-nitrophenol was extracted with 0.2 M NaOH. The enzymatic activity was quantified by reference to calibration curves corresponding to pnitrophenol standards incubated with each soil under the same conditions as for the samples (Saá,1995). ß-glucosidase activity was determined as described for acid phosphomonoesterase activity except that the substrate was 25 mM pnitrophenyl-ß-glucopyranoside and the p-nitrophenol released was extracted with 0.1 M THAM-NaOH at pH 12 (Eivazi and Tabatabai, 1988). Arylsulphatase activity was determined following the method described by Tabatabai and Bremner (1970), but using 2 M CaCl2 and 0.2 M NaOH instead of 0.5 M CaCl2 and 0.5 M NaOH, and quantifying the enzymatic activity by reference to calibration curves corresponding to p-nitrophenol standards incubated with each soil under the same conditions as for the samples (Saá,1995). For the latter four enzymes, the enzymatic activity was expressed in µmol p-nitrophenol g-1 h-1. For each soil sample, all determinations were made in triplicate and the average values were expressed on an oven-dried soil basis. Monitoring of plant survival, growth and ionome The maximum shoot height of trees was measured 0.5, 1, 2 and 3 years after incorporation of the compost. Tree mortality was also recorded and dead trees were replaced during the first year. To monitor the plant nutritive status and shoot metal accumulation, sub-samples of the plant aerial biomass (shoots in the case of A. capillaris, and both stems and leaves in the woody species) were collected at each sampling time. Plant material was carefully washed with pressurised tap water followed by deionised water, oven-dried at 45°C and ground. Dried plant 265 Monitoring the effectiveness of (aided) phytostabilisation in a field trial material (approximately 0.1 g) were digested in a 2:1 concentrated HNO 3 :HCl mixture on a hot plate at 120 °C, and the concentrations of P, K, Ca, Mg, Fe, Cd, Cu, Mn, Ni, Pb and Zn were measured by ICP-OES. For A. capillaris, the total biomass produced in each sub-plot was harvested after two and three years (June 2013 and 2014, respectively). For tree species, biomass production was determined in each sub-plot after four years of growth (harvested in March 2015). Biomass was expressed as tonnes shoot dry weight (DW) yield per hectare. Statistical analyses At each sampling period, the effects of compost amendment and plant cover on soil physicoand bio-chemical properties, and plant growth and ionome, was assessed using a one-way analyses of variance (ANOVA) followed by a multiple comparison of means using the post-hoc Tukey-test for independent means. When necessary, data were transformed to meet the requirements of ANOVA. The statistical analysis was performed using IBM SPSS Statistics v22 for Windows. 8.3. Results Effect of compost addition on soil physicoand bio-chemical properties Table 8.1 lists the mean values (and range) for the main soil properties of the non-amended mine tailings. The tailings soil present a highly acidic pH and low fertility (reflected through the low total C and N content, and available P). The effective CEC was low and the H + and Al 3+ cations were dominant. Trace element concentrations were highly variable across the tailings but, in general, elevated concentrations of Al, Cr, Cu, Fe, Mn or Ni were found (Table 8.1). Total Al and Fe concentrations fell within the same interval, with values varying from 33.4 to 58.9 g Al kg -1 and 81 to 137 g Fe kg -1 . Of the remaining TE, the principal metals were Cu and Mn: the mean total concentrations were 596 mg Cu kg -1 and 823 mg Mn kg -1 . Total concentrations of Cr and Ni ranged from 68 to 140 mg Cr kg -1 and from 27 to 181 mg Ni kg -1 . The highest values of available metals were found for Cu and Mn. The mean NaNO 3 -extractable Cu concentration was 61 mg kg -1 but values oscillated between 7 and 219 mg kg -1 across the experimental plot. The mean NaNO 3 -extractable concentration of Mn was 823 mg kg -1 (and ranged from 650 to 1142 mg kg -1 ). Chapter 8 266 Table 8.1. Physicochemical properties of the mine tailings and compost Mine tailings soil Compost Mean Range Mean (± SE) pH H2O 3.06 2.70 – 3.62 6.18 ± 0.04 pH KCl 2.90 2.30 – 3.42 6.15 ± 0.05 EC (mS cm -1 ) 1.85 0.34 – 3.97 - %C 0.60 0.10 – 1.08 15.4 ± 0.4 %N 0.10 0.08 – 0.13 1.4 ± 0.0 P olsen (mg kg -1 ) 2.5 <0.2 – 4.4 425.3 ± 33.0 CEC (cmol c kg -1 ) 21.4 7.9 – 49.3 57.6 ± 4.5 H + 9.4 3.5 – 23.7 < loq Al 3+ 8.0 3.8 – 18.9 < loq Ca 2+ 2.4 0.3 – 8.1 40.4 ± 1.2 K + < loq 4.4 ± 0.1 Mg 2+ 1.5 0.1 – 6.6 10.3 ± 0.5 Na + < loq 2.5 ± 0.1 Pseudo-total element concentrations g kg -1 P 0.7 0.5 – 1.0 11.8 ± 0.3 K 3.6 1.9 – 5.3 5.6 ± 0.2 Al 42.5 33.4 – 58.9 31.5 ± 0.4 Ca 5.9 2.9 – 12.8 30.1 ± 0.9 Mg 9.5 7.4 – 12.2 6.8 ± 0.2 Fe 105.0 81 – 137 40 ± 7 mg kg -1 Mn 823.0 650 – 1142 861 ± 20 Cu 596.0 308 – 1155 529 ± 84 Zn 123.0 95 – 366 530 ± 17 Cr 96.0 68 – 140 65.0 ± 1.6 Ni 43.0 27 – 181 59.2 ± 8.5 Cd 1.6 0.8 – 2.3 1.7 ± 0.2 Co 12.2 2.2 – 29.4 12.8 ± 0.5 NaNO 3 -extractable metal concentrations (mg kg -1 ) Mn 63.0 14 – 289 - Cu 61.0 7 – 219 - Zn 21.5 2.3 – 91.7 - Cr <loq - Ni 9.0 0.7 – 40.7 - Cd 0.12 0.01 – 0.60 - Co 7.0 0.8 – 23.5 - loq, limit of quantification: 0.01 cmolc kg-1 Al3+, K+FPROc kg-1 Na+PJ&UNJ-1 267 Monitoring the effectiveness of (aided) phytostabilisation in a field trial Table 8.2 gives an overview of the changes induced in general soil properties over time after amendment with compost (independent of plant cover). One year after compost addition, soil pH was significantly increased, and continued to increase over the following two years (p < 0.05). A progressive increase in total C and N content was also observed through time, reaching maximal values after 2 years. These increments were maintained at year 3: the total C and N content at this time was 7.90% and 0.65% (representing an increase of 12and 6-fold compared to non-amended soil, respectively). Compost addition also induced significant changes in the effective CEC. Soil CEC increased over the first year, reaching maximal values at this time of 33.2 cmolc kg-1. Thereafter, CEC progressively dropped and after 3 years similar values to those determined in the non-amended tailings were obtained. However, before amending the tailings the CEC was dominated by the acidic cations H+ and Al3+ (with a base saturation of <20%), while after three years the mean base saturation was >90% (predominantly dominated by Ca2+ 7DEOH  ([FKDQJHDEOH &D increased from 2.3 cmolc kg-1 in non-amended tailings soil to 29.3 cmolc kg-1 in year 1, thereafter decreasing to 14.8 cmolc kg-1 in year 3. Similar trends were observed in exchangeable Mg2+ and K+, although these were less pronounced. loq, limit of quantification: 0.01 cmolc K kg-1 T =0 T = 1 year T =2 years T = 3 years pH H2O 3.18 ± 0.02c 5.37 ± 0.15b 6.22 ± 0.11a 6.48 ± 0.11a pH KCl 3.06 ± 0.02c 5.16 ± 0.15b 5.71 ± 0.10a 6.04 ± 0.09a Cu concentration (mg kg -1 ) Pseudo-total 523 ± 21a 467 ± 22a 529 ± 41a 481 ± 29a NaNO 3 -extractable 33.0 ± 3.1a 1.2 ± 0.4b 0.3 ± 0.1b 0.4 ± 0.1b CEC (cmol c kg -1 ) 19.5 ± 3.5b 33.2 ± 3.9a 28.5 ± 1.6a 17.4 ± 0.8b H + 7.96 ± 1.22a 0.16 ± 0.11b < loq < loq Al 3 + 7.78 ± 1.28a 0.44 ± 0.28b 0.20 ± 0.09b 0.18 ± 0.09b Ca 2 + 2.3 ± 0.6c 29.3 ± 3.7a 24.8 ± 1.4ab 14.8 ± 0.7b K + < loq 0.85 ± 0.13a 0.95 ± 0.06a 0.64± 0.06a Mg 2 + 1.45 ± 0.50b 2.43 ± 0.42a 2.68 ± 0.15a 1.73 ± 0.09ab Polsen (mg kg -1 ) 2.1 ± 0.3c 243.0 ± 29.8a 194.3 ± 10.1ab 133.3 ± 10.6b %C 0.63 ± 0.08b 6.67 ± 0.72a 8.14 ± 0.72a 7.91 ± 0.94a %N 0.11 ± 0.00c 0.45 ± 0.07b 0.70 ± 0.06a 0.65 ± 0.05a Table 8.2. Evolution of soil physicochemical properties in the mine tailings soil after amendment with compost (considering the whole experimental plot). Different letters indicate significant differences with time (p < 0.05). Chapter 8 268 Available P (P olsen) also showed a pronounced increase after compost addition: NaHCO 3 -extractable P concentrations peaked at year 1 (representing an increase of approximately 100-fold compared to non-DPHQGHG WDLOLQJVVRLOS  0.05) and then progressively dropped during the following two years but never reached levels of non-amended tailings soil. Compost addition did not significantly alter the pseudo-total concentration of Cu (Table 8.2). However, a significant increase in the total concentration of Mn, Pb and Zn was observed: their concentrations (in mg kg -1 ) increased from 882 to 1119 for Mn, 16.8 to 46.4 for Pb, and 125.0 to 311.5 for Zn. Soil amendment had a strong effect on TE availability (estimated with the NaNO 3 -extractable concentration). The mean NaNO 3 - extractable Cu concentration in the mine tailings soil before compost addition was 61 mg kg -1 and decreased to 2.9 mg kg -1 after one year and to 0.5 mg kg -1 after three years (p < 0.05). NaNO 3 -extractable concentrations of TE such as Al, Cd, Co, Cr, Ni and Zn were below the quantification limit (0.025 mg kg -1 for Cd, Mn and Ni, 0.0125 mg kg -1 for Co and Cr and 0.125 mg kg -1 for Zn) after three years, and the mean Mn concentration in this labile pool was 1.4 mg kg -1 . The lowest soil enzyme activities were always found in the non-amended tailings soil, and the compost addition induced a significant increase in all assayed enzymes (Table 8.3). In non-amended soil the activity of the oxido-reductase enzymes was null, while one year after compost amendment mean values of catalase and dehydrogenase activities of 1.17 mmol H 2 O 2 consumed g -1 h -1 and 0.28 μmol INTF g -1 h -1 were recorded. For the hydrolase enzymes, enzymes activities were increased by 4to 23-fold (depending on the enzyme) after one year. The Table 8 . 3. Evolution of soil enzyme activities in the mine tailings soil after amendment with compost (considering the whole experimental plot). Different letters indicate significant differences with time (p < 0.05). T =0 T = 1 year T =2 years T = 3 years Catalase (mmol H 2 O 2 consumed g -1 h -1 ) 0.00 ± 0.00c 1.27 ± 0.14ab 0.84 ± 0.12b 1.56 ± 0.11a Dehydrogenase (ȝmol INTF g -1 h -1 ) 0.01 ± 0.00c 0.28 ± 0.02a 0.18 ± 0.02b 0.19 ± 0.02b Invertase (ȝmol glucose g -1 h -1 ) 0.50 ± 0.02c 4.64 ± 0.55ab 3.84 ± 0.36b 6.63 ± 0.63a ß-glucosidase (ȝmol PNP g -1 h -1 ) 0.12 ± 0.02c 0.43 ± 0.05c 0.74 ± 0.06b 1.06 ± 0.09a Urease (ȝmol NH 3 g -1 h -1 ) 0.16 ± 0.03c 3.74 ± 0.31ab 2.59 ± 0.22b 4.83 ± 0.38a Acid phosphomonoesterase (ȝmol PNP g -1 h -1 ) 0.41 ± 0.04c 1.57 ± 0.12ab 1.70 ± 0.11a 1.23 ± 0.06b Alkaline phosphomonoesterase (ȝmol PNP g -1 h -1 ) 0.21 ± 0.01b 1.16 ± 0.13a 0.70 ± 0.11ab 0.85 ± 0.1ab Arylsulphatase (ȝmol PNP g -1 h -1 ) 0.00 ± 0.00c 0.08 ± 0.01ab 0.06 ± 0.01b 0.12 ± 0.01a  269 Monitoring the effectiveness of (aided) phytostabilisation in a field trial most pronounced increases were observed in urease (mean values increased from 0.16 μmol NH3 g-1 h-1 to 3.74 μmol NH3 g-1 h-1) and arylsulphatase (no activity was observed in non-amended tailings and after one year a mean value of 0.08 μmol PNP g-1 h-1 was obtained). In general, these significant increments in activities were maintained over the experimental period in all enzymes except dehydrogenase, while a further significant increase in the activities of invertase and ȕ-glucosidase was observed between 2 and 3 years (Table 8.3). Evolution in plant growth and shoot ionome The maximum height (per individual) of both Salix spp. and P. nigra was recorded after the first growth season (6 months), and thereafter, on a yearly basis XQWLOKDUYHVWLQJDIWHU\HDUV)LJXUH$ODUJHYDULDELOLW\LQSODQWKHLJKWZDV observed between individuals of the same species across the three replicate subplots. After the first growth season the mean maximum height recorded was 106 cm for S. viminalis, 80 cm for S. caprea and 96 cm for P. nigra. Plant height did not change significantly between 6 months and 1 year, but individuals of all three 400 Tree height (cm) 300 200 100 0 Time (years) 0.5 1 2 3 4 S. viminalis S. caprea P. nigra Figure 8 . 2. Maximum shoot height (cm) of Salix viminalis, Salix caprea and Populus nigra.  Boxes represent the median (vertical line) and 25 and 75% quartiles. Whiskers represent the 90th and 10th percentile. Outliers are represented by unfilled circles. Different capital letters denote significant differences between years within each plant species and different lower case letters indicate significant differences between plant species within the same year (p < 0.05). Chapter 8 276 Plant cover had a significant effect on biochemical parameters. Figure 8.6 presents the soil enzyme activities for non-amended mine tailings at the beginning of the experiment, and the evolution in activities over time in unplanted soil, and soils cultivated with either S. viminalis or A. capillaris. The mean values of enzyme activities were always higher in planted soils than in unplanted soils, and this was the case for all sampling time points and all assayed enzymes (Figure 8.6). In general, enzyme activities in planted soils did not progressively increase over time, although urease, invertase, β-glucosidase and arylsulphatase activities were significantly higher after 3 years than at either 1 or 2 years. However, enzyme activities were influenced by plant species, and the mean values in soils under S. viminalis were generally higher than those recorded under A. capillaris (albeit not always statistically significant). For example, after 2 years the activities of catalase and dehydrogenase were 1.5and 1.6-fold higher in soil under S. viminalis than A. capillaris (p < 0.05). A similar effect was observed in P cycle enzymes: the activity of acid and alkaline phosphomonoesterase (in µmol PNP g-1 h-1) was 1.65 and 0.44 in soil under A. capillaris but increased to 2.05 and 1.16 in soil under S. viminalis (not significant in the case of acid phosphomonoesterase). Figure 8.5. Effect of plant species (Salix viminalis, S. caprea, Populus nigra and Agrostis capillaris) on soil NaNO 3 -extractable Cu concentrations. Boxes represent the median (vertical line) and 25 and 75% quartiles. Whiskers represent the 90th and 10th percentile. Outliers are represented by unfilled circles. Different capital letters denote significant differences between years within each plant species and different lower case letters indicate significant differences between plant species within the same year (p < 0.05). NaNO3-extractable Cu (mg kg-1) 0 20 40 60 80 100 120 140 Unplanted S. viminalis S. caprea P. nigra A. capillaris 00 05 10 15 20 25 Time (years) 0 1 3 4 0 1 2 3 1 2 3 [Document text truncated for crawler view.]