AGRICULTURAL USE OF ORGANIC WASTES AS SOURCE OF...
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UNIVERSIDAD POLITÉCNICA DE MADRID Escuela Técnica Superior de Ingenieros Agrónomos AGRICULTURAL USE OF ORGANIC WASTES AS SOURCE OF NITROGEN AND PHOSPHORUS: RISKS AND OPPORTUNITIES Tesis Doctoral María Isabel Requejo Mariscal Ingeniera Agrónoma 2015
Transcript of AGRICULTURAL USE OF ORGANIC WASTES AS SOURCE OF...
UNIVERSIDAD POLITÉCNICA DE MADRID
Escuela Técnica Superior de Ingenieros Agrónomos
AGRICULTURAL USE OF ORGANIC WASTES AS
SOURCE OF NITROGEN AND PHOSPHORUS:
RISKS AND OPPORTUNITIES
Tesis Doctoral
María Isabel Requejo Mariscal
Ingeniera Agrónoma
2015
Departamento de Química y Tecnología de los Alimentos
Escuela Técnica Superior de Ingenieros Agrónomos
AGRICULTURAL USE OF ORGANIC WASTES AS
SOURCE OF NITROGEN AND PHOSPHORUS:
RISKS AND OPPORTUNITIES
Autora:
María Isabel Requejo Mariscal,
Ingeniera Agrónoma.
Directores:
María Carmen Cartagena Causapé,
Catedrática Universidad Politécnica de Madrid.
Francisco Ribas Elcorobarrutia,
Doctor Instituto Regional de Investigación y Desarrollo
Agroalimentario y Forestal de Castilla-La Mancha.
Madrid
2015
Tribunal nombrado por el Magnífico y Excelentísimo Sr. Rector de la Universidad de la
Universidad Politécnica de Madrid, el día de de 2015.
Presidente:________________________________________________________________________
Secretario:________________________________________________________________________
Vocal:____________________________________________________________________________
Vocal:____________________________________________________________________________
Vocal:____________________________________________________________________________
Suplente:_________________________________________________________________________
Suplente:_________________________________________________________________________
Realizado el acto de defensa y lectura de Tesis el día de de 2015
en la E.T.S.I. Agrónomos.
EL PRESIDENTE LOS VOCALES
EL SECRETARIO
AUTORIZACIÓN PARA LA DEFENSA
María del Carmen Cartagena Causapé, Catedrática del Departamento de Química y
Tecnología de los Alimentos de la Universidad Politécnica de Madrid, y Francisco Ribas
Elcorobarrutia, Investigador del Centro de Investigación Agroambiental el Chaparrillo
perteneciente al Instituto Regional de Investigación y Desarrollo Agroalimentario y
Forestal de Castilla-La Mancha, como directores de la Tesis Doctoral titulada
“Agricultural use of organic wastes as source of nitrogen and phosphorus: risks and
opportunities”, realizada por María Isabel Requejo Mariscal, alumna del Programa de
Doctorado de Tecnología Agroambiental Para una Agricultura Sostenible del
Departamento de Química y Tecnología de los Alimentos, para aspirar al grado de
Doctor por la Universidad Politécnica de Madrid,
AUTORIZAN
La presentación de esta memoria para que se proceda al trámite de su lectura y
defensa ante el tribunal correspondiente.
Y para que conste a los efectos oportunos, firmamos el presente certificado en Madrid,
a de de 2015.
Fdo: Dra. Mª Carmen Cartagena Causapé Fdo.: Dr. Francisco Ribas Elcorobarrutia
Tras estos años de trabajo, cuando todo ya ha tomado forma y a punto de cerrar esta
etapa de mi vida, es el momento de echar la vista atrás, recordar y agradecer a todas
las personas e instituciones que han formado parte de esto y lo han hecho posible.
Quisiera dar las gracias en primer lugar a mis directores de tesis, Mª Carmen Cartagena
y Francisco Ribas. Gracias por darme la oportunidad de realizar esta tesis, por confiar en
mí, por apoyarme y formarme para empezar a desenvolverme en este mundo, por
vuestros consejos y ayuda para conseguir que todo salga adelante, y sobre todo, por
vuestra estima y cariño.
Mayte Castellanos, a ti te debo gran parte de lo que sé, gracias por tu tiempo y
dedicación, por tu amistad y por tantas risas y buenos ratos juntas. Raquel Villena,
gracias por tu ayuda con los análisis del laboratorio y tu buena disposición siempre.
Gracias Mª Jesús Cabello y Augusto Arce, por vuestra ayuda y dedicación en el
proyecto. A Ana Tarquis, por estar dispuesta siempre a resolver cualquier duda. Marta
García, gracias por todo lo que me enseñaste en los comienzos, por engancharme al
“mundo P”, por los congresos, las risas y tus ánimos y energía positiva en los momentos
de bajón.
El desarrollo de esta tesis doctoral ha sido posible gracias al apoyo económico del
Instituto Nacional de Investigación y Tecnología Agraria y Alimentaria a través de una
beca predoctoral y un proyecto de investigación (RTA2010-00110-C03).
Agradecer al Director del Centro de Investigación Agroambiental El Chaparrillo su
apoyo al proyecto y al desarrollo de esta tesis.
Al personal de la finca “La Entresierra”, Andrés, Adolfo, Rafa, Petro, Ángel, Vicky,
Domingo, Boni, Juan Carlos y a todos los que han pasado por allí, porque gracias a su
experiencia, conocimiento y buen hacer salen adelante los ensayos de campo. Gracias
a mis compañeros del grupo de Leñosos, Mari Carmen, Houssem y David, por su ayuda
y sus consejos, por tantos desayunos inolvidables en la cocinilla de la finca. A Sandra y
Marta, que se encargaron de enseñarme cómo funcionaba todo cuando llegué de
nuevas al Chaparrillo.
Agradecer también a todo el personal del Departamento de Química y Tecnología de
los Alimentos, Pilar Ortiz, Pilar López, Javi, Paloma y Ana, su paciencia y ayuda en el
laboratorio. Gracias a las chicas que han pasado por el Departamento para realizar su
Trabajo Fin de Carrera por su aportación a esta tesis, a Lara, Luna, y en especial a Marta
y Raquel.
Gracias a Juan Carlos y Félix Cuevas, de la empresa Agricompost S.L., por suministrarnos
el compost y los residuos de bodegas y alcoholeras utilizados en los diferentes
experimentos de esta tesis, así como por su buena disposición a la hora de facilitarnos
toda la información relativa a los mismos.
Gracias a las personas que he conocido en Ciudad Real, a Sara, Brenda, Jenni, Alicia
Fronti, Dulce…porque son los responsables de tantos buenos momentos y recuerdos que
me llevo del tiempo vivido allí. En especial a mi familia de Carmen 5, los “nómadas
felices”, Cris y Dani, porque cada vez que volvía de Madrid y abría la puerta de casa
me encontraba con un auténtico hogar del que me siento muy afortunada de haber
formado parte.
Me gustaría agradecer también a los grupos de investigación con los que he realizado
estancias así como a todas las personas a las que he conocido y que han hecho de mis
meses en el extranjero una gran experiencia. A Bettina Eichler-Löbermann, de la
Universidad de Rostock, por acogerme en su grupo y darme la oportunidad de aprender
una nueva técnica, por su ayuda, apoyo y confianza en mí. A Brigitte Klaus, por su gran
ayuda en el laboratorio y en todas las gestiones, siempre con una sonrisa. Al resto del
grupo y gente que conocí en Rostock, en especial a Sebastián y Adriana, por su
acogida durante mis meses allí. A Else Bünemann, Emmanuel Frossard y a todo el grupo
de Nutrición Vegetal del ETH de Zúrich, por darme la oportunidad de aprender de sus
técnicas y forma de trabajar. A todos mis compañeros de Eschikon, Nico, Gregor,
Chiara, Klaus, Eva B., Eva M., Salvatore, Andrea, Salomé, Arif…por tantos buenos ratos
en el laboratorio, porque la vida sin los “boffee break” es mucho más aburrida. A Edu y
Clara, por los paseos en las montañas suizas.
A mis mejores amigas, Bea y Ana, por tantos años de amistad, alegrías y penas
compartidas.
Gracias mamá y papá, por vuestro cariño infinito, por cuidarme siempre, porque gracias
a vuestra confianza en mí y a la libertad que siempre me habéis dado, hoy estoy aquí.
Gracias a mi hermano y al resto de mi familia, por apoyarme y creer en mí, sois los
mejores.
Jose, me cuesta encontrar palabras de agradecimiento, eres el mejor compañero que
podría desear. Gracias por quererme y comprenderme especialmente en los malos
momentos. Sueño con los días que nos esperan.
GRACIAS A TODOS!
María.
…A mi familia, a Jose.
List of contents
RESUMEN i
ABSTRACT v
LIST OF ABBREVIATIONS ix
LIST OF TABLES xiii
LIST OF FIGURES xv
CHAPTER 1: INTRODUCTION 1
1.1. ORGANIC WASTE GENERATION AND MANAGEMENT IN THE UE……………………..3
1.1.1. A specific case of Mediterranean countries: wastes derived from the winery
and distillery industries…………………………………………………………………………4
1.2. TOWARDS A SUSTAINABLE USE OF NITROGEN AND PHOSPHORUS IN
AGRICULTURE………………………………………………………………………………………..6
1.3. UNDERSTANDING SOIL NUTRIENT CYCLES: A KEY FACTOR……………………………..8
1.3.1. Soil nitrogen cycle……………………………………………………………………..8
1.3.2. Soil phosphorus cycle………………………………………………………………..10
1.4. NITROGEN, PHOSPHORUS AND THE ENVIRONMENT……………………………………13
1.5. METHODOLOGIES TO ASSESS NITROGEN AND PHOSPHORUS IN SOILS……………..17
1.5.1. Nitrogen………………………………………………………………………………...17
1.5.1.1. Inorganic N………………………………………………………………………17
1.5.1.2. Total and organic N……………………………………………………………17
1.5.2. Phosphorus…………………………………………………………………….……….18
1.5.2.1. Inorganic P……………………………………………………………….………18
1.5.2.2. Total and organic P…………………………………………………….………20
1.6. ORGANIC WASTES AS SOURCE OF NUTRIENTS IN AGRICULTURE……………….…….23
CHAPTER 2: OBJECTIVES 27
2.1. NITROGEN……………………………………………………………………………………...29
2.2. POSPHORUS……………………………………………………………………………………30
2.3. METHODOLOGY………………………………………………………………………………30
CHAPTER 3: NITROGEN MINERALIZATION IN SOIL AFTER ADDITION OF WINE-
DISTILLERY WASTE COMPOST: LABORATORY AND FIELD EVALUATION 33
3.1. ABSTRACT………………………………………………………………………………………35
List of contents
3.2. INTRODUCTION………………………………………………………………………………..36
3.3. MATERIAL AND METHODS…………………………………………………………………...37
3.3.1. Laboratory incubation experiment…………………………………………………37
3.3.2. Field experiment………………………………………………………………………..39
3.3.3. Nitrogen balance………………………………………………………………….…...41
3.3.4. Prediction of N mineralized in the field……………………………………….……42
3.3.5. Statistical analysis………………………………………………………………………43
3.4. RESULTS AND DISCUSSION…………………………………………………………………..43
3.4.1. Nitrogen mineralization observed in laboratory conditions……………………43
3.4.2. Nitrogen mineralization observed in the field: N balance……………………..47
3.4.3. Prediction of N mineralized in the field…………………………………………….49
3.5. CONCLUSIONS………………………………………………………………………………..51
CHAPTER 4: WINE-DISTILLERY WASTE COMPOST ADDITION TO A DRIP-IRRIGATED
HORTICULTURAL CROP OF CENTRAL SPAIN: RISK ASSESSMENT 53
4.1. ABSTRACT………………………………………………………………………………………55
4.2. INTRODUCTION………………………………………………………………………………..56
4.3. MATERIAL AND METHODS…………………………………………………………………...58
4.3.1. Experimental design and compost treatments…………………………………..58
4.3.2. Irrigation and drainage………………………………………………………………..61
4.3.3. Nitrogen leaching……………………………………………………………………...62
4.3.4. Environmental indices…………………………………………………………………63
4.3.5. Fruit yield, management efficiency and N uptake……………………………...63
4.3.6. Statistical analysis……………………………………………………………………....63
4.4. RESULTS AND DISCUSSION…………………………………………………………………..64
4.4.1. Irrigation and drainage……………………………………………………………….64
4.4.2. Nitrate concentration in the soil solution and N leaching……………………..65
4.4.3. Environmental indices…………………………………………………………………69
4.4.4. Fruit yield, management efficiency and N uptake……………………………...70
4.5. CONCLUSIONS………………………………………………………………………………...72
CHAPTER 5: WINERY AND DISTILLERY WASTES AS SOURCE OF PHOSPHORUS IN
CALCAREOUS SOILS 73
5.1. ABSTRACT………………………………………………………………………………………75
5.2. INTRODUCTION………………………………………………………………………………..76
List of contents
5.3. MATERIAL AND METHODS…………………………………………………………………...77
5.3.1. Organic materials………………………………………………………………………77
5.3.2. Incubation experiment………………………………………………………………..79
5.3.3. Statistical analysis………………………………………………………………………80
5.4. RESULTS AND DISCUSSION…………………………………………………………………..81
5.4.1. Soils and organic wastes characteristics…………………………………………..81
5.4.2. Evolution of pH………………………………………………………………………….83
5.4.3. Evolution of organic disolved P……………………………………………………...84
5.4.4. Evolution of Olsen P and recovery of applied P………………………………….86
5.4.5. Changes in soil P status: inorganic P fractions, P sorption capacity and
degree of P saturation………………………………………………………………………..89
5.5. CONCLUSIONS………………………………………………………………………………..93
CHAPTER 6: ORGANIC AND INORGANIC PHOSPHORUS FORMS IN SOIL AS
AFFECTED BY LONG-TERM APPLICATION OF ORGANIC AMENDMENTS 95
6.1. ABSTRACT………………………………………………………………………………………97
6.2. INTRODUCTION………………………………………………………………………………..98
6.3. MATERIAL AND METHODS…………………………………………………………………...99
6.3.1. Site characterization…………………………………………………………………...99
6.3.2. Soil analysis……………………………………………………………………………..101
6.3.3. NaOH–EDTA extraction and enzymatic incubation assay……………………101
6.3.4. Statistical analyses…………………………………………………………………....103
6.4. RESULTS AND DISCUSSION…………………………………………………………………104
6.4.1. Effect of long-term P amendment practice on inorganic and total soil P
pools…………………………………………………………………………………………….104
6.4.2. Effect of long-term P amendment practice on activities of enzymes……..107
6.4.3. NaOH–EDTA extractable P forms in manure and compost…………………..108
6.4.4. Effect of long-term P amendment practice on organic and inorganic P
forms in the NaOH–EDTA soil extracts…………………………………………………….109
6.5. CONCLUSIONS……………………………………………………………………………..113
CHAPTER 7: AGRONOMIC EVALUATION OF WINE-DISTILLERY WASTE COMPOST
ADDITION TO A MELON CROP: NITROGEN AND PHOSPHORUS 115
7.1. ABSTRACT……………………………………………………………………………………..117
7.2. INTODUCTION………………………………………………………………………………..118
7.3. MATERIAL AND METHODS………………………………………………………………….119
List of contents
7.3.1. Experimental design………………………………………………………………….119
7.3.2. Plant growth and leaf area index…………………………………………………121
7.3.3. Nitrogen and phosphorus content in aboveground organs…………………122
7.3.4. Crop yield and quality……………………………………………………………….122
7.3.5. Available N and P in soil……………………………………………………………..122
7.3.6. Statistical analysis……………………………………………………………………..122
7.4. RESULTS AND DISCUSSION…………………………………………………………………124
7.4.1. Plant growth and leaf area index…………………………………………………124
7.4.2. Nitrogen and phosphorus plant accumulation…………………………………126
7.4.3. Fruit yield and quality………………………………………………………………...129
7.4.4. N and P availability for the subsequent crops…………………………………..131
7.5. CONCLUSIONS……………………………………………………………………………….132
CHAPTER 8: GENERAL DISCUSSION 135
8.1. EFFECT OF THE APPLICATION OF WASTES FROM THE WINERY AND DISTILLERY
INDUSTRY ON NITROGEN DYNAMICS…………………………………………………………137
8.1.1. Effect on soil N mineralization rate………………………………………………...137
8.1.2. Effect of the application dose……………………………………………………..138
8.1.3. Effect of irrigation practices………………………………………………….……..138
8.1.4. Effect on the crop response………………………………………………………...139
8.1.5. Risk associated………………………………………………………………………...140
8.2. FACTORS INFLUENCING MOSTLY ON THE AVAILABILITY OF PHOSPHORUS IN SOILS
AFTER THE APPLICATION OF ORGANIC WASTES…………………………………………….141
8.2.1. Phosphorus forms in soil after the application of organic wastes…………...141
8.2.2. Effect of the soil type…………………………………………………………………142
8.2.3. Effect of the waste type……………………………………………………………..142
8.2.4. Effect on the crop response………………………………………………………...143
8.2.5. Risk associated………………………………………………………………………...145
CHAPTER 9: CONCLUSIONS 145
CHAPTER 10: REFERENCES 151
RESUMEN
i
RESUMEN
El nitrógeno (N) y el fósforo (P) son nutrientes esenciales en la producción de cultivos. El
desarrollo de los fertilizantes de síntesis durante el siglo XX permitió una intensificación de
la agricultura y un aumento de las producciones pero a su vez el gran input de nutrientes
ha resultado en algunos casos en sistemas poco eficientes incrementando las pérdidas
de estos nutrientes al medio ambiente. En el caso del P, este problema se agrava debido
a la escasez de reservas de roca fosfórica necesaria para la fabricación de fertilizantes
fosfatados.
La utilización de residuos orgánicos en agricultura como fuente de N y P es una
buena opción de manejo que permite valorizar la gran cantidad de residuos que se
generan. Sin embargo, es importante conocer los procesos que se producen en el suelo
tras la aplicación de los mismos, ya que influyen en la disponibilidad de nutrientes que
pueden ser utilizados por el cultivo así como en las pérdidas de nutrientes de los
agrosistemas que pueden ocasionar problemas de contaminación. Aunque la dinámica
del N en el suelo ha sido más estudiada que la del P, los problemas importantes de
contaminación por nitratos en zonas vulnerables hacen necesaria la evaluación de
aquellas prácticas de manejo que pudieran agravar esta situación, y en el caso de los
residuos orgánicos, la evaluación de la respuesta agronómica y medioambiental de la
aplicación de materiales con un alto contenido en N (como los residuos procedentes
de la industria vinícola y alcoholera). En cuanto al P, debido a la mayor complejidad de
su ciclo y de las reacciones que ocurren en el suelo, hay un mayor desconocimiento de
los factores que influyen en su dinámica en los sistemas suelo-planta, lo que supone
nuevas oportunidades de estudio en la evaluación del uso agrícola de los residuos
orgánicos.
Teniendo en cuenta los conocimientos previos sobre cada nutriente así como las
necesidades específicas en el estudio de los mismos, en esta Tesis se han evaluado: (1)
el efecto de la aplicación de residuos procedentes de la industria vinícola y alcoholera
en la dinámica del N desde el punto de vista agronómico y medioambiental en una
zona vulnerable a la contaminación por nitratos; y (2) los factores que influyen en la
disponibilidad de P en el suelo tras la aplicación de residuos orgánicos. Para ello se han
llevado a cabo incubaciones de laboratorio así como ensayos de campo que
permitieran evaluar la dinámica de estos nutrientes en condiciones reales.
Las incubaciones de suelo en condiciones controladas de humedad y
temperatura para determinar el N mineralizado se utilizan habitualmente para estimar
la disponibilidad de N para el cultivo así como el riesgo medioambiental. Por ello se llevó
a cabo una incubación en laboratorio para conocer la velocidad de mineralización de
RESUMEN
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N de un compost obtenido a partir de residuos de la industria vinícola y alcoholera,
ampliamente distribuida en Castilla-La Mancha, región con problemas importantes de
contaminación de acuíferos por nitratos. Se probaron tres dosis crecientes de compost
correspondientes a 230, 460 y 690 kg de N total por hectárea que se mezclaron con un
suelo franco arcillo arenoso de la zona. La evolución del N mineral en el suelo a lo largo
del tiempo se ajustó a un modelo de regresión no lineal, obteniendo valores bajos de N
potencialmente mineralizable y bajas contantes de mineralización, lo que indica que se
trata de un material resistente a la mineralización y con una lenta liberación de N en el
suelo, mineralizándose tan solo 1.61, 1.33 y 1.21% del N total aplicado con cada dosis
creciente de compost (para un periodo de seis meses). Por otra parte, la mineralización
de N tras la aplicación de este material también se evaluó en condiciones de campo,
mediante la elaboración de un balance de N durante dos ciclos de cultivo (2011 y 2012)
de melón bajo riego por goteo, cultivo y manejo agrícola muy característicos de la zona
de estudio. Las constantes de mineralización obtenidas en el laboratorio se ajustaron a
las temperaturas reales en campo para predecir el N mineralizado en campo durante
el ciclo de cultivo del melón, sin embargo este modelo generalmente sobreestimaba el
N mineralizado observado en campo, por la influencia de otros factores no tenidos en
cuenta para obtener esta predicción, como el N acumulado en el suelo, el efecto de
la planta o las fluctuaciones de temperatura y humedad. Tanto el ajuste de los datos
del laboratorio al modelo de mineralización como las predicciones del mismo fueron
mejores cuando se consideraba el efecto de la mezcla suelo-compost que cuando se
aislaba el N mineralizado del compost, mostrando la importancia del efecto del suelo
en la mineralización del N procedente de residuos orgánicos.
Dado que esta zona de estudio ha sido declarada vulnerable a la
contaminación por nitratos y cuenta con diferentes unidades hidrológicas protegidas,
en el mismo ensayo de campo con melón bajo riego por goteo se evaluó el riesgo de
contaminación por nitratos tras la aplicación de diferentes dosis de compost bajo dos
regímenes de riego, riego ajustado a las necesidades del cultivo (90 ó 100% de la
evapotranspiración del cultivo (ETc)) o riego excedentario (120% ETc). A lo largo del ciclo
de cultivo se estimó semanalmente el drenaje mediante la realización de un balance
hídrico, así como se tomaron muestras de la solución de suelo y se determinó su
concentración de nitratos. Para evaluar el riesgo de contaminación de las aguas
subterráneas asociado con estas prácticas, se utilizaron algunos índices
medioambientales para determinar la variación en la calidad del agua potable (Índice
de Impacto (II)) y en la concentración de nitratos del acuífero (Índice de Impacto
Ambiental (EII)). Para combinar parámetros medioambientales con parámetros de
producción, se calculó la eficiencia de manejo. Se observó que la aplicación de
RESUMEN
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compost bajo un régimen de riego ajustado no aumentaba el riesgo de contaminación
de las aguas subterráneas incluso con la aplicación de la dosis más alta. Sin embargo,
la aplicación de grandes cantidades de compost combinada con un riego
excedentario supuso un incremento en el N lixiviado a lo largo del ciclo de cultivo,
mientras que no se obtuvieron mayores producciones con respecto al riego ajustado.
La aplicación de residuos de la industria vinícola y alcoholera como fuente de P
fue evaluada en suelos calizos caracterizados por una alta capacidad de retención de
P, lo cual en algunos casos limita la disponibilidad de este nutriente. Para ello se llevó a
cabo otro ensayo de incubación con dos suelos de diferente textura, con diferente
contenido de carbonato cálcico, hierro y con dos niveles de P disponible; a los que se
aplicaron diferentes materiales procedentes de estas industrias (con y sin compostaje
previo) aportando diferentes cantidades de P. A lo largo del tiempo se analizó el P
disponible del suelo (P Olsen) así como el pH y el carbono orgánico disuelto. Al final de
la incubación, con el fin de estudiar los cambios producidos por los diferentes residuos
en el estado del P del suelo se llevó a cabo un fraccionamiento del P inorgánico del
suelo, el cual se separó en P soluble y débilmente enlazado (NaOH-NaCl-P), P soluble en
reductores u ocluido en los óxidos de Fe (CBD-P) y P poco soluble precipitado como Ca-
P (HCl-P); y se determinó la capacidad de retención de P así como el grado de
saturación de este elemento en el suelo. En este ensayo se observó que, dada la
naturaleza caliza de los suelos, la influencia de la cantidad de P aplicado con los
residuos en el P disponible sólo se producía al comienzo del periodo de incubación,
mientras que al final del ensayo el incremento en el P disponible del suelo se igualaba
independientemente del P aplicado con cada residuo, aumentando el P retenido en la
fracción menos soluble con el aumento del P aplicado. Por el contrario, la aplicación
de materiales orgánicos menos estabilizados y con un menor contenido en P, produjo
un aumento en las formas de P más lábiles debido a una disolución del P retenido en la
fracción menos lábil, lo cual demostró la influencia de la materia orgánica en los
procesos que controlan el P disponible en el suelo. La aplicación de residuos aumentó
el grado de saturación de P de los suelos, sin embargo los valores obtenidos no
superaron los límites establecidos que indican un riesgo de contaminación de las aguas.
La influencia de la aplicación de residuos orgánicos en las formas de P
inorgánico y orgánico del suelo se estudió además en un suelo ácido de textura areno
francosa tras la aplicación en campo a largo plazo de estiércol vacuno y de compost
obtenido a partir de biorresiduos, así como la aplicación combinada de compost y un
fertilizante mineral (superfosfato tripe), en una rotación de cultivos. En muestras de suelo
recogidas 14 años después del establecimiento del experimento en campo, se
determinó el P soluble y disponible, la capacidad de adsorción de P, el grado de
RESUMEN
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saturación de P así como diferentes actividades enzimáticas (actividad
deshidrogenasa, fosfatasa ácida y fosfatasa alcalina). Las diferentes formas de P
orgánico en el suelo se estudiaron mediante una técnica de adición de enzimas con
diferentes substratos específicos a extractos de suelo de NaOH-EDTA, midiendo el P
hidrolizado durante un periodo de incubación por colorimetría. Las enzimas utilizadas
fueron la fosfatasa ácida, la nucleasa y la fitasa las cuales permitieron identificar
monoésteres hidrolizables (monoester-like P), diésteres (DNA-like P) e inositol
hexaquifosfato (Ins6P-like P). La aplicación a largo plazo de residuos orgánicos aumentó
el P disponible del suelo proporcionalmente al P aplicado con cada tipo de fertilización,
suponiendo un mayor riesgo de pérdidas de P dado el alto grado de saturación de este
suelo. La aplicación de residuos orgánicos aumentó el P orgánico del suelo resistente a
la hidrólisis enzimática, sin embargo no influyó en las diferentes formas de P hidrolizable
por las enzimas en comparación con las observadas en el suelo sin enmendar. Además,
las diferentes formas de P orgánico aplicadas con los residuos orgánicos no se
correspondieron con las analizadas en el suelo lo cual demostró que éstas son el
resultado de diferentes procesos en el suelo mediados por las plantas, los
microorganismos u otros procesos abióticos. En este estudio se encontró una correlación
entre el Ins6P-like P y la actividad microbiana (actividad deshidrogenasa) del suelo, lo
cual refuerza esta afirmación.
Por último, la aplicación de residuos orgánicos como fuente de N y P en la
agricultura se evaluó agronómicamente en un escenario real. Se estableció un
experimento de campo para evaluar el compost procedente de residuos de bodegas
y alcoholeras en el mismo cultivo de melón utilizado en el estudio de la mineralización y
lixiviación de N. En este experimento se estudió la aplicación de tres dosis de compost:
1, 2 y 3 kg de compost por metro lineal de plantación correspondientes a 7, 13 y 20 t de
compost por hectárea respectivamente; y se estudió el efecto sobre el crecimiento de
las plantas, la acumulación de N y P en la planta, así como la producción y calidad del
cultivo. La aplicación del compost produjo un ligero incremento en la biomasa vegetal
acompañado por una mejora significativa de la producción con respecto a las parcelas
no enmendadas, obteniéndose la máxima producción con la aplicación de 2 kg de
compost por metro lineal. Aunque los efectos potenciales del N y P fueron parcialmente
enmascarados por otras entradas de estos nutrientes en el sistema (alta concentración
de nitratos en el agua de riego y ácido fosfórico suministrado por fertirrigación), se
observó una mayor acumulación de P uno de los años de estudio que resultó en un
aumento en el número de frutos en las parcelas enmendadas. Además, la mayor
acumulación de N y P disponible en el suelo al final del ciclo de cultivo indicó el
potencial uso de estos materiales como fuente de estos nutrientes.
ABSTRACT
v
ABSTRACT
Nitrogen (N) and phosphorus (P) are essential nutrients in crop production. The
development of synthetic fertilizers during the 20th century allowed an intensification of
the agriculture increasing crop yields but in turn the great input of nutrients has resulted
in some cases in inefficient systems with higher losses to the environment. Regarding P,
the scarcity of phosphate rock reserves necessary for the production of phosphate
fertilizers aggravates this problem.
The use of organic wastes in agriculture as a source of N and P is a good option
of management that allows to value the large amount of wastes generated. However, it
is important to understand the processes occurring in the soil after application of these
materials, as they affect the availability of nutrients that can be used by the crop and
the nutrient losses from agricultural systems that can cause problems of contamination.
Although soil N dynamic has been more studied than P, the important concern of nitrate
pollution in Nitrate Vulnerable Zones requires the evaluation of those management
practices that could aggravate this situation, and in the case of organic wastes, the
evaluation of the agronomic and environmental response after application of materials
with a high N content (such as wastes from winery and distillery industries). On the other
hand, due to the complexity of soil P cycle and the reactions that occur in soil, there is
less knowledge about the factors that can influence its dynamics in the soil-plant system,
which means new opportunities of study regarding the evaluation of the agricultural use
of organic wastes.
Taking into account the previous knowledge of each nutrient and the specific
needs of study, in this Thesis we have evaluated: (1) the effect of the application of
wastes from the winery and distillery industries on N dynamics from the agronomic and
environmental viewpoint in a vulnerable zone; and (2) the factors that influence P
availability in soils after the application of organic wastes. With this purposes, incubations
were carried out in laboratory conditions as well as field trials that allow to assess the
dynamic of these nutrients in real conditions.
Soil incubations under controlled moisture and temperature conditions to
determine N mineralization are commonly used to estimate N availability for crops
together with the environmental risk. Therefore, a laboratory incubation was conducted
in order to determine the N mineralization rate of a compost made from wastes
generated in the winery and distillery industries, widely distributed in Castilla-La Mancha,
a region with significant problems of aquifers contamination by nitrates. Three increasing
doses of compost corresponding to 230, 460 and 690 kg of total N per hectare were
mixed with a sandy clay loam soil collected in this area. The evolution of mineral N in soil
ABSTRACT
vi
over time was adjusted to a nonlinear regression model, obtaining low values of
potentially mineralizable N and low constants of mineralization, indicating that it is a
material resistant to mineralization with a slow release of N, with only 1.61, 1.33 and 1.21%
of total N applied being mineralized with each increasing dose of compost (for a period
of six months). Furthermore, N mineralization after the application of this material was also
evaluated in field conditions by carrying out a N balance during two growing seasons
(2011 and 2012) of a melon crop under drip irrigation, a crop and management very
characteristic of the area of study. The mineralization constants obtained in the
laboratory were adjusted to the actual temperatures observed in the field to predict N
mineralized during each growing season, however, this model generally overestimated
the N mineralization observed in the field, because of the influence of other factors not
taken into account for this prediction, as N accumulated in soil, the plant effect or the
fluctuations of temperature and moisture. The fitting of the laboratory data to the model
as well as the predictions of N mineralized in the field were better when considering N
mineralized from the soil-compost mixture rather than when N mineralized from compost
was isolated, underlining the important role of the soil on N mineralization from organic
wastes.
Since the area of study was declared vulnerable to nitrate pollution and is
situated between different protected hydrological units, the risk of nitrate pollution after
application of different doses compost was evaluated in the same field trial with melon
under two irrigation regimes, irrigation adjusted to the crop needs (90 or 100% of the crop
evapotranspiration (ETc)) or excedentary irrigation (120% ETc). Drainage was estimated
weekly throughout the growing season by conducting a water balance, samples of the
soil solution were taken and the concentration of nitrates was determined. To assess the
risk of groundwater contamination associated with these practices, some environmental
indices were used to determine the variation in the quality of drinking water (Impact
Index (II)) and the nitrates concentration in the groundwater (Environmental Impact
Index (EII)). To combine environmental parameters together with yield parameters, the
Management Efficiency was calculated. It was observed that the application of
compost under irrigation adjusted to the plant needs did not represent a higher risk of
groundwater contamination even with the application of the highest doses. However,
the application of large amounts of compost combined with an irrigation surplus
represented an increase of N leaching during the growing season compared with the
unamended plots, while no additional yield with respect to the adjusted irrigation
strategy is obtained.
The application of wastes derived from the winery and distillery industry as source
of P was evaluated in calcareous soils characterized by a high P retention capacity,
ABSTRACT
vii
which in some cases limits the availability of this nutrient. Another incubation experiment
was carried out using two soils with different texture, different calcium carbonate and
iron contents and two levels of available P; to which different materials from these
industries (with and without composting) were applied providing different amounts of P.
Soil available P (Olsen P), pH and dissolved organic carbon were analyzed along time.
At the end of the incubation, in order to study the changes in soil P status caused by the
different residues, a fractionation of soil inorganic P was carried out, which was separated
into soluble and weakly bound P (NaOH-NaCl- P), reductant soluble P or occluded in Fe
oxides (CBD-P) and P precipitated as poorly soluble Ca-P (HCl-P); and the P retention
capacity and degree of P saturation were determined as well. Given the calcareous
nature of the soils, the influence of the amount of P applied with the organic wastes in
soil available P only occurred at the beginning of the incubation period, while at the end
of the trial the increase in soil available P equalled independently of the amount of P
applied with each residue, increasing the P retained in the least soluble fraction when
increasing P applied. Conversely, the application of less stabilized materials with a lower
content of P resulted in an increase in the most labile P forms due to dissolution of P
retained in the less labile fraction, demonstrating the influence of organic matter addition
on soil P processes that control P availability in soil. As expected, the application of
organic wastes increased the degree of P saturation in the soils, however the values
obtained did not exceed the limits considered to pose a risk of water pollution.
The influence of the application of organic wastes on inorganic and organic soil
P forms was also studied in an acid loamy sand soil after long-term field application of
cattle manure and biowaste compost and the combined application of compost and
mineral fertilizer (triple superphosphate) in a crop rotation. Soil samples were collected
14 years after the establishment of the field experiment, and analyzed for soluble and
available P, P sorption capacity, degree of P saturation and enzymatic activities
(dehydrogenase, acid phosphatase and alkaline phosphatase). The different forms of
organic P in soil were determined by using an enzyme addition technique, based on
adding enzymes with different substrate specificities to NaOH-EDTA soil extracts,
measuring the hydrolyzed P colorimetrically after an incubation period. The enzymes
used were acid phosphatase, nuclease and phytase which allowed to identify
hydrolyzable monoesters (monoester-like P) diesters (DNA-like P) and inositol
hexakisphosphate (Ins6P-like P). The long-term application of organic wastes increased
soil available P proportionally to the P applied with each type of fertilizer, assuming a
higher risk of P losses given the high degree of P saturation of this soil. The application of
organic wastes increased soil organic P resistant to enzymatic hydrolysis, but no influence
was observed regarding the different forms of enzyme hydrolyzable organic P compared
ABSTRACT
viii
to those observed in the non-amended soil. Furthermore, the different forms of organic P
applied with the organic wastes did not correspond to those analyzed in the soil which
showed that these forms in soil are a result of multifaceted P turnover processes in soil
affected by plants, microorganisms and abiotic factors. In this study, a correlation
between Ins6P-like P and the microbial activity (dehydrogenase activity) of soil was
found, which reinforces this claim.
Finally, the application of organic wastes as a source of N and P in agriculture
was evaluated agronomically in a real field scenario. A field experiment was established
to evaluate the application of compost made from wine-distillery wastes in the same
melon crop used in the experiments of N mineralization and leaching. In this experiment
the application of three doses of compost were studied: 1 , 2 and 3 kg of compost per
linear meter of plantation corresponding to 7, 13 and 20 tonnes of compost per hectare
respectively; and the effect on plant growth, N and P accumulation in the plant as well
as crop yield and quality was studied. The application of compost produced a slight
increase in plant biomass accompanied by a significant improvement in crop yield with
respect to the unamended plots, obtaining the maximum yield with the application of 2
kg of compost per linear meter. Although the potential effects of N and P were partially
masked by other inputs of these nutrients in the system (high concentration of nitrates in
the irrigation water and phosphoric acid supplied by fertigation), an effect of P was
observed the first year of study resulting in a greater plant P accumulation and in an
increase in the number of fruits in the amended plots. In addition, the higher
accumulation of available N and P in the topsoil at the end of the growing season
indicated the potential use of this material as source of these nutrients.
List of abbreviations
ix
LIST OF ABBREVIATIONS
ACCE Active calcium carbonate equivalent
AcP Acid phosphatase
Alox Oxalate soluble aluminium
AlP Alkaline phosphatase
Al-P Aluminium-bound phosphorus
Ca-P Calcium-bound phosphorus
CBD-P Citrate-bicarbonate-dithionite extractable phosphorus
CCE Calcium carbonate equivalent
COM Communication from the European Commission
D Drainage
DAT Days after transplanting
DHA Dehydrogenase
DOC Dissolved organic carbon
DOCM Diario Oficial Castilla-La Mancha
DPS Degree of phosphorus saturation
EC European Community
EDGAR Emission Database for Global Atmospheric Research
EDTA Ethylenediaminetetraacetic acid
EEA European Environment Agency
EGM Exhausted grape marc
EII Environmental Impact Index
ET0 Reference evapotranspiration
ETc Crop evapotranspiration
EU European Union
EUROSTAT Statistical office of the European Union
FAO Food and Agriculture Organization of the United Nations
FDR Frequency Domain Resonance
Feox Oxalate soluble iron
Fe-P Iron-bound phosphorus
FY Fruit yield
GM Grape marc
List of abbreviations
x
GS Grape stalks
HCl-P HCl extractable phosphorus
II Impact Index
Ins6-P Inositol hexakisphosphate
Irr Irrigation
K First-order rate constant of mineralization
Kc Crop coefficient
LC Lees cake
MAGRAMA Ministerio de Agricultura, Alimentación y Medio Ambiente.
ME Management Efficiency
Mt Mega tonnes
N Nitrogen
NO3 Nitrate
NH4 Ammonium
N0 Potentially mineralizable nitrogen
NaOH-NaCl-P NaOH-NaCl extractable phosphorus
Nap Nitrogen applied
Nc Nitrogen applied with compost
Nl Nitrogen leaching
NM Nitrogen mineralized
Nr Reactive nitrogen.
Ns final Mineral nitrogen in the soil at the end of the crop cycle
Ns initial Mineral nitrogen in the soil at the beginning of the crop cycle
Nup Nitrogen uptake
Nw Nitrogen applied with irrigation water
P Phosphorus
Pdl Double lactate extractable phosphorus
Pi Inorganic phosphorus
31P-NMR 31P nuclear magnetic resonance
Po Organic phosphorus
Pox Oxalate soluble phosphorus
Prec Precipitation
PSC Phosphorus sorption capacity
List of abbreviations
xi
PSI Phosphorus sorption index
Pt Total phosphorus
Pw Water extractable phosphorus
Rf Runoff
SOM Soil organic matter
Ta Average air temperature
TMa Average maximum air temperature
Tma Average minimum air temperature
TPF Triphenylformazan
TSP Triple superphosphate
UH Hydrological unit
WL Wine lees
WSOC Water soluble organic carbon
Ѳν Volumetric soil water content
List of tables
xiii
LIST OF TABLES
Table 1.1. Main characteristics of the solid winery and distillery wastes and
by-products (Bustamante et al., 2008)………………………………................6
Table 1.2. Main forms of organic P in soil and its abundance (modified from
Condron et al., 2005). ………………………………………………....................12
Table 1.3. Methods for the determination of soil nitrogen………………………………18
Table 1.4. Methods for the determination of soil inorganic phosphorus……………..19
Table 1.5. Methods for the determination of soil organic phosphorus………………..21
Table 1.6. Kjeldahl N, total P concentration and N:P ratio in a wide range of organic
wastes with different origin and treatment (modified from García-
Albacete et al.,2012)…………………………….………………………………...24
Table 3.1. Characterization of the soil and compost used in the experiments……..38
Table 3.2. Net N mineralization (%) with the different doses of compost during the
incubation……………………………………………………………………………45
Table 3.3. N mineralization constants……………………………………………………….46
Table 3.4. Parameters for the N balance during the field experiment in 2012 and
2012……………………………………………………………………………………47
Table 3.5. Predicted values of N mineralized (NM) and observed values in field
plots……………………………………………………………………………………50
Table 4.1. The physicochemical characteristics of the soil (Ap horizon) at the field
experimental site……………………………………………………………………58
Table 4.2. Monthly climatic data of the experimental area during the growing season
of 2011 and 2012……………………………………………………………………59
Table 4.3. Characteristics of the compost used in 2011 and 2012…………………….59
Table 4.4. Characteristics of the irrigation water used in 2011 and 2012………….….60
Table 4.5. Fruit yield, N leached, management efficiency (ME) and N uptake in the
melon crop in 2011 and 2012…………………………………………………..…71
Table 5.1. Organic materials used in the incubation experiment……………………...78
Table 5.2. Physicochemical properties of the experimental soils………………………82
Table 5.3. Chemical properties of the organic materials………………………………..82
Table 5.4. ANOVA of the changes in pH (ΔpH), dissolved organic C (ΔDOC) and
Olsen P (ΔOlsen P) concentrations in soils treated with the organic
materials as affected by waste, soil type and time………………………….84
Table 5.5. Recovery of applied P (%) over time in soils treated with the different
organic materials…………………………………………………………………...87
Table 5.6. Correlation coefficients between the increment in Olsen P (ΔOlsen P) and
some selected properties of the organic materials at each incubation
interval in amended soil…………………………………………..……………....88
List of tables
xiv
Table 5.7. ANOVA of the changes in the inorganic P fractions (ΔNaOH-NaCl, ΔCBD-
P, ΔHCl-P), P sorption index (ΔPSI) and degree of P saturation (ΔDPS) in
soils treated with the organic materials as affected by waste and soil type
at the end of the incubation period………………………….…………….…..91
Table 5.8. Olsen P, P sorption index (PSI) and degree of P saturation (DPS) in soils
treated with the different organic materials at the end of the incubation
period…………………………............................................................................92
Table 6.1. Total P supply, total P uptake and total P balance accumulated after 14
years in field experiment…………………………………………………………100
Table 6.2. Properties of the topsoil collected in 2012 after 14 years of different P
amendment management and at the beginning of the experiment in
1998…………………………………………………………………………………..106
Table 6.3. Soil enzyme activities in soil samples collected in autumn 2012 after 14
years of different P amendment management…………………………….107
Table 6.4. Characteristics of the organic amendments applied in 2010……………109
Table 6.5. P classes in NaOH-EDTA soil extracts (0-30 cm) collected in autumn 2012
after 14 years of different P amendment management………………….112
Table 7.1. The physicochemical characteristics of the soil (Ap horizon) in the plots
used in 2011 and 2012 at the field experimental site………………………120
Table 7.2. Characteristics of the compost used in 2011 and 2012…………………...120
Table 7.3. Dry matter production by the aboveground melon plant organs and leaf
area index (LAI) at the middle of the growth cycle on 2011 and 2012 (48
and 47 DAT respectively)………………………………………………………...124
Table 7.4. Dry matter production by the aboveground melon plant organs and leaf
area index (LAI) at the end of the growth cycle of 2011 and 2012 (90 and
89 DAT respectively)………………………………………………………………125
Table 7.5. Fruit yield and quality data of the growing seasons of 2011 and
2012..................................................................................................................129
List of figures
xv
LIST OF FIGURES
Figure 1.1. Simplified diagram of the products generated during the winemaking
processes (translated from Bustamante, 2007)…………………………………5
Figure 1.2. Global recovery of N and P in crop production for 1900, 1950, 1970, 2000,
2050 (estimations) (Bouwman et al., 2013)………………………………………7
Figure 1.3. A preferred scenario for meeting long-term global phosphorus demand
(Cordell and White, 2011)…………………………………………………………..8
Figure 1.4. Squeme of major processes of N cycling (modified from Rennenberg et
al., 2009)………………………………………………………………………………..9
Figure 1.5. P dynamics in the soil-rizosphere continuum (modified from Shen et al.,
2011)…………………………………………………………………………………..11
Figure 1.6. Simplified summary of the nitrogen cascade illustrating losses,
transformations and effects of reactive nitrogen fertilizers and the
environment (Sutton et al., 2011)………………………………………………...13
Figure 1.7. Nitrate Vulnerable Zones in Europe (Joint Research Centre)………………14
Figure 1.8. Ecological water quality in EU member states (COM, 2012)………………..15
Figure 1.9. Fate and transport of P inputs in agricultural systems (Shigaki et al.,
2006)…………………………………………………………………..………………16
Figure 1.10. Principle of the enzyme additions assays………………………………………22
Figure 2.1. Experiments, variables studied and factors considered to achieve the aims
proposed in this Thesis………………………………………………………………31
Figure 3.1. Compost pile (Agricompost S.L.)…………………………………………………38
Figure 3.2. Incubation pots and soil samplings………………………………………………39
Figure 3.3. Details of compost application in the field……………………………………..40
Figure 3.4. Experimental plots at plantation and development of the melon crop….41
Figure 3.5. Evolution of inorganic N (a) nitrate-N (b) and ammonium-N (c) during 24
weeks of incubation………………………………………………………………..44
Figure 3.6. Average air temperature Ta (a); average maximum air temperature TMa
(b); average minimum air temperature Tma (c) during the field experiment
in 2011 and 2012…………………………………………………………………….48
Figure 3.7. Comparison between estimated and measured mineralized N from the
compost or from the soil-compost mixture……………………………………..51
Figure 4.1. Irrigation pond located at the experimental field station…………………..61
Figure 4.2. Placement of the Diviner probe and the suction cup in the plots (left) and
detail of the suction cup (right)………………………………………………….62
Figure 4.3. Cumulative crop evapotranspiration (ETc), estimated irrigation, actual
irrigation and drainage (D) of each dose of compost with various irrigation
rates: (a) 120% ETc, 2011; (b) 90% ETc, 2011; (c) 100% ETc, 2012…………..64
List of figures
xvi
Figure 4.4. NO3 concentration in the soil solution (a, b, c) and cumulative N leaching
(d, e, f) of each dose of compost with various irrigation rates: (a, d) 120%
ETc, 2011; (b, e) 90% ETc, 2011; (c, f) 100% ETc, 2012…………………………66
Figure 4.5. Impact Index (II) (a) and Environmental Impact Index (EII) (b) versus total
N applied with compost. The discontinuous line represents the maximum
limit equal to 50 mg L-1 established by the Drinking Water Directive (a) or a
NO3 concentration equal to the concentration of NO3 in the groundwater
(b)………………………………………………………………………………...……69
Figure 5.1. Incubation chamber used in the experiment………………………………….79
Figure 5.2. Evolution of pH in soils treated with the different organic materials during
the incubation period……………………………………………………………...83
Figure 5.3. Evolution of dissolved organic C in soils treated with the different organic
materials during the incubation period…………………………………………85
Figure 5.4. Evolution of Olsen P in soils treated with the different organic materials
during the incubation period……………………………………………………..86
Figure 5.5. Inorganic P fractions given by the sequential fractionation scheme of Kuo
(1996) in soils treated with the different organic materials at the end of the
incubation period…………………………………………………………………..90
Figure 6.1. Detail of the enzymatic incubation assay and the incubator used…….102
Figure 7.1. Melon crop at 40 DAT (left) and detail of the fruit set (right)………………121
Figure 7.2. Evolution of daily mean temperature and rainfall during the crop seasons
of 2011 and 2012…………………………………………………………………..123
Figure 7.3. Nitrogen accumulation in leaf, stem, fruit and in the whole plant at the
middle and at the end of the growing cycle…………………………………127
Figure 7.4. Phosphorus accumulation in leaf, stem, fruit and in the whole plant at the
middle and at the end of the growing cycle…………………………………128
Figure 7.5. Available nitrogen and phosphorus in soil before compost addition (Initial)
and at the end of the melon growing season for the different compost
doses……………………………………………………………………………..….132
CHAPTER 1:
INTRODUCTION
CHAPTER 1
3
1.1. ORGANIC WASTE GENERATION AND MANAGEMENT IN THE UE.
Waste generation and management poses a serious environmental challenge for the
modern society. The increasing population and its concentration in urban areas together
with the concentration of farming activities and the industrial development involve the
generation of larger amounts of wastes. In 2012, a total of 2.5 billion tonnes of wastes
were generated in the European Union (EU 28) (EUROSTAT, 2012).
According to the European Environment Agency (EEA, 2003), biowaste and
agricultural waste are included in the main waste streams in the EU. Each year, about
118 to 138 million of tonnes of biowastes are generated in the UE, and predictions stated
that these numbers will increase by 10% until 2020 (COM, 2010). The Waste Framework
Directive (2008/98/EC) defines biowaste as `biodegradable garden and park waste,
food and kitchen waste from households, restaurants, caterers and retail premises and
comparable waste from food processing plants´. Approximately 40% of the biowastes
produced in the UE are landfilled, and this proportion reaches the 100% in some countries.
In order to avoid landfilling and incineration, UE strategies regarding biowaste
management are focused to prevent its generation and promote the re-use and
recycling (2008/98/EC). The Waste Framework Directive encouraged the separate
collection of biowaste and the treatment by composting and digestion to produce
environmentally safe materials. Accordingly, the flagship initiative for a Resource Efficient
Europe (COM, 2011), within the Europe 2020 Strategy, aims to use wastes as resources.
Agricultural organic wastes include animal excreta, plant residues, soiled water
and silage effluent (EEA, 2013). In 2012, the generation of vegetal and animal wastes
exceed 100 million tonnes (EUROSTAT, 2012). Among these wastes, those coming from
livestock uses are especially relevant since about 70% of the agricultural land in the UE is
Introduction
4
dedicated to animal production (Oenema, 2012). The increasing consumption of animal
proteins in the last decades has enhanced the number of animals for livestock
production. Consequently, the intensification and specialization of animal production in
the recent decades have been difficulted the management of these wastes. However,
the implementation of the European Community Nitrates Directive (Council Directive
91/676/EEC) has promoted the improvement of manure management for agricultural
use. On the other hand, in plant production (horticulture, fruit, wine and grass sectors),
important amount of wastes are generated as well at the farm and post-harvest levels,
including processing and marketing units (e.g. wineries).
1.1.1. A specific case of Mediterranean countries: wastes derived from the winery
and distillery industries.
Although the wine production is distributed worldwide, the European Union concentrates
over 50% of the world production, being this industry mainly located in Mediterranean
countries as France, Italy and Spain (FAO, 2013). The wine industry is one of the most
important sectors of the agro-food industry of Spain, with a total production per year of
3.2 million tonnes of wine (FAO, 2013). Additionally, Spain holds the largest area under
vines in the world (International Organization of Vine and Wine, 2014).
Problems related to the waste generation in the wine industry are an especial
concern in a very short period of time, especially during the harvesting, which takes place
between September and October. During the winemaking, different wastes and by-
products are generated at different stages of the process (Figure 1.1).
The main solid wastes and by-products generated in the wineries are the grape
stalks, grape marc and wine lees. The grape stalks are generated during the crushing-
destemming process, and are constituted by the bunch peduncle and the support
structure of the grapes. The next waste produced is the grape marc, which is formed by
the skins, pulp and pips generated during the pressing process carried out to obtain the
must. The wine lees are the residue deposited at the bottom of the fermentation tanks
during wine fermentation, obtained after the racking process. The grape marc and wine
lees are considered as by-products as they are sent to the alcohol distilleries to extract
the alcohol and tartrates presented in these materials, producing other solid wastes:
exhausted or spent grape marc and lees cake. Additionally, the laundering operations
carried out during the vinification processes generates large volumes of winery
wastewater, while the distillation of wine, grape marc and wine lees resulted in an effluent
known as vinasse (Torrijos and Moletta, 2000).
CHAPTER 1
5
Figure 1.1. Simplified diagram of the products generated during the winemaking
processes (translated from Bustamante, 2007).
There are different alternatives to treat and recovery the wastes generated in the
wineries and distilleries (Arvanitoyannis et al., 2006). Traditionally, these wastes were
incorporated to vineyard soils without any previous treatment. The use of these raw
materials in agriculture has been studied in the past, specifically the application of grape
marc (Sorlini et al., 1998) and winery sludge (Mariotti et al., 2000; Masoni et al., 2000).
According to Bustamante et al. (2008a), the solid winery and distillery wastes are
characterized by a low pH and electrical conductivity, notable contents of nitrogen,
phosphorus and potassium, and a lower content of heavy metals compared with other
organic wastes as manures or urban wastes (Table 1.1). Otherwise, these wastes contain
soluble phenolic compounds that can cause phytotoxicity problems. For this reason, the
composting of these wastes has been found as a feasible alternative of recycling these
wastes in agriculture (Ferrer et al., 2001; Bertrán et al., 2004; Bustamante et al., 2008b).
Introduction
6
Table 1.1. Main characteristics of the solid winery and distillery wastes and by-products
(Bustamante et al., 2008a).
Characteristics Grape
stalk
Grape
marc
Wine
lees
Exhausted
grape marc
Winery-
sludge
pH 4.4 3.8 4.0 5.5 6.8
EC (dS m-1) 4.44 3.40 5.59 1.62 6.15
Organic matter (g kg-1) 920 915 759 912 669
Organic C (g kg-1) 316 280 300 276 257
Water-soluble C (g kg-1) 74.5 37.4 87.8 18.3 26.9
N (g kg-1) 12.4 20.3 35.2 21.3 44.3
P (g kg-1) 0.94 1.15 4.94 1.64 10.9
K (g kg-1) 30.0 24.2 72.8 11.9 20.7
Ca (g kg-1) 9.5 9.4 9.2 14.6 81.6
Mg (g kg-1) 2.1 1.2 1.6 1.2 4.6
Fe (mg kg-1) 128 136 357 370 3128
Mn (mg kg-1) 25 12 12 17 91
Cu (mg kg-1) 22 28 189 23 262
Zn (mg kg-1) 26 24 46 19 227
Water-soluble
polyphenols (g kg-1) 19.0 2.6 7.5 1.6 1.2
1.2. TOWARDS A SUSTAINABLE USE OF NITROGEN AND PHOSPHORUS IN
AGRICULTURE.
Nitrogen (N) and phosphorus (P) are key nutritional elements for plant growth and thus
limiting nutrients in crop production. Up to the beginning of the 20th century, N and P
plant needs were covered with natural levels of these elements in soil as a result of the
addition of organic matter, or via N fixation by legumes (Bouwman et al., 2013).
Afterwards, the need to increase crop production in order to sustain the growing
population led to the development of synthetic fertilizers production.
The invention of the Haber-Bosch process to produce artificial N fertilizers
represented an increase of relatively inexpensive fertilizers which contributed to the
intensification of the agriculture (Smil, 1996). In 1950, global annual consumption of N
fertilizers was less than 4 Mt, but increased to 32 Mt by 1970 and to greater than 80 Mt by
the 1990s (Roy and Hammond, 2004). This great effort of geo-engineering has more than
double the production of N compared with pre-industrial levels (Galloway et al., 2008)
thus it is only through synthetic fertilizers that humanity has been able to reach 6 million
people (Erisman et al., 2008).
CHAPTER 1
7
However, regarding P, the problem lies in the source to produce chemical
fertilizers, rich phosphate rocks which are located in very few parts of the world and are
a finite resource (Cordell and White, 2011). Currently, total P reserves are estimated at
67000 Mt whereas mining production in 2013 was of 220 Mt (Schoumans et al., 2015).
Approximately 85% of the remaining rock phosphate reserves are located in Morocco
and Western Sahara which means that geopolitical changes may increase the price of
phosphate fertilizers and consequently the food (Cordell and White, 2011). In addition,
the increasing population together with the changes in the diet are exerting greater
pressure on the demand for P, to increase agricultural production and to meet the food
demand. The total consumption of phosphate fertilizers in the world was quadrupled
during the period 1961-2012 (Schoumans et al., 2015) and the future demand is
forecasted to increase (FAO, 2006).
This change from low-input systems that relied on natural inputs to others that
highly relies on synthetic fertilizers has led to a decrease of nutrient recovery in crop
systems in the last decades (Figure 1.2). The low efficiency of the agricultural practices
have resulted in a large surplus of N and P and in increasing losses to the environment
(Erisman et al., 2011; Withers et al., 2015). Less than half of the total nutrients applied as
inorganic or organic fertilizers is efficiently used for cultivation, while the rest is dissipated
to the environment creating a complex net of impacts occurring through air, land and
water that threaten our global environment (Erisman et al., 2007; Galloway et al., 2008).
Figure 1.2. Global recovery of N and P in crop production for 1900, 1950, 1970,
2000, 2050 (estimations) (Bouwman et al., 2013).
NA: North America, SCA: South and Central America, EUR: Europe, AFR: Africa, NAS: North Asia,
SAS: South Asia, OCE: Oceania, WRLD: world, IND: industrialiced countries, DEV: developing
countries.
There are therefore pressing environmental and resource justifications for
increasing the sustainability of N and P in agricultural systems. The scientific communities
and policy makers are making a great effort focused on quantifying and managing the
environmental impacts derived from the increasing N inputs (Sutton et al., 2011).
Introduction
8
Regarding P, although it exists a debate about the depletion timeline of the phosphate
rock reserves (estimates range from 30 to 300 years) (Fixen, 2009; Smit et al., 2009; Van
Kauwenbergh, 2010); there is no dispute about the inefficient use of P in the food chain
which is a threat to the future food security which requires new sustainable P use initiatives
(Figure 1.3). Recently, the scientific community have proposed a package of strategies
in order to close the P cycle in Europe, the 5R strategy (Withers et al., 2015): Realign P
inputs, Reduce P losses to waters, Recycle P in bio-resources, Recover P from waste and,
if necessary, Redefine our food system.
Figure 1.3. A preferred scenario for meeting long-term global phosphorus demand
(Cordell and White, 2011).
1.3. UNDERSTANTING SOIL NUTRIENT CYCLES: A KEY FACTOR.
The development of sustainable land management practices for all agroecosystems
requires a fundamental understanding of the nutrients cycling and processes in soil that
may affect the availability to plants and the losses to the environment. While the global
geochemical N cycling is dominated by microbial processes in soils, P cycling is rather
complicated as it is influenced by different processes involving the inorganic and organic
solid phases, the biological activity and the chemistry of the soil solution.
1.3.1. Soil nitrogen cycle.
Nitrogen cycling in soil is characterized by a variety of N transformations involving both
inorganic (ammonium and nitrate) and organic species and uptake/immobilization of N
by microbes and plants, as shown in Figure 1.4. About 90% of N in the surface layer of
P RESUSE &
SUPPLY
MEASURES
P
EFFICIENCY
MEASURES
CHAPTER 1
9
most soils is in organic form (Bremner, 1965) which occurs as consolidated amino acids or
proteins, or free amino acids, amino sugars and other complexes. Inorganic N in soils is
predominantly in form of ammonium (NH4) and nitrate (NO3).
Figure 1.4. Squeme of major processes of N cycling
(modified from Rennenberg et al., 2009).
*Green lines represent plant processes, orange lines represent microbial processes and blue
dashed lines represent hydrological transport pathways
Litter production and decomposition are the major drivers of the N turnover.
During the decomposition process, extracellular enzymes of fungi and bacteria cleave
the large polymers of organic matter to largely bioavailable monomers which are
accessible for plants and microbes. Microorganisms can further degrade the organic
monomers to form ammonium (mineralization/ammonification). Depolymerizing and
ammonifying heterotrophic microbes are supposed to be C-limited which illustrates the
tight coupling of the C and N cycles.
Ammonium can also be oxidized to nitrate (nitrification). Ammonium and nitrate
can be taken up by plants or immobilized by microorganisms. Nitrification has been
found to be a key process in N cycling because it increases the probability of N losses by
leaching along hydrological pathways since the end products (nitrite/nitrate) are
susceptible to losses by leaching along hydrological pathways or further reduction to
Introduction
10
gaseous NO, N2O and N2 via denitrification. Nitrification is generally an aerobic process
thus it is affected by soil water content (optimum 30%-40% of the water-filled pore space
(Bouwman, 1998)), but also by soil pH (optimum pH 5.5-6.5 (Machefert et al., 2002)) and
temperature (Machefert et al., 2002). Otherwise, ammonium can also be fixed by clays
reducing its availability for the nitrification process (Sánchez-Martín et al., 2008).
Nitrogen is crucial for the growth and activity of heterotrophic microorganisms
thus ammonium and nitrate and various organic N species can be taken up by bacteria
and fungi, process known as N immobilization. N assimilation has been shown to be
dependent on available C (Azam et al., 1998; Trinsoutrot et al., 2000), since C stimulate
the growth and activity of these microorganisms.
Denitrification is an anaerobic bacterial process by which nitrate is reduced to
nitrite and further reduced to nitrous oxide (N2O) or dinitrogen (N2) gas. This occurs
generally with extensive moisture and is strongly influenced by soil, plant and
environmental factors increasing with higher nitrate and ammonium concentrations,
temperature and readily decomposable compounds. The other process of gaseous loss
from soil to the atmosphere is volatilization, the loss of ammonia, which is influenced by
soil, environmental and management factors.
Plants predominantly take up inorganic N forms (ammonium and nitrate) to cover
their demand of N for growth (Harrison et al., 2007). However, in N limiting ecosystems,
aminoacids and organic monomers are readily taken up by plants. Nitrogen acquisition
by plants depends on the plant species and on the respective growth potential but also
on other factors such as temperature (which affects root activity as well as N availability),
soil pH and competition by microbes (Jackson et al., 2008).
Nitrogen cycling is strongly affected by the human management. Fertilization,
irrigation, drainage, tillage practices, amendments, soil rotation, soil compaction, grazing
or human induced N deposition can alter the ecosystems having effects on substrate
availability, aeration, microbial community composition, soil moisture or even on
temperature profiles (Butterbach-Bahl et al., 2011).
1.3.2. Soil phosphorus cycle.
Inorganic orthophosphate (HPO42-, H2PO4-) in the soil solution is the primary source for P.
However, the equilibrium concentration of orthophosphate in the soil solution is low (<5
µM), so orthophosphate removed by plant or microbial uptake is continually replenish
from the solid-phase to sustain plant growth (Wild, 1988). This involves a combination of
desorption and dissolution of inorganic P and mineralization of organic P (Figure 1.5).
CHAPTER 1
11
Figure 1.5. P dynamics in the soil-rizosphere continuum
(modified from Shen et al., 2011).
Inorganic P in soil usually ranges from 35% to 70% of total P (Harrison, 1987). Primary
P minerals as apatite are very stable and the release of available P from these minerals
is generally too slow. However, secondary P minerals as calcium (Ca), iron (Fe) and
aluminion (Al) phosphates vary in their dissolution rates depending on the size of mineral
particles and soil pH (Pierzynski et al., 2005; Oelkers and Valsami-Jones, 2008). Phosphorus
sorption and desorption on clays and Al or Fe oxides also equilibrate with P in the soil
solution. These forms of soil co-exist in a complex equilibrium and represent very stable,
sparingly available and plant available P pools. In acidic soils, P availability in the soil
solution is primarily controlled by P adsorption by Al/Fe oxides and hydroxides such as
gibbsite, hematite and goethite (Parfitt, 1989). In neutral to calcareous soils, P retention is
generally dominated by precipitation reactions (Lindsay et al., 1989) although adsorption
Introduction
12
reactions on the surface of Ca carbonates (Larsen, 1967) and clay minerals (Devau et
al., 2011) can also occur.
Organic P forms in soil generally represents between 30% and 65% of total P
(Harrison, 1987) and have been reported to contribute significantly to plant P nutrition
(Firsching and Claassen, 1996; Oehl et al., 2001; Chen et al., 2002). Soil organic P
originates from animal and plants remains, and is synthesized by soil organisms (Condron
et al., 2005). Beyond the organic P forms are included orthophosphate monoesters (sugar
phosphates, phosphoproteins, mononucleotides and inositol phosphates),
orthophosphate diesters (phospholipids, nucleic acids) and phosphonates
(predominantly 2-aminoethyl phosphonic acid). Orthophosphate monoester are the
dominant form of organic P in most soils, mainly in the form of inositol phosphates while
phosphodiesters and phosphonates usually account for less than 10% of organic P.
(Table 1.2).
Table 1.2. Main forms of organic P in soil and its abundance
(modified from Condron et al., 2005)
The turnover of organic P in soil is primarily determined by the immobilization or
mineralization rates. Phosphorus mineralization is controlled by a combination of
biological and biochemical factors. The biological mineralization of organic C
determines de solubility of organic P which in turn controls the rate of biochemical
hydrolysis by extracellular phosphatase enzymes produced by plant roots, mycorrhizae
and microorganisms (Magid et al., 1996; Gressel and McColl, 1997; Joner et al., 2000). The
mineralization of soil organic P is also determined by its chemical nature and reactivity.
Orthophosphate monoesters (especially inositol hexakisphosphate) can be adsorbed by
ligand exchange to the same sites as orthophosphate (Celi et al., 1999) which limits its
susceptibility to mineralization (Stewart and Tiessen, 1987; Gressel et al., 1996). On the
Organic P form Concentration range Reference
(% of total organic P)
Orthophosphate monoesters
Inositol phosphates 1-100 Turner et al. (2002a)
Phosphoproteins Trace Stevenson and Cole (1999)
Mononucleotides Trace Stevenson and Cole (1999)
Sugar phosphates Trace Stevenson and Cole (1999)
Orthophosphate diesters
Nucleic acids 0-2 Magid et al. (1996)
Phospholipids 0-5 Magid et al. (1996)
Teichoic acid 0-20 Guggenberger et al. (1996)
Makarov et al. (2002)
Phosphonates 0-12 Cade-Menun et al. (2000)
CHAPTER 1
13
contrary, orthophosphate diesters are relatively soluble in soil and thus are degraded
rapidly (Magid et al., 1996).
1.4. NITROGEN, PHOSPHORUS AND THE ENVIRONMENT.
The low use-efficiency of nitrogen and phosphorus in agriculture has resulted in losses to
the environment with different impacts on human health, biodiversity and climate
change.
The `nitrogen cascade´, which includes the different nitrogen forms and its
associated environmental impacts is presented in Figure 1.6 (Sutton et al., 2011).
Figure 1.6. Simplified summary of the nitrogen cascade illustrating losses,
transformations and effects of reactive nitrogen fertilizers in the environment
(Sutton et al., 2011).
As seen in Figure 1.6, there are different environmental impacts associated with
each of these N forms, being some of these environmental problems local (soil and
groundwater pollution), or including the regional to global scales (gas emissions).
Nitrate leaching has been found the major N loss mechanism from agricultural
soils in humid climates and in irrigated cropping systems (Tisdale et al., 1999). Nitrate
pollution of groundwater poses a recognized risk for its use as drinking water, while
Introduction
14
eutrophication of surface water can lead to algal growth, oxygen deficiencies and fish
death. According to the EEA (2009), although N concentrations remained relatively
constant in lakes and even slightly decreases in rivers, the groundwater nitrates
concentration have remained high in some regions of Europe. Groundwater is an
important resource in Europe, providing water for domestic use for about two third of the
population but it is a finite and slowly renewed resource and overexploitation associated
with degradation of water quality is putting in danger an important source of drinking
water (Grizzeti et al., 2011). The European Drinking Water Directive (Council Directive
98/83/EC) stablishes the limit of nitrates allowed in drinking water in 50 mg L-1, but
recommends not to exceed 25 mg L-1.
The Nitrates Directive (91/676/EEC) aims to control N losses and requires Member
States to identify areas contributing to N pollution of groundwater and surface water,
establishing a threshold value of nitrates concentration of 50 mg L-1 to protect water
bodies. In these agricultural areas the application of fertilizers should balance the need
of the crops and manure application should not exceed 170 kg N ha-1. According to the
last assessment report on the implementation of the Nitrates Directive (COM, 2013), in the
period 2008-2011, 14.4% of groundwater stations monitored in EU27 exceeded 50 mg L-1
and 5.9% were between 40 and 50 mg L-1.
Figure 1.7. Nitrate Vulnerable Zones in Europe
(Joint Research Centre, European Commission
http://fate-gis.jrc.ec.europa.eu/geohub/MapViewer.aspx?id=2)
CHAPTER 1
15
Exceedances of the nitrate standards is a common problem across Europe,
particularly from shallow wells. For example, in Castilla-La Mancha (central Spain), where
seven areas have been declared zones vulnerable to nitrate pollution (DOCM Nº32,
2010), during a monitoring period carried out from 2005 to 2009 in the water supply and
in the wells located in agricultural fields, 43% of the measurements exceeded the limit
established at 50 mg L-1 and this percentage was of 86% in the highest contaminated
zone (Denia, 2010). This situation came from the irrigated agriculture, the inadequate use
of N fertilizers in this region (mainly due to excess applications and an incorrect moment
and techniques of application) and the filtrations from livestock farms.
The nitrogen released to the atmosphere consists mainly on ammonia (NH3),
nitrous oxide (N2O) and nitric oxide (NOx). The emissions of NH3, together with O2 and H2O
result in a secondary pollutant (HNO3) which is implied in acid rain resulting in the
acidification of soil and water. The agricultural sources of NH3 includes animal houses and
manure storages, application of manures and mineral fertilizers to soil, grazing animals,
and other sources as plants (Hertel et al., 2011). Nitrous oxide poses an important risk of
global warming since it is a powerful greenhouse gas and contributes to the ozone layer
depletion. Nitric oxide contributes indirectly to the ozone layer depletion and it is also
implicated in the acid rain. In Europe, 65% of these emissions are associated with
ecosystems denitrification (EDGAR, 2010).
Figure 1.8. Ecological water quality in EU member states (EEA, 2012).
Introduction
16
The environmental significance of P lies in its dominant role in the eutrophication
of aquatic ecosystems, particularly in lakes, as it is normally the limiting nutrient for algal
growth, as the other major nutrients required for eutrophication (N and C) can be
obtained directly from the atmosphere. Eutrophication increases the primary production
producing an imbalance between consumption and production of plants and algae
with negative effects on the species diversity and water quality. According to current
computations, the percentage of the European natural area at risk of eutrophication in
2000 was 77% (Dise et al., 2011). Although the ecological status of surface waters has
been improved over the last decades it is still poor in many European lakes and water
pathways (Kristensen, 2012). Currently there is no European Directive or other regulation
concerning P use in agriculture, however, some European countries or regions have
implemented legislations restricting P fertilization (e.g. Belgium, Germany, Netherlands)
(Amery and Schoumans, 2014).
Figure 1.9. Fate and transport of P inputs in agricultural systems (Shigaki et al., 2006).
The transfer of P from agricultural land can occur through a variety of flow
pathways, being mainly overland flow (runoff) and subsurface flow (leaching).
Phosphorus losses by runoff depend on many factors, including the concentration and
forms of P in soils, the runoff volume, the amount and solubility of P present in mineral or
organic fertilizers as well as the method of application (Sharpley et al., 1994; Withers et
al., 2001). Phosphorus leaching has been traditionally considered insignificant due to the
high P affinity to Fe and Al oxides, clay minerals and carbonates in soil. However, more
recent studies indicated that leaching losses should be considered (Sims et al., 1998;
Hooda et al., 2000), especially in waterlogged soils (Khalid et al., 1977), in acidic sandy
soils under heavy fertilization (Breeuwsma and Silva, 1992) or high in organic matter
CHAPTER 1
17
content (Fox and Kamprath, 1971). Phosphorus leaching losses are dependent on soil
characteristics (P sorption capacity, preferential flow pathways, reducing conditions),
agronomic management (e.g. amount and forms of P added, liming) and climatic
conditions (Elliot et al., 2002; Delgado and Torrent, 2000; Jensen et al., 2000).
1.5. METHODOLOGIES TO ASSESS NITROGEN AND PHOSPHORUS IN SOILS.
As previously pointed, it is important to know in which forms are the nutrients in soil in order
to evaluate the agricultural practices either agronomically or environmentally. The
determination of the inorganic forms is more frequent as these forms are the main
responsible for crop nutrition and environmental pollution. However, the importance of
characterize the organic forms is that these can be converted to available forms through
microbial processes.
1.5.1. Nitrogen.
1.5.1.1. Inorganic N.
The determination of these forms is usually taken as an estimation of available N in
agricultural soils. Additionally, soil incubation methods are usually employed to quantify
the mineralizable pool of soil organic N thus assessing soil N availability (Standford and
Smith, 1972; Chae and Tabatabai, 1986), by determining inorganic N forms periodically.
Nitrate is water soluble and a number of solutions including water have been used
as extractants (Table 1.3). Exchangeable NH4 is defined as NH4 that can be extracted at
room temperature with a neutral potassium salt solution. The most common extractant
for NO3 and NH4 is 2 M KCl (Maynard and Kalra, 1993).
1.5.1.2. Total and organic N.
Soil total N determination can be performed by wet oxidation (Kjeldahl method) or dry
combustion (Dumas method) (McGill and Figueiredo, 1993). Wet-oxidation techniques
are the most common and involve conversion of organic and inorganic N to NH4 and its
subsequent measurement. However, N determination by Kjeldahl digestion does not
include all NO3 forms. Determination of the chemical forms of soil organic N involves a
two-step procedure. First the soil is hydrolyzed by a hot mineral acid and secondly the
Introduction
18
hydrolyzate is either distilled or chromatographically separated for identification of
common aminoacids (Chae, 1993).
Table 1.3. Methods for the determination of soil nitrogen.
Form of N Procedure
Total N Wet-oxidation (Kjeldahl)
Dry-combustion (Dumas)
Organic N Amino-acids
Soil hydrolysis and distillation Amino-sugars
Inorganic N
NO3 Soil extraction
Water
Saturated (0.35%)CaSO4.2H2O
0.03 M NH4F
0.015 M H2SO4
0.01 M CaCl2
0.5 M NaHCO3 (pH 8.5)
0.01 CuSO4
0.01 M CuSO4 containing Ag2SO4
2 M KCl
Exchangeable
NH4 Soil extraction
0.05 M K2SO4
0.1 M KCl
1.0 M KCl
2 M KCl
1.5.2 Phosphorus.
1.5.2.1 Inorganic P.
a) Soil P tests.
A soil test is generally defined as a rapid extraction and analysis to assess the plant
available nutrient status of a soil. In order to estimate P availability to a crop over a
growing season, a soil test must extract that portion of soil P which is involved in buffering
the soil solution P intensity, called the labile soil P (Beegle, 2005). A variety of soil test
extractants have been developed based on fundamental understanding of the forms of
P in soil which make up the labile P for plant uptake. Many of them employ acids to
dissolve Ca, Al and Fe phosphates determined to be the main inorganic sources of labile
P (Bray and Kurtz, 1945; Mehlich, 1984). Others use buffered alkaline solutions (primarily in
calcareous soils) (Olsen et al., 1954). Besides the standard soil extractions, there are other
alternative tests as water-extractable P, 0.01 M CaCl2 (Houba et al., 1990), resin-
extractable P (Amer et al., 1955), resin extractable P and iron oxide strip extractable P
(Menon et al., 1989).
CHAPTER 1
19
Table 1.4. Methods for the determination of soil inorganic phosphorus.
b) Sequential fractionations.
Other common approach is the use of fractionation procedures that rely on extract
sequentially P from soil with selective solvents that isolate P pools of different solubility
(Pierzynski et al., 2005). Phosphorus fractionation methods are widely used to investigate
the transformations of applied P fertilizers and understand how P fractions could be
related to soil available P. In most studies, the P fractions are grouped into available, non-
available and recalcitrant pools. The first major fractionation method was developed by
Procedure Form of P Extractants
Soil P tests
Bray P1
Labile P
0.03 M NH4F + 0.025 M HCl
Bray P2 0.03 M NH4F + 0.1 M HCl
Double lactate 0.02 M Ca lactate + 0.02 M HCl
Mehlich-3
0.015 M NH4F + 0.2 M CH3COOH
+ 0.25 M NH4NO3 + 0.013 M
HNO3 + 0.001 M EDTA
Olsen 0.5 M NaHCO3 at pH 8.5
Water
extractable P Soluble and readily
desorbable P
Deionized water
CaCl2 0.01 M CaCl2
Resin P
Bioavailable P
Chloride saturated resin + water
Iron oxide strip
Filter paper strip coated with Fe
oxide + 0.01 M CaCl2
Sequential
fractionations
Chang and
Jackson
Al-P 0.5 M NH4F
Fe-P 0.1 M NaOH
Ca-P 0.5 M H2SO4
Reductant-soluble P Citrate-bicarbonate-dithionite
Residual P Digestion with concentrated
H2SO4 and H2O2
Hedley
Loosely bound P Anion exchange resin in
bicarbonate form
Readily labile
Al-P+ Fe-P 0.5 M NaHCO3
Al-P+ Fe-P 0.1 M NaOH
Occluded P Sonication + 0.1 M NaOH
Ca-P 1 M HCl
Residual P Digestion with concentrated
H2SO4 and H2O2
Kuo
(calcareous
soils)
Soluble and loosely
bound P + Al-P 0.1 M NaOH + 1 M NaCl
Reductant soluble P Citrate-bicarbonate-dithionite
Ca-P 0.5 M HCl
Introduction
20
Chang and Jackson (1957) focused on inorganic P compounds in calcareous soils and
sediments. A significant modification of the method allowed the simultaneous
determination of both inorganic and organic P fractions from a single soil sample (Hedley
et al., 1982). The fractionation procedure most widely used is based on the modifications
summarized by Kuo (1996).
1.5.2.2 Total and organic P.
The determination of total P in soil needs the solubilization of P through the decomposition
of mineral and organic containing materials in the soil. Currently the most common
approaches involve either alkali or acid oxidation or digestion of the soil samples:
H2SO4/H2O2/HF digestion method (Bowman, 1988); NaOBr/NaOH (Dick and Tabatabai,
1977) and H2SO4/H2O2/Li2SO4 (Parkinson and Allen, 1975).
Total organic P is estimated indirectly by ignition or extraction. The difference of P
extracted by an acid solvent between ignited and unignited samples is attributed to
organic P (Saunders and Williams, 1955). The extraction of organic P with Na2EDTA and
NaOH (Bowman and Moir, 1993) has been found to be highly efficient to recover most
organic P (Turner et al., 2003a; Doolette et al., 2011) as NaOH solubilizes organic P
associated with organic matter while the EDTA chelates metal cations to increase the
efficiency of the organic matter extraction. Organic P is calculated as the difference of
P in the extracts before and after oxidation of organic matter by digestion.
Both ignition and extraction methods are prone to errors that can over or
underestimate the soil organic P concentration. The solubility of P may change with
ignition whereas extraction procedures might cause hydrolysis or incomplete extraction
of organic P compounds (Condron et al., 2005). For this reasons, by the moment there is
no method to determine exactly organic P and these measurements should be
considered as estimates. There are three general approached to characterize organic
P: sequential fractionation, spectroscopy analysis and enzymatic hydrolysis.
a) Sequential fractionation.
Although these protocols have been initially developed to determine inorganic P, also
organic P is dissolved in some fractions (Bowman and Cole, 1978). Inorganic P is
determined in each fraction before and after digestion to calculate organic P. However,
this methods may be unreliable for the determination of organic P because the hydrolysis
of organic P and the readsorption and/or precipitation of inorganic P compounds may
overestimate organic P. Additionally, the extractants used are not specific for any
particular group of organic P compound (Magid et al., 1996) and the conventional
classifications of labile P can only approximate the availability to plant and microbes
CHAPTER 1
21
(Tiessen and Moir, 1993) so the information obtained by sequential fractionations should
be interpreted with caution.
Table 1.5. Methods for the determination of soil organic phosphorus.
b) Nuclear magnetic resonance spectroscopy.
Due to its magnetic properties, 31P is easily detected by nuclear magnetic resonance
spectroscopy (31P-NMR), making it a good tool for to characterize of organic P (Condron
et al., 1997). This technique may be used on solid samples (solid-state spectroscopy) or
extracted samples (solution spectroscopy). For solution 31P-NMR spectroscopy, soil is
usually extracted with 0.25 M NaOH plus 0.05 M EDTA, concentrated by lyophilization and
subsequently reconstituted and analyzed after pH adjustment to 12 (Cade-Menun and
Preston, 1996; Crouse et al., 2000; Cade-Menun, 2005). The assignment of peaks in NMR
spectra is based on literature reports and on the position of known organic P compounds
added during extraction or analysis (Cade-Menun and Preston, 1996). An advantage of
31P-NMR spectroscopy is the simultaneous observation of all organic species in a single
Procedure Form of P Extractants
Ignition Total organic P
NaOH-EDTA Extraction
Sequential
fractionations
Chang
and
Jackson
Labile P 1 M NH4Cl
Al-bound P 0.5 M NH4F
Fe-bound P 0.1 M NaOH
Ca-bound P 0.5 M H2SO4
Reductand-soluble P Citrate-bicarbonate-dithionite
Refractory P 0.1 M NaOH
Hedley
Labile P (Pi) 0.5 M HCl, resin strip
Labile P 0.5 M NaHCO3
Microbial P Fumigation, 0.5 M NaHCO3
Fe-bound P 0.1 M NaOH
Ca-bound P 0.1 M HCl
Recalcitrant P Conc. HCl
Residual P Digestion, conc. H2SO4, H2O2
31P-NMR
Phosphonates
NaOH-EDTA
Orthophosphate monoesters
Orthophosphate diesters
Pyrophosphate
Polyphosphates
Enzyme additions
Ortophosphate monoesters
NaOH-EDTA Ortophosphate diesters
Inositol hexakisphosphate
Introduction
22
extract (Doolette and Smernik, 2011). However, overlapping peaks can hinder the
quantification, especially in the monoester region, being not possible to distinguish
between single species. Another inconvenient of this technique is the high cost of the
equipment for analysis.
c) Enzyme additions.
The conversion of organic P forms to orthophosphate is mediated by hydrolytic enzymes
(Quiquampoix and Mousain, 2005) which are released actively by microorganisms and
plant roots and subsequent stabilized in soil coloids or associated to active cells
(Nannippieri et al., 2011). In soil, the availability of organic P for plants was proposed to
depend more on the availability of a hydrolyzable substrate rather than on enzymatic
activity (Tarafdar and Claassen, 1988). Thus, the enzymatic hydrolysis of soil organic P can
be used as an analytical tool (Bünemann, 2008). If an excess of hydrolytic enzymes is
added to a soil sample, the enzyme labile substrate is hydrolyzed releasing
orthophosphate that can be measured colorimetrically (Figure 1.10). Knowing the
specificity of each enzyme preparation allows to know which specific form of organic P
has been hydrolyzed. Additions of alkaline phosphatase, nuclease, phosphodiesterase,
phytase and phospholipase are frequently used in soils (Bünemann, 2008).
Figure 1.10. Principle of the enzyme additions assays.
The addition of enzymes (alone or in combination) to soil extracts to identify
different classes of organic P has been proposed by Turner et al. (2002b). They classified
CHAPTER 1
23
water extractable P into hydrolyzable orthophosphate monoesters, orthophosphate
diesters and inositol hexakisphosphate (Ins6P) by adding alkaline phosphatase, a
combination of alkaline phosphatase and phosphodiesterase, and phytase respectively.
One advantage of this technique is that it could be performed with relatively little lab
equipment and costs compared with 31P-NMR spectroscopy and the P classes obtained
are comparable using both methods (He et al., 2007; Johnson and Hill, 2010).
1.6. ORGANIC WASTES AS SOURCE OF NUTRIENTS IN AGRICULTURE.
Organic wastes have been traditionally applied to agricultural soils in order to replenish
soil organic matter and to improve soil structure and fertility (Giusquiani et al., 1995;
Crecchio et al., 2004). Moreover, the addition of organic wastes to soil has gained
importance in the last decades as a consequence of the worldwide environmental
concern about several aspects as the need of recycling the increasing amounts of
wastes generated, the scarcity of nutrient sources (e.g. phosphorus) or the contribution
of agriculture to the atmospheric concentration of greenhouse gases (restoring soil
organic matter as sink of CO2).
The incorporation of organic wastes can improve soil conditions increasing N and
P availability. A significant proportion of total N applied to soil with organic wastes is
organically bound but this organic N pool mineralizes slowly into ammonium and
subsequently into nitrate in the soil but only a part becomes available for plant
production in the year of application. Nitrogen mineralization occurs simultaneously with
N immobilization depending primarily of the C:N ratio of the material applied (Wild, 1988;
Tisdale et al., 1999). N immobilization has been reported to be especially large with the
decomposition of residues with high C (Cliff et al., 2002; Moritsuka et al., 2004). The
availability of N in soils during the short and long term is strongly linked to the soil organic
matter and the magnitude of the active pool and the more passive pool of soil organic
matter (Amlinger et al., 2003).
Total concentrations of P in organic wastes vary widely according to the origin
and the treatment process employed being inorganic P usually the main form (García-
Albacete et al., 2012). The availability of P in soil after incorporation of organic materials
requires an understanding not only of the forms and amounts of P in these materials but
also of other elements which can modify P behaviour in soil. In addition to the available
P that contain these wastes, organic matter in residues contains organic P which is
converted into orthophosphate through mineralization. In addition, the humic
substances and organic acids released during the partial decomposition of organic
Introduction
24
wastes can be adsorbed into soil surfaces, decreasing the potential P adsorption by
blocking sites for the formation of complexes with Al, Fe and Ca (Iyamuremye et al.,
1996a; Haynes and Mokolobate, 2001; Akhtar et al., 2002; Mkhabela and Warman, 2005).
The organic matter incorporated into soil with the organic wastes has an influence on
the composition and enhancement of the microbial biomass producing changes in the
enzymatic activity (García-Gil et al., 2000; Marinari et al., 2000). The addition of organic
wastes also may produce changes in soil pH induced by decomposition of organic acids
or associated with N mineralization (Mokolobate and Haynes, 2002; Hedley and
McLaughlin, 2005).
Table 1.6. Kjeldahl N, total P concentration and N:P ratio in a wide range of
organic wastes with different origin and treatment
(modified from García-Albacete et al., 2012).
On the other hand, in order to establish the correct application dose it is important
to consider the environmental risks associated to the use of organic wastes in agriculture.
The application of organic wastes high in NH3 (e.g. slurry) or readily decomposable
organic N to the soil surface can result in losses by NH3 volatilization, particularly if the soil
or the organic material have an alkaline nature. Amending soils with organic wastes
generally increases the potential of denitrification because they are a source of
available C and they increase soil moisture holding capacity, increasing N2O and NOx
emissions. As with the use of mineral fertilizers, the application of organic wastes with a
large potential of N mineralization can led to nitrate losses especially in humid or irrigated
areas. The Nitrates Directive established an application limit of 170 kg N ha-1 year-1 for
animal manure but the use of other organic wastes in agriculture is not clearly regulated
Waste Description of the original material Kjeldahl
N
Total
P N:P
(mg kg-1) (mg kg-1)
Compost Municipal solid waste + vegetal waste 19500 6448 3.02
Compost Municipal solid waste + vegetal waste 13700 5889 2.33
Digestate Municipal solid waste + vegetal waste 19500 3637 5.36
Digestate Municipal solid waste 34700 8535 4.07
Compost Biosolid + vegetal waste 26900 14164 1.90
Compost Biosolid + vegetal waste 14059 5640 2.49
Compost Biosolid + pig slurry + dairy manure 16400 20794 0.79
Compost Dairy manure + vegetal waste 26100 19104 1.37
Compost Exhausted grape marc 21400 3793 5.64
Compost Pig slurry 14100 53407 0.26
Compost Champ 23900 11350 2.11
Slurry Pig slurry 32300 12074 2.68
Heat drying sludge Sewage sludge 42000 31247 1.34
Dewatered sludge Water treatment sludge 35000 4722 7.41
CHAPTER 1
25
by this Directive. The Codes of Good Agricultural Practice encourage the calculation of
N balances to establish N application doses which requires to know how much N is
mineralized in soil during the crop season.
A significant proportion of P contamination in water pathways come from the use
of organic wastes employed as soil amendments (Sharpley et al., 1994; Haygarth and
Jarvis, 1997; Sims et al., 2000). This is partly due to the dosage of these wastes, which is
usually make as function of its N content. However, depending on the N:P ratio of the
organic material applied (Table 1.6), this criteria results in some cases in an over-
application of P and consequently in an accumulation of this element in soil being
susceptible of transport through water pathways (Mozaffari and Sims, 1994; Sims et al.,
1998; Sharpley et al., 2007). In Europe there are emerging P legislations at country or
region level which establish maximum P application standards by using different systems:
maximum application limits (e.g. northern Belgium, Estonia), P balance systems (e.g.
Germany) or soil P tests (e.g. Northern Ireland).
In conclusion, understanding the fate of N and P in organic waste amended soils
is needed to manage agricultural soils in order to minimize potential environmental
impacts. A nutrient management plan should ensure the efficient use of organic wastes
so that limit nitrate leaching, prevent accumulation of phosphorus in soil and reduce the
risk of nitrous oxide losses to the atmosphere. The dosage of these materials should be
given by the specific conditions of each agricultural system as the soil type,
management and its associated potential risks.
CHAPTER 2:
OBJECTIVES
CHAPTER 2
29
The foregoing introduction has been made clear the importance of managing nitrogen
and phosphorus in agricultural systems, from an agronomic and environmental
viewpoint. Organic wastes are employed in agriculture to improve soil conditions and
plant growth, as source of organic matter and nutrients. The processes that occur in the
soil after the application of these materials are of great importance since they control
the availability of nutrients in soil for plant uptake (opportunities) as well as the nutrient
losses from the agro-systems which can cause contamination problems (risks).
The main aim of this Thesis is to extend the understanding of the processes that
could influence N and P dynamics in the soil-plant system after the agricultural use of
different organic wastes.
2.1. NITROGEN.
Although N processes in soil have been widely studied in the past, the challenge at this
point is to evaluate the result of different agricultural practices to achieve both
agronomic and environmental goals, especially in zones vulnerable to nitrate pollution
with a higher risk of N contamination. There are many organic fertilizers, such as the
wastes derived from the winery and distillery industry, whose behaviour in soil is still poorly
understood. These wastes are generated in Spain mainly in vulnerable zones, so before
its use in agriculture it is necessary to evaluate N dynamics and the agronomic and
environmental response. Hence arises one of the specific objectives of this Thesis:
1. To evaluate the effect of the application of wastes from the winery and distillery industry
on N dynamics from an agronomic and environmental viewpoint. This objective aims to
answer the following questions:
- What is the soil N mineralization rate after the application of these wastes?
- What is the effect of the application dose?
- How can other agricultural practices (irrigation) affect N dynamics?
- What is the crop response associated with this element?
- What is the risk associated?
Objectives
30
2.2. PHOSPHORUS.
This element has been less studied than nitrogen. However, due to the perspectives of
global phosphorus scarcity, there is an increase interest to find new P sources and to find
more efficient agricultural practices in order to ensure food security. Due to the
complexity of the soil P cycle, there are many factors that can influence P dynamics in
the soil-plant system after the incorporation of organic materials and that we still should
understand. The other specific objective of this Thesis is:
2. To determine which factors influence mostly on the availability of phosphorus in soils
after the application of organic wastes. This objective aims to answer the following
questions:
- What are the P forms in soil after the application of organic wastes?
- What is the effect of the soil type?
- What is the effect of the waste type?
- What is the crop response associated with this element?
- What is the risk associated?
2.3. METHODOLOGY.
The methodology used to achieve the objectives presented based on the bibliography
reviewed is shown in Figure 2.1.
Two type of experiments were carried out to evaluate the effect of winery and
distillery wastes addition on N dynamics: soil incubations and field experiments. An
incubation trial was included to determine the nitrogen potential mineralization potential
of compost made with these materials under defined conditions of temperature and
moisture. The field experiment was necessary to study the effect of organic wastes
addition in a real agricultural system typical of the area where these wastes are
generated in Spain: an irrigated horticultural crop. The agronomic benefits of compost
addition together with the nitrate leaching risk were evaluated under field conditions.
CHAPTER 2
31
Figure 2.1. Experiments, variables studied and factors considered to achieve the aims
proposed in this Thesis.
The availability of P in agricultural soils with the application of different organic
wastes was studied under different scenarios and using both field experiments and
laboratory incubations. A soil incubation experiment was carried out to assess more
accurately soil P availability along time and the effect of the application of different
wastes derived from the winery and distillery industry on inorganic P fractions in soil. The
results obtained were validated under real field conditions. The study of organic P forms
was carried out in a long term field experiment where different organic fertilizers were
evaluated (cattle manure and biowaste compost).
CHAPTER 3:
NITROGEN MINERALIZATION IN SOIL AFTER ADDITION
OF WINE-DISTILLERY WASTE COMPOST: LABORATORY
AND FIELD EVALUATION
Requejo, M.I., Cartagena, M.C., Villena, R., Giraldo, L., Arce, A., Ribas, F., Cabello, M.J.,
Castellanos M.T. (2015). Soil Research. Accepted.
CHAPTER 3
35
3.1. ABSTRACT.
The application of wastes from the wine-distillery industry as source of organic matter and
nutrients could be a good option of agricultural management. This study is focused on
soil nitrogen (N) mineralization after addition of compost derived from this industry at
different doses (7, 13 and 20 t ha-1). An aerobic soil incubation in controlled conditions
was carried out to study N mineralization from the soil-compost mixture as well as isolating
the compost from the soil. The data were fitted to a non-linear regression obtaining low
values of potentially mineralizable N (N0) and constants of mineralization (k) (from 81 to
104 mg kg-1 and from 0.008 to 0.013 days-1 for the soil-compost mixtures, and from 42 to
71 mg kg-1 and from 0.009 to 0.015 1 days-1 for the increasing doses of compost) which
indicates that it is a mature compost very resistant to mineralization. Nitrogen mineralized
(NM) in the field during two growing seasons (2011 and 2012) of a melon crop was
calculated through a N balance, taking into account N imputs and outputs in the soil-
plant system. NM in the unamended plots accounted to 31 kg ha-1 and 24 kg ha-1 in 2011
and 2012, respectively, and increased proportionally to the dose of compost applied
until 113 kg ha-1 and 98 kg ha-1 in the consecutive years. The constants of mineralization
obtained in the laboratory were adjusted by field temperatures to predict NM in the field
and a general overestimation was observed. The best estimates were obtained when
considering the mixture of soil and compost, which reflects the important role of the soil
to evaluate N mineralization caused by the addition of organic wastes.
N mineralization from wine-distillery waste compost
36
3.2. INTRODUCTION.
Spain produces more than 3 million tonnes of wine each year being the third producer
in the world (FAO, 2013). During the winemaking process, solid residues like grape stalks
are generated, as well as grape marc (composed of the skin, pulp and seeds of the
grape) and wine lees as by-products. A large proportion of the grape marc and wine
lees produced are sent to the distilleries obtaining exhausted grape marc and lees cake.
These residues are generated within a short period of the year (from August to October)
and have polluting characteristics such as low pH and phytotoxic and antibacterial
phenolic substances (Bustamante et al., 2008a). For this reason, in the recent years there
is an effort to develop potential uses of these wastes to produce different substances
with a market value (Arvanitoyannis et al., 2006; Bustamante et al., 2007). The composting
of these wastes has been shown to be a feasible option for their use in agriculture, as
source of nutrients and organic matter (Lasaridi et al., 2000; Ferrer et al., 2001; Brinton and
York, 2003; Requejo et al., 2014). Several authors have reported positive effects after the
application to soil of winery and distillery waste composts. These include the
enhancement of carbon stocks and soil microbial activity (Bustamante et al., 2010;
Mosetti et al., 2010; Paradelo et al., 2011), the availability of macronutrients (Bustamante
et al. 2011) and the suppression of plant pathogens (Borrero et al., 2004; Diánez et al.,
2007; Segarra et al., 2007).
The winery and distillery wastes contain important amounts of nitrogen (N),
sometimes higher than other organic wastes used in agriculture, as sludge, municipal
solid waste or manure (Bustamante et al., 2008a). As this element is often a limiting factor
for crop growth, the organic wastes could be a source of available N, reducing the need
for mineral fertilizers (Cordovil et al., 2005; Odlare et al., 2011). However, this N is mostly in
organic form. Some of the organic N forms in organic wastes are readily mineralized to
inorganic N, while others are chemically or structurally protected being recalcitrant
(Chadwick et al., 2000), so to make a good diagnosis of N fertilization and to establish
the correct doses of application, it is necessary to know how much N is mineralized during
the crop growing season. Methods involving soil incubations to determinate N
mineralized have been generally accepted to estimate N availability for crop growth by
mineralization of soil organic matter (Stanford et al., 1974; Gianello and Bremner, 1986;
Serna and Pomares, 1992a). Several studies have evaluated N mineralization after
addition of raw or composted wine-distillery wastes to soil using laboratory incubations
and reported a slow mineralization rate (Flavel et al., 2005; Bustamante et al., 2007;
Paradelo et al., 2011). Other studies have studied the availability of N by determining
nutrient plant uptake (Masoni et al., 2000; Montemurro et al., 2010). However, to better
CHAPTER 3
37
understand the release of N from these materials, further investigation under field
conditions would be required.
It is accepted that the incorporation of organic materials to soil can affect
mineralization of soil organic matter, what is known as ̀ priming effect´ (Bernal et al., 1998;
Pascual et al., 1998; Leifeld et al., 2002). However, in studies of mineralization of organic
wastes, mainly through soil incubations, it is generally accepted to calculate N
mineralized from the residue independently (Chae and Tabatai, 1986; Cordovil et al.,
2005; Sistani et al., 2008), without considering the soil-residue effect.
The objective of this study was to study N mineralization after the addition to soil
of different doses of compost derived from the wine-distillery industry: (i) using aerobic
soil incubations and fitting the data to a mathematical model; (ii) applying a N balance
in field conditions to know N mineralized during the growing season of a selected crop;
(iii) comparing the results obtained by the model with the values observed in the field. In
addition, in order to evaluate the different approaches to study N mineralization after
application of organic wastes to soil, N mineralized was calculated: (i) by considering N
mineralized in the soil-compost mixture; (ii) by substracting the amount of N mineralized
in the unamended soil from N mineralized in the compost-treated soil.
3.3. MATERIALS AND METHODS.
3.3.1. Laboratory incubation experiment.
A laboratory aerobic incubation experiment was carried out using the arable layer of soil
(0-30 cm) collected in `La Entresierra’ field station at Ciudad Real (central Spain; 3º 56’
W; 39º 0’ N; 640 m altitude), in the same plots used in the field experiment. Soil
characteristics are reflected in Table 3.1. The soil is a shallow sandy clay loam, classified
as Petrocalcic Palexeralfs (Soil Survey Staff, 2010). Soil samples were air-dried at room
temperature and sieved to 2 mm before the incubation.
The compost used in the incubation and in the field experiment was obtained
from a composting plant located in Socuéllamos (Ciudad Real) which collects residues
from the wineries and distilleries around this area and through an aerobic degradation
produces a commercially compost composed of grape stalks, exhausted grape marc
and lees cake. The main properties of the compost used are shown in Table 3.1. This
compost fulfilled the criteria established by the Spanish legislation concerning the use of
organic materials in agriculture.
N mineralization from wine-distillery waste compost
38
Figure 3.1. Compost pile (Agricompost S.L.) and compost aspect during the application.
Table 3.1. Characterization of the soil and compost used in the experiments.
Three kilograms of soil were mixed with increasing rates of compost (from the
batch applied in the field in 2012) before fill in plastic pots. Each amount of compost
applied correspond to the different doses applied in the field experiment corresponding
to 230 (S+D1), 460 (S+D2) and 690 (S+D3) kg N ha-1. A control pot without compost
addition was also established (S). Treatments were replicated three times. The pots were
maintained at 70% of the water holding capacity by adding periodically distilled water
and the incubation was performed at 25 °C. A plastic covered the pots during the whole
experiment to avoid volatilization losses. A soil sample was taken from each pot at 1, 7,
14, 21, 35, 49, 63, 77, 119 and 169 days after the beginning of the incubation. In each
sampling date, the soil was extracted with deionized water to determine nitrates (NO3)
with an ion-selective electrode and with 1 M KCl to measure ammonium (NH4) using
spectrophotometric techniques. Inorganic N was calculated as the sum of NO3-N and
NH4-N at every sampling date.
Soil Compost 2011 Compost 2012
Sand (%) 70.4 − −
Silt (%) 8.0 − −
Clay (%) 21.6 − −
pH 8.4 9.2 9.8
EC (mS cm-1) 0.2 1.2 1.0
Organic carbon (g kg-1) 14.0 316.9 348.8
N (Kjeldahl) (g kg-1) 1.0 32.9 33.0
NO3--N (mg kg-1) 15.3 0.2 0.3
NH4+-N (mg kg-1) 1.7 1.5 1.8
C:N 14.0 10.1 11.2
CHAPTER 3
39
Figure 3.2. Incubation pots and soil samplings.
To fit the results, a non-linear regression model proposed by Stanford and Smith
(1972) was used:
Nm= N0(1- e -kt) (1)
where Nm represents accumulated mineralized N at time t; N0 is the potentially
mineralizable organic N; k is the first-order rate constant of mineralization; and t is the time
of incubation.
The fraction of N mineralized from the total N applied with each dose of compost
was calculated as:
% Mineralized N= 100 x [(NO3-N(t) - NO3-N(0)) - (NO3-N(t)c - NO3-N(0)c)]/N(0) (2)
where NO3-N(t) is the concentration of nitrate-N present at each time in the compost-
amended soil, NO3-N(0) is the concentration of nitrate-N present initially in the amended
soil, NO3-N(t)c is the concentration of nitrate-N present at each time in the unamended
soil, NO3-N(0)c is the concentration of nitrate-N present initially in the unamended soil, and
N(0) is the total N added with each dose of compost.
3.3.2. Field experiment.
The field experiment was carried out in La Mancha, a region in central Spain which holds
the highest production of wine in the country (MAGRAMA, 2012a). As the ideal scenario
is the application of these materials close to the area where are generated, the N
balance was carried out during two growing seasons of a melon crop (Cumumis melo
L.), traditionally cultivated in this region. The trial was performed during the April to August
seasons of 2011 and 2012. The area is characterized by a Mediterranean climate, with
strong continental character and widely fluctuating daily temperatures. The rainfall
season occurs outside the melon growing period. During the three years previous to the
N mineralization from wine-distillery waste compost
40
experiment, the plots did not receive nor fertilizer nor organic amendments and were
used to grow non-irrigated winter wheat (Triticum aestivum L.).
Due to low rainfall during the melon crop season and high evapotranspiration
rates, irrigation is necessary. Drip-irrigation strategies in combination with plastic-much
have been adopted in this area for horticultural production. All years, the water used was
groundwater from a well near the experimental plots.
Four treatments were set up both years: a control with no compost addition (S),
and three doses of wine-distillery waste compost application corresponding to 7 (S+D1),
13 (S+D2) and 20 (S+D3) t ha-1. A randomized complete-block design was used, with the
dose of compost as factor of variation. Each treatment included four replications plots
(12x15 m) formed by ten rows with eight plants each. The compost was applied on 20
April in 2011 and 19 April in 2012 localized in the crop row and incorporated into the soil
at about 10-20 cm depth.
Figure 3.3. Details of compost application in the field.
A `Piel de sapo´ type melon (Cucumis melo L. cv. Trujillo) was used as crop. Each
year of experimentation, melon seeds were germinated from late March under
greenhouse conditions, until they had sprouted two or three real leaves. Subsequently,
the seedlings were transplanted in the field plots approximately 20 days after compost
incorporation (11 May 2011 and 9 May 2012) onto plastic mulch at a density of 4444
plants ha-1 (1.5 by 1.5 m). After transplanting, all plots received initially 28 mm of water in
order to facilitate crop establishment. The irrigation schedule was calculated from 27 to
90 days after transplanting in 2011, and from 20 to 97 in 2012, as a single and daily
irrigation in order to replenish the ETc calculated daily using FAO method (ETc = Kc x ET0,
Doorenbos and Pruitt, 1977). ET0 is the reference evapotranspiration and was estimated
by the FAO Penman-Monteith method (Allen et al., 2002), using daily data from a
meteorological station sited near the experimental field. Kc is the crop coefficient
CHAPTER 3
41
obtained in previous years for melon crop in the same conditions (Ribas et al., 1995).
Rainfall was negligible, so the water to be applied in each treatment every week was
calculated as the ETc of the previous week divided by the efficiency of the system, which
considers the salt tolerance of the crop, the quality of irrigation water, the soil texture and
the homogeneity of the irrigation system (Rincón and Giménez, 1989), estimated at 0.81
in our conditions. This result was then divided by the number of days to obtain the daily
irrigation needs.
Figure 3.4. Experimental plots at plantation and development of the melon crop.
3.3.3. Nitrogen balance.
In the field experiment, the N mineralized (NM) over the growing season was determined
using a N balance:
NM = Plant N uptake + N losses + N soil final – N applied – N soil initial (3)
Soil samples from the arable layer (0-30 cm) were collected before compost
addition and at the end of the growing season. The samples were extracted with
deionized water or 1 M KCl for the determination of NO3 and NH4 respectively. The NO3
concentration in the extracts was determined with an ion-selective electrode and the
NH4 concentration using spectrophotometric techniques.
Four melon plants per treatment were collected at the end of the growing season
and separate intro leaves, stems and fruits. The plant samples were dried at 80 °C,
weighted and ground to a fine powder. Nitrogen content was determined using the
Kjeldahl method (Association of Official Analytical Chemists, 1990). The N accumulated
by each organ was obtained as the product of N concentration and dry weight. The
plant N uptake was determined as the sum of the N accumulations of each above-
ground organ of the plant.
N mineralization from wine-distillery waste compost
42
Nitrogen losses may occur via denitrification, volatilization and/or leaching. In our
experimental conditions, N losses by denitrification were considered negligible (Sánchez-
Martín et al., 2008). Due to the clay proportion of this soil, the ammonium is fixed by the
soil avoiding volatilization losses (Bremner, 1965). However, during the irrigation period,
the use of higher volumes of water to flush out the salts deposited in the soil, primarily by
drip systems, probably results in N leaching losses (Doorenbos and Pruitt, 1977; Ayers and
Westcott, 1987) due to the presence of a discontinuous and fragmented petrocalcic
layer below the first 0.60 m of soil with cracks that provide vertical permeability locally,
being this area declared vulnerable to nitrate pollution (Directive 91/676/EEC). The N
leaching to below 0.60 m was calculated as the product of NO3-N concentration and
drainage. To estimate drainage, the following water balance was used (Doorenbos and
Pruitt, 1977):
D=Irr + Prec - ETc- Rf ± ∆Ɵv (4)
where Irr is the irrigation, Prec is the precipitation, Rf is the runoff that was assumed to be
negligible and Ѳν is the volumetric soil water content. A tube was installed in the center
of each plot and between two consecutives plants to measure Ѳν on a weekly basis, in
a straight line at 0.375 m from the drip line, with a probe (Diviner 2000) based on FDR
(Frequency Domain Resonance). A porous ceramic cup was used to extract samples of
the soil solution every week simultaneously with the FDR measurements. The cups were
installed vertically in each plot between two consecutives plants to a depth of 0.60 m
and 0.375 m from the irrigation line. A suction of −0.7 bar was applied to the cups with a
pressure vacuum pump to extract the soil water. Water samples were stored in a freezer
until analysis. The nitrate concentration in the soil water solution was measured with an
ion-selective electrode.
Nitrogen received in the irrigation water as well as mineral N added with the
compost were considered as N applied. Nitrogen inputs from irrigation water were
calculated from the amount of irrigation water used every week and the N
concentrations in water samples (measured weekly) summed over the irrigation
schedule.
3.3.4. Prediction of N mineralized in the field.
The exponential model was used to predict N mineralized in the field during the melon
crop season. However, the rate constant of mineralization (k) needed to be adjusted to
field air temperature using the correction proposed by Standford and Smith (1972):
k1=k . 2(
T daily-T incubation
10) (5)
CHAPTER 3
43
where k1 is the adjusted rate of mineralization to air temperature, k is the constant
obtained in the model, Tdaily is the average daily air temperature and Tincubation is the
temperature used for the soil incubation. The prediction of N mineralized is usually
corrected by soil water content (Cabrera and Kissel, 1988; Zinati et al., 2007) using the
next factor:
W = soil water content in the field/optimum soil water content (6)
The optimum soil water content was considered the water content at which the
soil was incubated (70% of water holding capacity). In our experimental conditions,
under drip irrigation, the moisture in the soil profile was never below the optimum
considered, so this correction was not required.
The day following the initial soil sampling in the field was considered the first day
to carry out the prediction of NM, and the day of the final soil sampling the end.
3.3.5. Statistical analysis.
Non-linear regression by the Levenberg-Marquardt method was used to calculate N0 and
k using the SPSS statistical software. A confidence interval of 95% was selected in order
to assess the statistical significance of the curve-fitting.
Analysis of variance (ANOVA) was carried out separately for each year of the
field experiment to determine the effects of the treatments on the measured parameters.
Significant statistical differences of variables between the different treatments were
established by the Tukey test (p≤0.05).
Linear regressions were calculated between estimated and observed values of
mineralized N.
3.4. RESULTS AND DISCUSSION.
3.4.1. Nitrogen mineralization observed in laboratory conditions.
The evolution of inorganic N (NO3-N, NH4-N and their sum) in the unamended (S) and
amended soil (S+D1, S+D2, S+D3) during the incubation period is shown in Figure 3.5.
Nitrate dominates over ammonium during all the time of incubation, being the
concentrations of ammonium almost zero in the last weeks of incubation. One week after
compost incorporation, the concentration of inorganic N started to increase for all the
doses. Bustamante et al. (2007), applying wine and distillery wastes (including grape
N mineralization from wine-distillery waste compost
44
marc, stalks, exhausted grape marc and wine lees) to different type of soils, reported an
initial N immobilization in the first 20 days of incubation in all cases. They explained this
fact by the high C:N of the residues, the polyphenol content and low concentration of
soluble compounds. This N immobilization was also observed by Flavel et al. (2005) in a
soil amended with composted grape marc, but they attributed this fact to an incomplete
decomposition of the material. Our results showed that with a correct process of
composting, the stabilized material added to soil should no cause immobilization of N
(Table 3.2).
Figure 3.5. Evolution of inorganic N (a) nitrate-N (b) and ammonium-N (c) during 24
weeks of incubation. S: unamended soil; S+D: soil-compost mixture. The bars indicate
standard deviation at n=3.
CHAPTER 3
45
Inorganic N increased in the control and amended soils during all the incubation
period. In all the sampling dates, the content of inorganic N in the amended soils
remained higher than the control, increasing with respect to the dose of compost and
tended to increase until 119 days of incubation. Since this moment, the concentration of
inorganic N in the soil amended with the different doses of compost remained constant
or even a slight decrease was observed. The same trend was reported by other authors
applying different organic wastes to calcareous soils (Hernández et al., 2002; Zarabi and
Jalali, 2013).
Table 3.2. Net N mineralization (%) with the different doses of compost during the
incubation.
After 24 weeks of incubation, the amended soils increase the inorganic N content
by 31 (S+D1), 45 (S+D2) and 66 mg kg-1 (S+D3) with respect to the control (Figure 3.5).
Although inorganic N increased with increasing compost application rates, our study
showed that doubling the rate did not double the inorganic N content in soil. This is in
agreement with the findings of Lindeman and Cárdenas (1984) and Hernández et al.
(2002) in soils amended with sewage sludge, and with the results of Mantovani et al.
(2006) for urban waste compost. This fact suggests that mineralization of N does not only
rely on the presence of a mineralizable pool of N, but also in the activity of soil microbial
community, which could be diminished by the application of large amounts of compost
(Xue and Huang, 2013).
As seen in Table 3.2, net N mineralization from the compost after 6 months of
incubation accounted for 1.61, 1.33 and 1.21% of the total N added with the different
rates of compost (D1, D2 and D3 respectively). These values were lower than those
reported by Cabrera et al. (2005) for composted olive mill sludge with a longer period of
Net N mineralization
(%)
Days D1 D2 D3
7 0.40 0.67 0.58
14 0.32 0.32 0.23
21 0.22 0.50 0.51
35 0.61 0.30 0.39
49 0.77 0.60 0.71
63 0.80 1.03 1.05
77 1.15 1.14 0.86
119 1.83 1.69 1.69
169 1.61 1.33 1.21
N mineralization from wine-distillery waste compost
46
incubation, but lower than those reported by Mantovani et al. (2006) for urban waste
compost during 24 weeks of incubation.
The estimates of potentially mineralizable N (N0) and the constant of
mineralization rate (k) obtained by using the model of Standford and Smith (1972) are
presented in Table 3.3. In all cases, the model showed a good fit to the amounts of
mineralized N obtained over the soil incubation. The adjustment of the model is better
when considering the soil-compost mixture, possibly because this model was originally
developed for soils and not for substrates.
Table 3.3. N mineralization constants.
Values between parentheses represent the standard error. S: unamended soil; S+D: soil-compost
mixture; D: compost. Nm: N mineralized during the whole incubation, N0: Potentially mineralizable
N, k: First-order rate constant of mineralization, r: correlation coefficients obtained as a result of the
non-linear least-squares equation.
The N0 and k values obtained for the unamended soil were according to the
range reported by other authors for wide different soils. Stanford and Smith (1972)
reported values of N0 between 20 to 300 mg kg-1 and k between 0.005 and 0.014 days-1.
Cabrera and Kissel (1988) obtained values of N0 and k between 9 to 280 mg kg-1 and
0.002 to 0.006 days-1 respectively. The N0 value of 41.0 mg kg-1 corresponds to 4.1% of the
soil Kjeldahl N (1.0 g kg-1) (Table 3.1). The values of N0 in D1 (42 mg kg-1), D2 (53 mg kg-1)
and D3 (71 mg kg-1) accounted for 2.7, 1.7 and 1.5% of the organic N added with the
compost which indicates that the major part of organic N in compost derived from winery
and distillery wastes is resistant to mineralization. The k constants increased with respect
to the dose of compost applied ranging from 0.009 to 0.015 days-1. These values were low
to those reported for sewage sludge by Serna and Pomares (1992b) (0.013-0.126 days-1)
and by Hernández et al. (2002) (0.031-0.163 days-1). However, the mineralization
constants obtained in this study are comparable with those obtained by Cabrera et al.
(2005) (0.006 days-1) applying composted olive mill sludge and by Gil et al. (2011) (0.002-
0.009 days-1) for composted bovine manure and sewage sludge. These values of k may
indicate the stability of the compost obtained from the wine-distillery wastes so that a
slow mineralization of organic N after the addition of this compost to soil is presented.
Treatment Nm
(mg kg-1)
N0
(mg kg-1)
k
(days-1)
r
S 26 41 ( ± 10) 0.006 ( ± 0.003) 0.92
S+D1 56 81 ( ± 16) 0.008 ( ± 0.002) 0.96
S+D2 67 87 ( ± 7) 0.011 ( ± 0.002) 0.90
S+D3 82 104 ( ± 27) 0.013 ( ± 0.004) 0.91
D1 30 42 ( ± 17) 0.009 ( ± 0.003) 0.96
D2 41 53 ( ± 6) 0.013 ( ± 0.002) 0.87
D3 56 71 ( ± 23) 0.015 ( ± 0.005) 0.88
CHAPTER 3
47
3.4.2. Nitrogen mineralization observed in the field: N balance.
The components used to carry out the N balance (Eq. (2)) for the melon crop under field
conditions are presented in Table 3.4.
Table 3.4. Parameters for the N balance during the field experiment in 2012 and 2012.
N uptake (Nup), N leaching (Nl), N applied with irrigation water (Nw), N applied with compost (Nc),
N applied (Nap=Nw+Nc), soil mineral N at the beginning (Ns initial) and at the end (Ns final) of the
crop cycle and N mineralized (NM).S: unamended plots, S+D: compost-amended plots. Within
each column and year, means followed by the same letter are not significantly different at p≤0.05.
The main input of N for the melon crop was the N applied with the irrigation water.
The N applied (Nap) was higher in 2012 due to a higher concentration of nitrates in the
irrigation water and a longer irrigation period. The total amount of N applied with
irrigation during the total schedule period was 89 kg ha-1 in 2011 and 132 kg ha-1 in 2012.
These important inputs of N from irrigation water come from the high concentration of
nitrates in the groundwater, the source of water for the irrigation systems in this area
(Castellanos et al. 2013; Requejo et al. 2014). For this reason, the inclusion of this input of
N in the balance in these conditions is necessary for fertilizer recommendations
(Castellanos et al., 2013). On the other hand, the contribution of inorganic N applied with
the compost to the total Nap was insignificant (from 2 to 6 kg N ha-1), so the differences
in Nap between the different treatments were low in each year of experimentation. It is
important to notice that although the total N applied with the three doses of compost
ranged from 230 to 690 kg ha-1, only about 5% was initially available.
There were no differences in nitrogen uptake (Nup) between the different
treatments in both growing seasons. Castellanos et al. (2012) recommended nitrogen
applications about 90-100 kg N ha-1 for melon crop in the same conditions. These
recommendations were totally covered just with the N applied with irrigation water, what
could explain the absence of response in Nup with compost application. In 2011, the
compost addition did not affect the N losses by leaching (Nl). In 2012, the lowest losses
of Nl occurred with D3 (7.8 kg ha-1) and the greatest in the unamended plots (15.9 kg ha-
Treatment Nup Nl Nw Nc Nap Ns final Ns initial NM
(kg ha-1)
Growing
season
2011
S 94 a 14 a 89 0.0 89 73 a 61 31 a
S+D1 96 a 10 a 89 1.6 91 89 b 59 45 b
S+D2 104 a 9 a 89 3.3 93 138 c 60 99 c
S+D3 114 a 10 a 89 4.9 94 142 c 58 113 d
Growing
season
2012
S 130 a 16 b 132 0.0 132 65 a 55 24 a
S+D1 138 a 13 ab 132 2.0 134 84 b 65 37 b
S+D2 134 a 8 a 132 4.0 136 115 c 57 65 c
S+D3 143 a 8 a 132 5.9 138 137 d 52 98 d
N mineralization from wine-distillery waste compost
48
1), but significant differences between compost doses were not observed. Nitrogen in soil
at the end of the growing season increased with respect to the dose of compost applied,
as consequence of the organic N mineralized from the compost applied as found by
other authors applying compost from different sources (Iglesias-Jiménez and Álvarez,
1993; Bar-Tal et al., 2004).
Figure 3.6. Average air temperature Ta (a); average maximum air temperature TMa (b);
average minimum air temperature Tma (c) during the field experiment in 2011 and
2012.
CHAPTER 3
49
Nitrogen mineralized (NM) during the melon crop season in the plots amended
with compost was higher than in the control plots and was positively correlated with total
N applied with the compost (r=0.93). In 2011, the NM ranged from 31 to 113 kg ha-1, and
from 24 to 98 kg ha-1 in the growing season of 2012. For all treatments, the mineralization
was higher in 2011 than in 2012. As soil N mineralization is influenced by temperature
(Stanford et al., 1973; Cassman and Munns, 1980), differences in the field air temperatures
along the respective growing seasons could explain this fact (Figure 3.6). In 2012,
average, maximum and minimum temperatures were lower since the compost
application until the middle of May, which could have influenced N mineralization.
During August (coinciding with the end of the growing season), there was also a
decrease in the minimum air temperatures in 2012 with respect to 2011. Moreover, the
higher inputs of N during the growing season of 2012 could have delayed N
mineralization. Castellanos et al. (2013) found a negative relationship between available
N in soil and N mineralization during the growing season of a melon crop. Other authors
have also reported an effect of N in the soil solution on the N mineralization from organic
residues (Fangueiro et al., 2012) and a decrease of the active microbial pools in the
presence of sufficient amounts of N (Fog, 1988).
Based on the results of the N balance, N mineralized (%) from the compost during
the growing season with respect to the total N applied ranged between 5 and 15% and
increased with respect to the dose of compost application. This observation contrasts
with the results obtained in laboratory conditions. The presence of plants in the field
experiment that remove N from the soil results in a competence with the soil
microorganisms for available N which could accelerate N mineralization (Kaye and Hart,
1997).
3.4.3. Prediction of N mineralized in the field.
The suitability of the exponential model to predict N mineralization after compost
addition was evaluated for the two approaches studied: (i) considering the constants
obtained for the soil-compost mixture and comparing the prediction directly with the
values observed in the field amended plots; (ii) considering the constants obtained for
the compost and comparing the prediction with the values observed in the amended
plots after subtracting the mineralization observed in the unamended plots. The
predicted mineralized N as well as the observed values for the unamended soil and for
the compost-amended soil using both approaches are shown in Table 3.5. The error was
calculated in each case as proposed by Cabrera and Kissel (1988):
[(Predicted NM – Observed NM)/ Observed NM] x 100 (7)
N mineralization from wine-distillery waste compost
50
In 2011, the model made a good prediction of NM in the unamended plots.
However, in 2012 an overprediction of 31% was observed. Using the same model,
Cabrera and Kissel (1988) observed overpredictions from 114 to 343% of N mineralized in
fields cropped with sorghum and explained this error to an incorrect adjustment of the
water content factor. Meanwhile, Zinati et al. (2007) obtained a better prediction for
sugar beet production, with underestimations or overpredictions of less than 10%.
Table 3.5. Predicted values of N mineralized (NM) and observed values in field plots.
S: unamended soil; S+D: soil-compost mixture; D: compost.
Generally, the predicted amounts of N mineralized in the compost-amended soil
using the exponential model exceeded the amounts measured in the field through the
N balance, as reflected in Figure 3.7. However, high correlations were observed between
predicted NM and observed NM for both approaches, similar or better than the obtained
in other studies (De Neve and Hofman, 1998; Mikha et al., 2006; Cordovil et al., 2007). The
exponential model predicted better the amount of NM in 2011 than in 2012. The
overestimations of the model might be due to other factors rather than air temperature
or soil moisture that could have had an influence on soil N mineralization. The difference
in other environmental factors between laboratory and field conditions (nitrate
accumulation in the soil, plant root effect, temperature and moisture fluctuations) could
deteriorate the prediction of the model (Delphin, 2000).
When considering the mixture of soil and compost, the average error for all the
doses was 16% in 2011 and 60% in 2012. On the other hand, when subtracting the NM
Days Treatment NM predicted NM observed Error
(kg ha-1) (kg ha-1) (%)
Growing
season
2011
107
S 30 31 -2
S+D1 65 45 44
S+D2 94 99 -5
S+D3 122 113 8
D1 40 14 180
D2 62 68 -9
D3 89 82 8
Growing
season
2012
119
S 31 24 31
S+D1 75 37 105
S+D2 96 65 49
S+D3 125 98 27
D1 42 13 229
D2 63 41 54
D3 91 74 22
CHAPTER 3
51
from the soil to calculate the NM derived only from compost, the overestimation
increased, being 60 and 102% for 2011 and 2012 respectively. This reflects that it is more
accurate to consider the mixture of soil and compost rather than isolating the substrate
applied. As the soil type influences highly N mineralization in soils (Chae and Tabatabai,
1986) and the addition of organic matter through organic materials could have an
influence on soil organic matter mineralization (Fontaine et al., 2003), the soil-organic
substrate interactions should be taken into account in order to evaluate the
mineralization of N in amended soils.
Figure 3.7. Comparison between estimated and measured mineralized N from the
compost or from the soil-compost mixture. The 1:1 line is shown as a dashed line
.
3.5. CONCLUSIONS.
The N mineralization pattern of the wine-distillery waste compost suggests that it is a
mature compost very resistant to mineralization, with a slow release of inorganic N. For
this reason, when this compost is applied to agricultural soils, a small proportion of the
total N applied is available for the crop or susceptible of being leached in the short term,
even at high doses of application. This should be taken into account when establishing
application limits by the administrations.
The experimental data obtained in the laboratory incubation fitted well the
model of N mineralization generally accepted and allowed to estimate the potentially
mineralizable organic N (N0) and the constant of mineralization (k). The model obtained
(adjusted for real field temperature) generally overpredicted N mineralization in the field
obtained through a N balance, probably due to the influence of other factors that were
N mineralization from wine-distillery waste compost
52
not considered in the model, as nitrogen in soil, plant uptake or moisture/temperature
fluctuations. The model adjustment and the prediction of N mineralization were better
when considering the soil-compost mixture rather than when isolating the organic
substrate, which indicates the important role of the soil to evaluate N mineralization from
the agricultural use of organic wastes.
Acknowledgments
This work was funded by the Instituto Nacional de Investigación y Tecnología Agraria y
Alimentaria, Spanish Agency (RTA2010-00110-C03). The authors would like to thank
Agricompost S.L. for providing the compost and the personal of `La Entresierra´ field
station for their technical assistance during the field experiment.
CHAPTER 4:
WINE-DISTILLERY WASTE COMPOST ADDITION
TO A DRIP-IRRIGATED HORTICULTURAL CROP
OF CENTRAL SPAIN: RISK ASSESSMENT
Requejo, M.I., Cartagena, M.C., Villena, R., Arce, A., Ribas, F., Cabello, M.J., Castellanos M.T.
(2014). Biosystems Engineering 128: 11-20.
CHAPTER 4
55
4.1. ABSTRACT.
The assessment of environmental effects is important before agricultural use of organic
wastes, especially in irrigated agricultural areas vulnerable to nitrate pollution. A field
study with melon crop was conducted during 2011 and 2012 in central Spain, using
different rates of wine-distillery waste compost to quantify nitrogen (N) leaching under
two regimes of irrigation. To evaluate the groundwater pollution risk associated with these
practices, some environmental indices were used to determine the variation in the
quality of drinking water (Impact Index (II)) and in the nitrate concentration of the
groundwater (Environmental Impact Index (EII)). To combine environmental together
with yield parameters, the Management Efficiency (ME) was calculated. Considering II
and EII, compost addition together with adjusted irrigation (90-100% ETc) do not represent
a risk to increase groundwater contamination, and, in high doses, may contribute to
remove nitrate from drainage water and improve the quality of the drinking water. In
contrast, applying large amounts of compost together with an excess of water (120%
ETc) increased N leaching significantly and poses a higher risk of groundwater
contamination. The rate corresponding to 13 t ha-1 of compost together with an efficient
irrigation is sufficient to achieve the highest yield, does not exceed the maximum
allowable limits established by II and EII and has an adequate ME.
N leaching from wine-distillery waste compost
56
4.2. INTRODUCTION.
The winery industry is widespread in Mediterranean countries, especially in Spain, France
and Italy which together represent almost 50% of the total worldwide wine production
(FAO, 2013). Within the region of Castilla-La Mancha (central Spain), 468000 ha under
vines hold most of the wine production in the country, producing more than 1.5 Gl per
year (MAGRAMA, 2012a). During the winemaking process, solid residues like grape stalks
are generated, as well as grape marc and wine lees as by-products. Grape marc and
wine lees are usually sent to alcohol distilleries to extract alcohol and calcium tartrate,
generating exhausted grape marc and lees cake.
The addition of organic matter to soil to improve soil fertility has been widely
studied (Hernando et al., 1989; Stevenson, 1994; Giusquiani et al., 1995). Previous studies
have demonstrated positive effects in plant growth after the application of wine and
distillery wastes or composts in agriculture (Mariotti et al., 2000; Masoni et al., 2000; Ferrer
et al., 2001; Manios, 2004; Carmona et al., 2012). Descriptions have also been made of
the suppressive character of these materials against plant pathogens (Borrero et al.,
2004; Segarra et al., 2007; Rivera and Aballay, 2008).
The nature of the organic wastes applied to an arable area is closely related to
the location of the waste generating industries. Melon is one of the most important
horticultural crops traditionally cultivated in Castilla La-Mancha and almost the whole
area dedicated to this crop is dependent on irrigation (MAGRAMA, 2012b), because
semiarid conditions predominate during the growing season. This crop is grown mainly in
zones declared vulnerable to nitrate (NO3) pollution (Directive 91/676/EEC), around three
protected hydrological units (UH) designated by the local administration: UH.04.04
`Mancha Occidental´, UH.04.05 `Campo de Calatrava´, and UH.04.06 `Campo de
Montiel´. The local administration limits nitrogen (N) fertilization to 115 or 135 kg ha-1
according to the zone of application (DOCM Nº32, 2010). Recent research undertaken
in this area demonstrated that the high NO3 concentration in irrigation water, which is
around 100-150 mg L-1, together with the mineral N applied with the fertilizer, result in N
supplies that usually exceed 250 kg ha-1 (Castellanos et al., 2013).
Drip-irrigation strategies in combination with plastic-much have been adopted in
this area for crop production because this system provides precise water management,
increases yields and controls weeds (Lament, 1993; Hartz, 1996). However, the high salt
content in irrigation water requires a `leaching fraction´ to wash out the salts
accumulated in the soil (Doorenbos and Pruitt, 1977; Ayers and Westcott, 1985) to avoid
problems of crop development. In this condition, N leaching is quite probable and thus
CHAPTER 4
57
the increasing pollution of the aquifers is a major concern. This is especially relevant given
that the aquifers in question are the main source of water supply for human consumption
in the area (Miner-Moptma, 1994).
Improving irrigation and fertilizer management is essential to reduce N leaching
in vulnerable zones. Recent research on melon in semiarid regions has focused on
optimization of water and mineral N fertilizer requirements in irrigated systems, concluding
that increasing the amount of water applied does not result in higher yields
(Panagiotopoulus, 2001; Kirnak et al., 2005) and has a negative effect on crop yield and
quality (Pew and Gardner, 1983). It is also assumed that a moderate deficit of water does
not affect the crop yield negatively (Cabello et al., 2009). Previous studies carried out in
the study area for melon concluded that N applied with irrigation water (approximately
100 kg N ha-1 for all the growing season) is sufficient for adequate plant growth
(Castellanos et al., 2011). Less information is found in the literature regarding the use of
organic fertilizers in irrigated horticultural crops. Bryan et al. (1995) reported an
improvement in growth and yield of tomato after application of municipal solid waste
compost and a trend to increased yield at higher rates of irrigation. Lopedota et al. (2013)
studied the effect of different organic fertilizers under two irrigation rates on melon yield
and found an increase in yield with a higher irrigation, but no differences between the
different fertilizer treatments.
In this critical scenario, it is crucial to assess the environmental effects and best
management options for the agricultural use of organic wastes, especially regarding to
N. Compost N forms added to soil are mostly organic, so organic N can be mineralized
during the crop period and thus be taken up by the plants, immobilized, or leached.
Some authors have reported a slow N mineralization rate and even an initial
immobilization using incubations of soil amended with different winery and distillery
wastes (Flavel et al., 2005; Bustamante et al., 2007) or vermicompost derived from
exhausted grape marc (Paradelo et al., 2011). In experiments with open-air lysimeters,
the addition of winery sludge to maize and soyabean was shown to increase N leaching
(Masoni, et al, 2000; Mariotti et al., 2001). However, there is little information relating the
use of composts derived from winery and distillery wastes in field conditions.
The evaluation of the use of organic fertilizers in horticultural crops of central Spain
involves the study of agronomic and environmental responses, especially in zones
vulnerable to N pollution. The objectives if this work were: (1) to quantify melon yield,
drainage and N leaching response to different doses of wine-distillery waste compost
under two different irrigation strategies; (2) to evaluate the risk of groundwater NO3
contamination using environmental indices.
N leaching from wine-distillery waste compost
58
4.3. MATERIAL AND METHODS.
4.3.1. Experimental design and compost treatments.
The field experiment was conducted during the growing seasons of 2011 and 2012, from
April to August, at the experimental field station `La Entresierra´, located in Ciudad Real,
central Spain (3° 56’ W; 39° 0’ N; 640 m altitude). The soil is a shallow sandy clay loam,
classified as Petrocalcic Palexeralfs (Soil Survey Staff, 2010), with a discontinuous and
fragmented petrocalcic layer below the first 0.60 m. Cracks within this layer enhance
vertical permeability locally. Its generic horizons are: Ap (0-0.3 m), Btk (0.3-0.6 m), Ckm
(0.6-1.0 m), Cck (1.0 to >1.5 m). The presence of plant available water is therefore
restricted to the top 0.60 m. The volumetric water content for the first 0.3 m was 0.228 m3
m-3 at field capacity (soil matric potential of -0.03 MPa) and 0.121 m3 m-3 at wilting point
(soil matric potential of -1.5 MPa) and from 0.3 to 0.7 m it was 0.430 and 0.211 m3 m-3,
respectively.
Soil characteristics (Ap horizon) are provided in Table 4.1. The field plots had been
sown for the previous three years with non-irrigated winter wheat (Triticum aestivum L.)
receiving neither organic amendments nor fertilizers.
Table 4.1. The physicochemical characteristics of the soil (Ap horizon) at the field
experimental site.
Parameters
pH 8.4
EC (mS cm-1) 0.2
Organic matter (g kg-1) 24.0
Available P (mg kg-1) 22.1
Available K (mg kg-1) 410.7
Available Ca (mg kg-1) 1649.3
Available Mg (mg kg-1) 461.2
N (Kjeldahl) (g kg-1) 1.0
NO3--N (mg kg-1) 15.3
NH4+-N (mg kg-1) 1.7
Climatic data of the melon growing season are shown in Table 4.2. The area is
characterized by a Mediterranean climate, with a strong continental character and
widely fluctuating daily temperatures. The rainfall season occurs outside the melon
growing period.
CHAPTER 4
59
Table 4.2. Monthly climatic data of the experimental area during the growing
season of 2011 and 2012.
The compost used in this experiment was obtained from a composting plant
located in Socuéllamos (Ciudad Real) which collects residues from the wineries and
distilleries around this area and through an aerobic degradation produces commercially
available compost composed of grape stalks, exhausted grape marc and lees cake. The
main properties of the compost used are shown in Table 4.3. This compost fulfilled the
criteria established by Spanish legislation concerning the use of organic materials in
agriculture.
Table 4.3. Characteristics of the compost used in 2011 and 2012.
Parameters 2011 2012
pH 9.2 9.8
EC (mS cm-1) 1.2 1.0
N Kjeldahl (g kg-1) 32.9 33.0
NH4+-N (g kg-1) 1.5 1.8
NO3--N (g kg-1) 0.2 0.3
Total P (g kg-1) 6.8 7.6
Total K (g kg-1) 33.7 21.1
Organic matter (g kg-1) 545 600
C:N ratio 10.1 11.2
Four treatments were set up in both years: a control with no compost addition
(D0), and three doses of wine-distillery waste compost application corresponding to 7
(D1), 13 (D2) and 20 (D3) t ha-1. A randomized complete-block design was used, with the
Climatic parameters April May June July August
Growing season 2011
Minimum air temperature (°C) 7.9 10.2 13.8 14.8 18.4
Maximum air temperature (°C) 23.6 26.3 32.3 35.0 35.2
Average temperature (°C) 15.7 18.2 23.0 24.9 26.8
Rainfall (mm) 61.5 47.7 5.1 - 0.3
Relative humidity (%) 69.9 63.6 45.5 49.5 57.3
Growing season 2012
Minimum air temperature (°C) 4.3 8.7 13.0 14.2 14.7
Maximum air temperature (°C) 16.8 26.9 33.4 35.6 36.4
Average temperature (°C) 10.5 17.8 23.2 24.9 25.5
Rainfall (mm) 54.1 27.5 - 6.7 -
Relative humidity (%) 76.6 59.0 50.7 48.1 50.5
N leaching from wine-distillery waste compost
60
dose of compost as factor of variation. Each treatment included four replicate plots
(12x15 m) formed by ten rows with eight plants each.
The compost was applied on 20 April in 2011 and 19 April in 2012 localized in the
crop row with a tow-behind spreader and immediately incorporated into the soil at 10-
20 cm depth with a cultivator. A `Piel de sapo´ type melon (Cucumis melo L. cv. Trujillo)
was used as crop. Each year of experimentation, melon seeds were germinated from
late March under greenhouse conditions, until they had sprouted two or three true
leaves. Subsequently, the seedlings were transplanted in the field plots approximately 20
days after compost incorporation (11 May 2011 and 9 May 2012) into plastic mulch
(transparent polyethylene, 20 µm thick, 110 cm width) at an areal density of 4444 ha -1
(1.5 by 1.5 m).
The low rainfall during the melon crop season (total rainfall from May to August of
53.1 in 2011 and 34.2 in 2012) and the high evapotranspiration rates (between 700 and
800 mm from May to August) make irrigation necessary. The irrigation system consisted of
one drip line per crop row with emitters every 0.5 m, providing water at 2 L h-1 per emitter.
In both years, groundwater from a well near the experimental plots was used. The
irrigation water quality was measured weekly to determine the amount of nitrates
applied (Table 4.4).
Table 4.4. Characteristics of the irrigation water used in 2011 and 2012. Values represent
means of weekly measurements throughout the experiment.
Parameters 2011 2012
pH 7.9 7.9
EC (mS cm-1) 3.1 2.9
K+ (mg L-1) 3.6 3.5
Mg+ (mg L-1) 172.0 169.0
Ca2+ (mg L-1) 300.0 275.0
NO3- (mg L1) 116.0 125.0
NH4+ (mg L-1) 0.1 0.1
SO42- (mg L-1) 1233.7 1195.5
Cl- (mg L-1) 203.5 188.8
All treatments received 120 kg ha-1 of phosphorus fertilizer (phosphoric acid) for
the season, injected daily through fertigation. Nitrogen fertilizers were not applied due to
the high NO3 concentration in the irrigation water. A regular programme of disease and
insect control was followed throughout the growing period, according to standard
management practices.
CHAPTER 4
61
Figure 4.1. Irrigation pond located at the experimental field station.
4.3.2. Irrigation and drainage.
After transplanting, all plots initially received 28 mm of water to facilitate crop
establishment. The irrigation schedule was calculated from 27 to 90 days after
transplanting (DAT) in 2011, and from 20 to 97 DAT in 2012. The crop evapotranspiration
(ETc) was calculated daily using FAO method as ETc= Kc x ET0 (Doorenbos and Pruitt,
1977). ET0 is the reference evapotranspiration and was estimated by the FAO Penman-
Monteith method (Allen et al., 2002), using daily data from a meteorological station sited
500 m from the experimental field. Kc is the crop coefficient obtained in previous years
for melon crop in the same conditions (Ribas et al., 1995). In 2011, two irrigation strategies
were evaluated and consisted of a single and daily irrigation of 90 and 120% of the ETc,
respectively. In 2012, a single irrigation method of 100% ETc was established. Previous
research in the same area demonstrated that 90 and 100% ETc should be considered as
adjusted irrigation to meet the plant requirements whereas 120% ETc should be
considered as overwatering (Cabello et al., 2009). Rainfall was negligible, so the water
to be applied in each treatment every week was calculated as the ETc of the previous
week divided by the efficiency of the system, which considers the salt tolerance of the
crop, the quality of irrigation water, the soil texture and the homogeneity of the irrigation
system (Rincón and Giménez, 1989), estimated at 0.81 in our conditions. This result was
then divided by the number of days to obtain the daily irrigation needs (estimated
irrigation). To ensure the efficiency of the system, all irrigation treatments had to be
increased by 23%. Due to that, irrigation and ETc were not coincident. For each irrigation
treatment, daily irrigation rates were the same for all plots. Water meters were installed
at the outflow of each electrovalve supplying the water to obtain the actual irrigation
applied.
N leaching from wine-distillery waste compost
62
To estimate drainage (D), the following water balance was used
D = Irr + Prec - ETc - Rf ± ∆ ∫ θvz=0.6
0dz (Doorenbos and Pruitt, 1977), where Irr is the
irrigation, Prec is the precipitation, Rf is the runoff, which was assumed to be negligible,
and ∆θv is the change in soil water content measured between two consecutive weeks
to a depth Z of 0.60 m. A tube was installed in the centre of each plot, between two
plants in the row and 0.375 m from the drip line (Castellanos et al., 2013), to measure θv
on a weekly basis with a probe (Diviner, 2000) based on Frequency Domain Resonance
(FDR).
4.3.3. Nitrogen leaching.
Porous ceramic cups were used to extract samples of the soil solution every week
simultaneously with the FDR measurements. The cups were installed vertically in each plot
between two consecutive plants using a mechanical soil drill to a depth of 0.60 m and
0.375 m from the irrigation line. A soil slurry was made by mixing the extracted soil
(previously sieved) with water. The slurry was put into the hole before inserting the suction
cup and then around the tube to maximize the contact between the cup and the soil.
A suction of −70 kPa was applied to the cups with a vacuum pump to extract the soil
water. Water samples were stored in a freezer until analysis. The NO3 concentration in the
soil water was measured with an ion-selective electrode. The N leaching to below 0.60 m
was calculated as the product of nitrate concentration and D.
Figure 4.2. Placement of the Diviner probe and the suction cup in the plots (left) and
detail of the suction cup (right).
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63
4.3.4. Environmental indices.
Taking into account N leaching and D, two environmental indices were calculated to
evaluate the risk of groundwater pollution. These indices had previously been used by
Castellanos et al. (2013) to assess the effect of different rates of N fertilizer in a fertigated
melon crop. The first one, the Impact Index (II) was defined as the ratio between the NO3
concentration in the leachate and the maximum concentration considered by the
European Drinking Water Directive (Council Directive 98/83/EC) established at 50 mg L-1.
The second index, the Environmental Impact Index (EII), was expressed as the ratio
between the NO3 concentration in the leachate and the concentration in groundwater.
4.3.5. Fruit yield, management efficiency and nitrogen uptake.
Melons were harvested when a significant number of fully ripe fruits were observed in the
field and harvesting was carried out weekly. Each melon was weighed at harvest to
determine the total fruit yield (FY).
To consider both agronomic and environmental indices with a unique index,
Castellanos et al. (2013) defined the Management Efficiency (ME) as the ratio between
FY and the amount of N leached. This index was calculated in this study to evaluate the
best management practice regarding the use of wine-distillery waste compost in
vulnerable zones.
Four plants per treatment were sampled from each plot around 90 days after
transplanting (DAT). Sampling was performed so as to avoid border effects. The parts
from the individual plants (leaves, stems and fruits, including ripe fruits) were separated
and weighed to obtain the fresh mass. The dry mass of these separated organs were
determined following drying to constant mass in an oven at 80°C. The dried samples
(leaves, stems and fruits) were ground to a fine powder using an electrical grinder.
Sample N content was determined using the Kjeldahl method (Association of Official
Analytical Chemists, 1990). The N accumulated by each organ was obtained as the
product of N concentration and biomass. The N uptake was determined as the sum of
the N accumulations of each above-ground organ of the plant.
4.3.6. Statistical analysis.
Analysis of variance (ANOVA) was carried out separately for each year to determine the
effects of the compost treatments (D0, D1, D2 and D3) on the measured parameters.
Significant statistical differences of variables between the different treatments were
N leaching from wine-distillery waste compost
64
established by the Tukey test (p≤0.05). Regression analysis was conducted to identify
relationships between N leaching and N applied with compost.
4.4. RESULTS AND DISCUSSION.
4.4.1. Irrigation and drainage.
Figure 4.3. Cumulative crop evapotranspiration (ETc), estimated irrigation, actual
irrigation and drainage (D) of each dose of compost with various irrigation rates: (a)
120% ETc, 2011; (b) 90% ETc, 2011; (c) 100% ETc, 2012.
CHAPTER 4
65
Figure 4.3 shows the cumulative ETc, the estimated irrigation, the actual irrigation applied
and drainage obtained through the water balance. The ETc for all the irrigation schedule
was 282.1 mm in 2011 and 386.6 mm in 2012, while irrigation amounted to 441.9 mm and
341.3 mm for the 120% ETc and 90% ETc rates respectively in 2011, and 461 mm in 2012.
Throughout the irrigation period, when irrigation was higher than ETc, there was drainage
and consequently, N leaching occurred.
In 2011, the total amount of NO3-N applied with irrigation during the total schedule
period was 115.8 and 89.4 kg ha-1 for the excessive (120% ETc) and adjusted (90% ETc)
irrigation strategies respectively. In 2012, as the NO3 concentration in the groundwater
increased (Table 4.4) and the irrigation period was longer, the NO3-N applied with water
at 100% ETc was 131.6 kg ha-1.
Drainage followed a similar pattern for all the tubes corresponding to the same
irrigation rate, and no significant differences were obtained between compost
treatments for either year. In 2011, the mean D during the entire irrigation schedule was
151.8 mm and 72.5 mm for 120% ETc and 90% ETc irrigation doses respectively whereas in
2012, with an irrigation adjusted to the 100% ETc, D was 92.4 mm. Knowles et al. (2011)
also reported non-significant differences in the volume of drainage when applying
biosolids compared to the control.
4.4.2. Nitrate concentration in the soil solution and nitrogen leaching.
The evolution of the concentration of NO3 in the soil solution during the crop cycle
showed a different pattern with respect to the amount of water applied (Figure 4.4 a, b,
c). The maximum concentrations were achieved during the first stages of the crop
development in all cases. During this stage, concentrations around 200 mg L-1 or even
higher were obtained in the water samples due to mobilization of the nitrate contained
in soil, water and compost at the beginning of the irrigation schedule. During the
beginning of the vegetative stage of melon crop, N demand is low (Cabello et al., 2011),
so mineral nitrogen remains in the soil until drainage occurs. Flores et al. (2005) reported
the same fact with a pepper crop cultivated with mineral and organic fertilizers. The
addition of compost did not increase the concentration of NO3 in the soil in this initial
phase of the plant growth, due to the low NO3 concentration initially present in this
material (Table 4.3). Furthermore, other authors stated that immobilization processes
initially dominate over mineralization when this kind of waste is applied to soil (Flavel et
al., 2005; Bustamante et al., 2007; Paradelo et al., 2011).
Nitrate concentration decreased with time progressively due to plant growth and
N uptake. According to Castellanos et al. (2012), the highest N uptake rate for melon
N leaching from wine-distillery waste compost
66
plant takes place from 30-35 to 70-80 DAT, coinciding with fruit development. About 50
DAT, all the plants have set fruits and the first set fruits are in the second half of their
development. The lowest concentrations of NO3 were measured from this time until about
75 DAT, and in some cases were close to zero. When applying an excess of irrigation
(120% ETc), a different pattern was observed with the highest dose of compost
application (D3) where soil solution concentrations remained above 50 mg L-1 throughout
this period. By this time, compost mineralization should have started and the treatment
with the highest dose of compost could have an excess of NO3 susceptible to be lost by
drainage. It is well known that N leaching losses occur when there is a high amount of
NO3 in soil in conjunction with a high drainage volume (Di and Cameron, 2002).
Figure 4.4. NO3 concentration in the soil solution (a, b, c) and cumulative N leaching (d,
e, f) of each dose of compost with various irrigation rates: (a, d) 120% ETc, 2011; (b, e)
90% ETc, 2011; (c, f) 100% ETc, 2012. The bars represent the least significant difference (p
≤ 0.05).
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67
From 75 DAT until the end of the crop period, the concentration of NO3 in the soil
solution tended to increase in all cases because N accumulation in the plant became
constant (Castellanos et al., 2012) and compost addition had an influence. In 2011, two
different patterns were observed according to the water regime. With an overwatering
(Figure 4.4 a), NO3 concentration increased significantly when applying compost, and
differences were greater with an increased dose applied. On the other hand, with
adjusted irrigation (Figure 4.4 b), no significant differences were found between the
control and the treatments with compost application, but a tendency to increase NO3
concentration in the soil solution was observed for D0 and D1 (p≤0.11). In 2012 (Figure 4.4
c), NO3 concentration followed the same pattern as the adjusted strategy of 2011, being
in this case significantly higher for treatments D0 and D1 from 69 to 97 DAT, in comparison
with D2 and D3.
Nitrogen leaching (Figure 4.4 d, e, f) followed a similar pattern as the drainage,
but the differences between treatments increased due to differences on the NO3
concentrations in the soil solution. With overwatering (Figure 4.4 d), N losses increased
progressively from 50 DAT until the end of the study and the control remained below the
compost treatments throughout this period. The total amount of N lost by leaching
ranged from 10.2 (D0) to 39.7 (D3) kg N ha-1. If an excess of the crop water needs is
applied, the water that the plants do not use could drain quickly to the subsoil due to the
high permeability of the topsoil in this site. Consequently, the NO3 contained in the topsoil
or in the irrigation water, as well as organic N mineralized during the crop period, is
transported through drainage to deeper soil layers, being susceptible to leaching.
Differences between treatments could be attributed to mineralization of the organic N
contained in the compost, which increased with the dose applied. An influence of the
rate of organic fertilizer application on N soil levels and leaching has been reported by
several authors (Hartl et al., 2003; Correa et al., 2006).
With respect to the adjusted irrigation (Figure 4.4 e, f), the amount of N leached
from the beginning of the irrigation schedule until 76 DAT remained constant and
accounted for <50% of the total N leached during the whole irrigation period. With this
type of irrigation, drainage losses were very small or even did not occur during most of
the growing season, until the senescence of the melon crop is reached, when the
amount of drainage started to increase. This is the time when the greatest increases of N
leached occur, being higher for the control (D0) than for the other treatments. In 2011,
with 90% ETc, N leaching losses were between 9.4 (D2) to 13.9 (D0) kg N ha-1, but no
significant differences were found. In 2012, these losses ranged between 7.8 (D3) to 15.9
(D0), with D0 and D1 significantly higher than D2 and D3 (p≤0.05). Díez et al. (1997)
reported the same fact using municipal solid waste compost in a maize-wheat crop
N leaching from wine-distillery waste compost
68
rotation under conventional or efficient irrigation. They found nitrate leaching losses were
less than the control when using efficient irrigation due to a decrease in the
concentration of nitrates in the soil solution.
Castellanos et al. (2013) found N leaching amounts from 1.5 to 59.6 kg N ha-1 for
all the melon crop season when applying different amounts of mineral N via fertigation
(11 to 393 kg N ha-1). In their study, N leaching losses increased exponentially with the
amount of N available (considering mineral N from the soil, the fertilizer and the irrigation
water) during the crop cycle, obtaining losses around 30 kg N ha-1, a little lower than the
highest value obtained in this study, when N available was 314 kg ha-1. This means that
although available N applied with compost is initially insignificant (around 5 kg ha-1 with
the highest dose) and total N available amounted to 180 kg ha-1, mineralization of
organic N contained in compost during the crop cycle results in an important pool of
available N in soil.
Another factor that can have an effect on nitrate leaching is plant N uptake (Di
and Cameron, 2002; Cameron et al., 2013). However, in this experiment compost
addition did not increase plant N uptake in either water regime (Table 4.5). Thus, the
reduction of NO3 losses by leaching when applying compost with irrigation adjusted to
the plant needs cannot be attributed to N removal by the crop.
In some cases, the water input to the soil may stimulate NO3 losses by
denitrification rather than increasing leaching (Bar-Yosef and Kafkafi 1972; Di et al., 1998).
Application of organic fertilizers such as compost or manures has been shown to
enhance losses by denitrification in soil by increasing soil carbon availability (Vallejo et
al., 2006). The basic soil pH and elevated temperatures during the crop season could
favour denitrifier activity in our experimental conditions (Granli and Bockmanm, 1994;
Stanford et al., 1975). Otherwise, a primary requirement for denitrification is the NO3
concentration in the soil solution that depend on mineralization and nitrification rate,
plant N uptake, microbial immobilization and NO3 movement by leaching and diffusion
(Tiedje, 1988). With an overwatering regime, NO3 from soil, water or that mineralized from
compost is leached to deeper soil layers from the first stages of the crop development.
By contrast, the adjusted irrigation management maintained a low drainage until the
end of the crop cycle, so higher levels of NO3 are accumulated in the topsoil, and
together with the high input of soluble carbon, could stimulate the denitrification process.
Sánchez-Martín et al. (2010) showed an increase in the average N denitrification rate of
7 mg N m-2 day-1 when applying digested pig slurry in a drip-irrigated melon crop with
similar environmental conditions to those in our study. The use of compost for removing
nitrate from drainage water by denitrification has been reported by other authors (Tsui et
al., 2007; Su and Puls, 2007).
CHAPTER 4
69
4.4.3. Environmental indices.
The relationship between the environmental indices calculated and the amount of N
applied with the different rates of compost was studied. When II was plotted against total
N applied with compost, two linear regressions were obtained for the overwatering
treatment (r=0.97), and for the treatments with adjusted irrigation (r=0.78) (Figure 4.5 a).
For the treatments with adjusted irrigation the gradient is negative and shows a decrease
of II as total N applied increased. Using an excess of irrigation the gradient is positive and
shows an increase of II as total N applied increased. For adjusted irrigation treatments,
D0 and D1 presented a nitrate concentration in the leachate above 50 mg L-1, and
therefore, have an II>1 and are located above the horizontal dotted line. By contrast, D2
and D3 gave concentrations which were beneath the maximum allowable value.
Figure 4.5. Impact Index (II) (a) and Environmental Impact Index (EII) (b) versus total N
applied with compost. The discontinuous line represents the maximum limit equal to 50
mg L-1 established by the Drinking Water Directive (a) or a NO3 concentration equal to
the concentration of NO3 in the groundwater (b).
N leaching from wine-distillery waste compost
70
Both treatments, D0 and D1, with adjusted irrigation and quality previously shown,
may be considered practices that pose a risk to the use of groundwater as drinking water,
as they exceeded the limit established by the Drinking Water Directive, mainly due to the
high concentration of NO3 in irrigation water and in the soil profile. Castellanos et al.
(2013) obtained a value of 0.99 for II, considering an optimal dose of N available around
160 kg ha-1, slightly below the D0 treatment of our study.
Considering the excess irrigation, treatments D1, D2 and D3 presented II>1, and
only D0 showed a nitrate concentration beneath the maximum allowable value. It can
be concluded that the addition of wine-distillery waste compost to soil under an excess
of irrigation could contribute to a deterioration of the quality of groundwater for human
consumption.
Figure 4.5 b shows the relationship between the EII and the total N applied with
compost. As occurred with II, two linear regressions were obtained according to the
irrigation strategy, also showing good correlations (r=0.97 and r=0.68 for overwatering
and adjusted irrigation respectively). As nitrate concentration in groundwater in the area
of study exceeds the limit of 50 mg L-1, the EII index is an approach to determine whether
agricultural practices contribute to increase groundwater pollution. All treatments with
adjusted irrigation showed nitrate concentrations in the leachate below the nitrate
concentration in the groundwater, obtaining an EII<1. In these conditions of irrigation and
total N applied, any treatment means a risk of increasing groundwater contamination.
For the irrigated plots with an excess of water, only D3 presented an EII>1 and the rest of
treatments gave concentrations which were below that of groundwater. So in these
conditions, only the D3 treatment supposes a risk of groundwater contamination.
4.4.4. Fruit yield, management efficiency and N uptake.
In 2011 and 2012, there were four harvests conducted weekly starting at 76 and 78 DAT
respectively. In both years, compost addition had a positive effect on FY (Table 4.5). In
2011, under both irrigation strategies, FY was significantly lower in D0 with respect to D2
and D3, and no significant differences were found between the control and D1. In 2012,
with adjusted irrigation, yield was significantly higher when applying compost and no
effect between the different doses was found.
The ME expresses the fruit yield per kg of N leached. With overwatering, the
highest ME was obtained for the control, because although FY increased with the
addition of compost, the increment of N losses by leaching was higher. When an adjusted
irrigation strategy is used, ME increased with increasing rate of compost applied because
CHAPTER 4
71
there is not only an increment in the fruit obtained, but also a reduction in the N leached
to groundwater.
Although no statistically comparisons were done between the different irrigations
in 2011, it can be noticed that an excess of irrigation did not increase FY in any of the
treatments studied. This fact is widely reported in the literature (Kirnak et al., 2005; Cabello
et al., 2009) for melon crop fertilized with mineral N. With respect to the use of organic
fertilizers under different regimes of irrigation, Lopedota et al. (2013) found an increase in
yield when applying higher irrigation rates to a melon crop fertilized with manure,
anaerobic digestate or composted municipal organic waste compared to a deficit
irrigation strategy. Abdel-Mawgoud (2006) studied the effect of different rates of drip-
irrigation and compost in green-bean and found a positive and cumulative effect of
both factors on crop yield.
Table 4.5. Fruit yield, N leached, management efficiency (ME) and N uptake in the
melon crop in 2011 and 2012.
Within each column, year and irrigation rate, means followed by the same letter are not
significantly different at p≤0.05.
Plant N uptake was not affected by any of the doses of compost applied in both
irrigation strategies (Table 4.5). In our conditions, the large amounts of NO3-N applied with
irrigation water (116 and 89 kg N ha-1 for 120 and 90% ETc in 2011; 132 kg N ha-1 for 100%
ETc in 2012) seem to be enough to meet crop needs. Castellanos et al. (2012)
recommended nitrogen applications close to 90-100 kg ha-1 for melon crop in the same
Treatment
Fruit yield
(t ha-1)
N leaching
(kg ha-1)
ME
(t kg-1)
N uptake
(kg ha-1)
Growing
season
2011
120% ETc
D0 28.3 a 10.2 a 2.8 99.3 a
D1 31.8 ab 26.4 b 1.2 99.5 a
D2 36.5 b 33.9 bc 1.1 108.5 a
D3 36.8 b 39.7 c 0.9 109.0 a
90% ETc
D0 29.6 a 13.9 a 2.1 94.4 a
D1 31.9 ab 10.4 a 3.1 96.2 a
D2 36.4 b 9.4 a 3.9 104.2 a
D3 36.1 b 10.3 a 3.5 113.8 a
Growing
season
2012
100% ETc
D0 49.9 a 15.9 b 3.1 130.1 a
D1 55.2 b 13.2 ab 4.2 138.0 a
D2 55.0 b 8.2 a 6.7 134.2 a
D3 56.1 b 7.8 a 7.2 142.8 a
N leaching from wine-distillery waste compost
72
conditions. Other beneficial effects of the use of this compost seem to have influenced
the increase in crop yield.
4.5. CONCLUSIONS.
The irrigation rate plays an important role in N losses by leaching when applying wine-
distillery waste compost to irrigated crops in vulnerable zones. Excessive irrigation
throughout crop development resulted in higher N leaching in the treatments with
compost addition that increased with the dose applied. On the contrary, when an
adjusted irrigation strategy was used, the addition of compost reduced N losses by
leaching with no reduction in crop yield. Compost addition together with irrigation
adjusted to the actual plant needs is recommended. Treatments D2 and D3 with
adjusted irrigation showed a NO3 concentration in the leachate below the limit
established by the Drinking Water Directive as well as giving the highest fruit yields. An
excess of water together with large amounts of wine-distillery waste compost pose a risk
of groundwater contamination.
Acknowledgments
This work was funded by the Instituto Nacional de Investigación y Tecnología Agraria y
Alimentaria, Spanish Agency (RTA2010-00110-C03). The authors would like to thank
Agricompost S.L. for providing the compost and the personal of `La Entresierra´ field
station for their technical assistance during the field experiment.
CHAPTER 5:
WINERY AND DISTILLERY WASTES AS
SOURCE OF PHOSPHORUS IN
CALCAREOUS SOILS
Requejo, M.I., Fernández-Rubín de Celis, M., Martínez-Caro, R., Castellanos, M.T., Ribas, F.,
Arce, A., Cartagena, M.C. (2015). Catena. Under review.
CHAPTER 5
75
5.1. ABSTRACT.
The depletion of phosphate rock reserves makes necessary the finding and
characterization of new phosphorus (P) sources to ensure a sustainable P fertilization. This
research deals with the effect of the application of wastes derived from the winery and
distillery industries on P availability in two calcareous soils. Soils were treated with five
wastes providing different amounts of P: two raw wastes (exhausted grape marc and
grape stalk) and three composts made with different proportions of these materials or
including lees cake as well. The soils were incubated for 16 weeks determining soil pH,
dissolved organic carbon and Olsen P in five samplings along time. At the end of the
incubation, inorganic P was fractionated to NaOH-NaCl-P, citrate-bicarbonate-
dithionite-P (CBD-P), and HCl-P and the P sorption index (PSI) and degree of P saturation
(DPS) were also determined. The increment in Olsen P was firstly positively related to the
amount of P applied, being higher in the composted materials, but along time the effect
of other factors such as the organic matter incorporation influenced these changes. After
16 weeks of incubation, the enhancement of Olsen P produced by the different wastes
began to equalize (7.1 mg kg-1 on average), despite of the different amounts of P applied
with each one. Increasing the P applied with the organic wastes in these soils enhanced
P retained in the sparingly soluble P fraction (HCl-P) whereas the addition of raw materials
slightly released P from this fraction, recovering more P in the most labile pool (NaOH-
NaCl-P). It can be concluded that the winery and distillery wastes had a positive effect
on soil P availability and that P in excess would be retained due to the calcareous nature
of these soils, according to the P saturation of the soil.
P availability from wine-distillery wastes
76
5.2. INTRODUCTION.
The European Union concentrates over 50% of the world wine production, being this
industry mainly located in Mediterranean countries as France, Italy and Spain (FAO,
2013). The modern wine industry generates a huge amount of wastes and by-products
which are an important concern in the grape growing regions due to their seasonality.
The organic wastes produced include the grape stalks (GS) (12%), composed of the
stems obtained after the destemming process; the grape marc (GM) (62%), produced
during the grape press and constituted by skin and seeds; and the wine lees (WL) (14%),
generated in the clarification of the wine fermentation process (Ruggieri et al., 2009).
Grape marc and wine lees are usually sent to alcohol distilleries to extract alcohol and
calcium tartrate, obtaining exhausted grape marc (EGM) and lees cake (LC).
In the recent years there is an effort to develop methods to recover and re-use
these wastes to produce different substances with a market value. The grape skins and
pips contain natural compounds as tannins, tartaric acid or polyphenols which are
commercially appreciated (Louli et al., 2004; Yalcin et al., 2008; Ping et al., 2011). These
materials could also be employed for bioethanol production (Rodríguez et al., 2010) and
for biomass power generation (Gómez et al., 2010). However, the stalks are less valuable
due to its low industrial and economical value (Ruggieri et al., 2009). The recycling of
these wastes as organic fertilizers in agriculture, by direct addition to soil or after
composting, has also been proposed as a good alternative to replenish soil organic
matter and nutrients (Arvanitoyannis et al., 2006). Several studies have evaluated the
effect of the application of these materials on soil characteristics and processes: nitrogen
mineralization (Flavel et al., 2005; Bustamante et al., 2007), soil carbon dynamics
(Bustamante et al., 2010; Paradelo et al., 2011) or immobilization of heavy metals and
pesticides (Andrades et al., 2004; Martínez et al., 2006; Romero et al., 2006). However, as
far as we know there are no studies focused on phosphorus (P) dynamics and availability
in soil from the use of these materials in agriculture.
Agricultural production is highly reliant on the application of P fertilizers. As the
depletion of phosphate rock is becoming an important concern (Cordell et al., 2009; Van
Vuuren et al., 2010), there is an increasing interest to find strategies to mobilize P in soil as
well as to find new P sources (Stutter et al., 2012, 2015). Phosphorus in calcareous semiarid
soils of Mediterranean regions is often a limiting nutrient for crop production due to the
high retention of P in soil constituents, as carbonates and iron oxides, which limit the
availability of P in these soils (Solis and Torrent, 1989; Delgado and Torrent, 2000; Delgado
et al., 2000). On the other hand, the continuous additions of P fertilizer to raise optimum
CHAPTER 5
77
levels for plant growth could lead to P accumulation in surface horizons, increasing the
potential of P losses to surface and groundwater that can cause eutrophication
problems (Sims et al., 2000; McDowell et al., 2001). Recently, with the purpose to close
the P cycle in Europe, the scientific community has proposed to implement a package
of nutrient management strategies known as the 5R strategy (Withers et al., 2015): Realign
P imputs, Reduce P losses to waters, Recycle P in bio-resources, Recover P from waste
and if necessary Redefine our food system.
Organic amendments usually contain important amounts of P (García-Albacete
et al., 2012). Besides being a potential P source, organic amendments provide a
significant amount of easily available carbon resulting in a higher microbial activity
(biomass build-up, turnover, respiration and substrate mineralization) (Kao et al., 2006;
Saha et al., 2008) which affects P cycling. There are also evidences that organic
amendments may influence P dynamics in soil by modifying pH in the soil solution
(Iyamuremye et al., 1996a), by competition between low-molecular organic acids and
phosphates for sorption sites delaying P sorption (Sui and Thompson, 2000) or by inhibiting
P precipitation (Inskeep and Silverthooth, 1988).
To ensure a sustainable P fertilization, the characterization of alternative P sources
in terms of P availability in soils is necessary. The objective of this work is to study the effect
of different wastes derived from the winery and distillery industry on the availability of P
(Olsen-P) in two calcareous soils of Spain, country which holds the largest area under
vines in the world (International Organization of Vine and Wine, 2014).
5.3. MATERIAL AND METHODS.
5.3.1. Organic materials.
We used five organic residues from the winery and distillery industry applied freshly or
after being subjected to a composting stabilization. The raw wastes used were exhausted
grape marc (EGM) and grape stalks (GS). Composts made with different proportions of
these raw materials were also included in the experiment: C1 (85% EGM + 15% GS); C2
(15% EGM + 85% GS). In addition, a compost made with EGM, GS and a little proportion
of lees cake was also included in the experiment: C3 (15% EGM + 75% GS + 10% LC). The
composts used were provided by a commercial composting plant located in
Socuéllamos (Ciudad Real), a grape-growing area in central Spain. The mixtures were
composted in open-air piles during 2-3 months subjected to pile-turning aeration, being
the temperature and moisture periodically monitored during all the process. The stalks
P availability from wine-distillery wastes
78
were previously crushed using a hammer mill in order to reduce the particle size and
composted independently several months before incorporating to the mixture piles, as a
longer time to stabilize them is required due to their lignocellulosic nature.
Table 5.1. Organic materials used in the incubation experiment.
All organic materials were air-dried and ground before using them for incubation
and analysis. EC and pH in the waste samples were determined in a 1:25 (w/v) water
extract; organic carbon and water soluble organic carbon (WSOC) were determined by
dichromate oxidation according to Ciavatta et al. (1989), being WSOC determined in a
1:10 (w/v) water extract, after desiccation. Total, organic and inorganic P were
determined following the Standards, Measurements and Testing (SMT) procedure for
phosphorus fractionation (European Commission) as reviewed by García-Albacete et al.
(2012) for organic wastes. Available P (Olsen P) was determined by extraction with 0.5 M
Waste ID Composition Aspect
EGM Exhausted grape marc
GS Grape stalks
C1
Compost 1
(85% exhausted grape
marc + 15% grape stalk)
C2
Compost 2
(15% exhausted grape
marc + 85% grape stalk)
C3
Compost 3
(15% exhausted grape
marc + 75% grape stalk
+ 15% lees cake)
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79
NaHCO3, pH 8.5, for 30 min. Phosphorus in the extracts was determined colorimetrically
by the method of Murphy and Riley (1962). Total nitrogen was analyzed by the Kjeldahl
method and total Ca and Mg by atomic absorption spectrophotometry after digestion
with nitric and sulfhydric acid.
5.3.2. Incubation experiment.
Soil samples were collected from the surface layers of two arable sites at the
experimental field station `La Entresierra´ in Ciudad Real (S1) and at the experimental
field station of the `Technical University of Madrid´ in Madrid (S2), both located in central
Spain. All the organic materials were applied to 100 g of each soil at a rate equivalent to
100 t ha-1 (fresh weight) mixed thoroughly, and placed in 100 mL polypropylene
containers. An unamended control was also established. The pots were watered to reach
70% of the soil water-holding capacity by adding deionized water and incubated at 27
°C in a controlled incubation chamber for 16 weeks. The moisture was controlled
periodically to replenish water evaporation losses. Destructive samples were collected at
7, 14, 28, 56 and 112 days of incubation (three replicates per treatment at each sampling
date).
Figure 5.1. Incubation chamber used in the experiment.
For all the sampling dates, the moisture content of the soil samples was
determined gravimetrically, after drying to constant weight at 105 °C. Dissolved organic
carbon (DOC) was determined using an automatic analyzer for liquid samples (TOC
3100, Analytik Jena) after a 1:5 soil/water extraction. Soil samples were air-dried and
P availability from wine-distillery wastes
80
analyzed for pH and Olsen P. Soil pH was determined in 1:2.5 soil/water suspensions with
a pH electrode (Micro pH 2001, Crison). Available P was measured by the Olsen method
(Olsen et al., 1954), using a 1:20 extraction with 0.5 M sodium bicarbonate.
Soil inorganic P was fractionated at the end of the incubation following the
scheme proposed by Kuo (1996) for calcareous soils, based on three sequential chemical
extractions: 1) 0.1 M NaOH + 1 M NaCl (NaOH-NaCl-P); 2) 0.3 M sodium citrate + 0.1 M
NaHCO3 and sodium dithionite at 85 °C (CBD-P) ; and 3) 0.5 M HCl (HCl-P). After each
successive sequential extraction, soils were washed twice with a saturated NaCl solution
and the supernatant was included in the corresponding extracts. NaOH-NaCl-P
represents soluble and loosely bound P, which refers mainly to P adsorbed on active
surfaces (Ruíz et al., 1997) that undergoes adsorption and/or precipitation reactions.
CBD-P is the reductant soluble P, related to P occluded within Fe oxides and HCl-P is P
precipitated as poorly soluble Ca-P (mainly lithogenic apatite) (Delgado and Torrent,
2000; Delgado et al., 2000; Ruíz et al., 1997). The P sorption index (PSI) of the soil samples
at the end of the experiment was determined according to Bache and Williams (1971):
one gram of air-dried soil was equilibrated with 20 mL of 75 mg P L-1 solution (1.5 g P kg-1
soil) and PSI was calculated as PSI (mg kg-1)= XV/S (where X is initial menus final solution
P in mg L-1, V is the solution volume in L and S is the soil weight in kg). Phosphorus content
in all the extracts was determined by the method of Murphy and Riley (1962). To evaluate
the potential environmental risk derived from the use of these wastes in calcareous soils,
the degree of P saturation of the soils samples at the end of this study was evaluated by
the ratio between available P to P fixation maximum according to Pautler and Sims
(2000): DPS(%)= Olsen P x100/(Olsen P + PSI).
5.3.3. Statistical analysis.
The effect of soil, waste type and incubation time on variations in pH, DOC and Olsen P
induced by the application of the organic wastes were tested by factorial ANOVA. At
the end of the incubation period, the effect of soil and waste type on variations in
inorganic P fractions, PSI and DPS were evaluated using the same method. One way
ANOVA was carried out to compare the effect of the different treatments on the
inorganic P fractions, PSI and DPS separately in each soil. Significant statistical differences
between the different treatments were established by the Tukey test (p≤0.05). Pearson
correlation coefficients were used to determine relationships between soil available P
and organic wastes properties.
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81
5.4. RESULTS AND DISCUSSION.
5.4.1. Soils and organic wastes characteristics.
Selected chemical and physical properties of the soils of study are shown in Table 5.2.
The soils were calcareous, with a similar pH but having S1 higher CaCO3 content than S2.
Sand was the dominant size fraction in both soils, but they differed in the clay and silt
content, being sandy clay loam and sandy loam respectively. Total Fe content was
higher in S1. In S1, Olsen P was slightly above the critical limit of 10 to 20 mg P kg-1 soil for
common field crops (Delgado and Torrent, 2000), whereas in S2 this value was well above
this limit. The dominant inorganic P fraction in S1 was CBD-P (73%), indicating that most
soil P was occluded within Fe oxides. On the contrary, in S2 the major fraction is HCl-P
(54%), being P in this soil mainly associated to poorly soluble Ca phosphates. The soil P
sorption capacity, as indicated by the PSI, was higher in S1.
The characteristics of the winery and distillery wastes used in the experiment are
shown in Table 5.3. The organic wastes displayed marked variations in pH which ranged
from 5.4 to 9.5. Total P content ranged from 0.7 to 7.6 g kg-1, being higher in the
composted materials. The highest P content observed in C3 was a result of the high
concentration of P in the wine lees with respect to the other wastes generated from this
industry (Bustamante et al., 2008a). The proportion of organic P was higher in the raw
materials, whereas in the composted mixtures P was mainly in inorganic forms. The
organic materials used were also different in terms of organic matter, raw materials
contained more total and water soluble organic C than the composted wastes. The
higher C:N and C:P ratios of the raw materials indicated a lower degree of organic
matter stabilization.
P availability from wine-distillery wastes
82
Table 5.2. Physicochemical properties of the experimental soils.
aCCE: Calcium carbonate equivalent; bACCE: Active calcium carbonate equivalent; cPSI: Phosphorus sorption index; dInorganic P fractions according to Kuo
(1996); eCBD: citrate-bicarbonate-dithionite.
Table 5.3. Chemical properties of the organic materials.
Organic materiala pH CE Organic C WSOCb TNc Ptd Pie Pof Olsen P
C:N C:P Ca Mg
(mS cm-1) (g kg-1)
(g kg-1)
EGM 5.4 0.9 286 11 22 1.0 0.5 0.5 0.4 13 272 0.1 0.3
GS 7.2 0.9 325 9 15 0.7 0.5 0.2 0.3 21 458 0.1 0.6
C1 9.5 0.6 200 6 24 2.8 2.4 0.5 0.3 8 71 1.5 0.7
C2 7.9 0.6 257 3 37 2.9 2.5 0.4 0.2 7 90 0.2 0.7
C3 9.2 1.2 174 4 23 7.6 7.3 0.3 0.3 8 23 0.3 0.6 aEGM: exhausted grape marc; GS: grape stalk; C1: Compost 1 (85% exhausted grape marc+15% grape stalk); C2: Compost 2 (15% exhausted grape marc+85%
grape stalk); C3: Compost 3 (15% exhausted grape marc+ 75% grape stalk + 10% lees cake). bWSOC: Water soluble organic C; cTN: Total N; dPt: Total P; ePi: Inorganic P; fPo: Organic P.
Soil pH EC Texture Organic
C CCEa ACCEb Fe
Exchangeable
Ca+Mg PSIc Olsen P
Inorganic P fractionsd
Clay Silt Sand NaOH-NaCl-P CBD-Pe HCl-P
(mS cm-1) (g kg-1) (g kg-1) (g kg-1) (g kg-1) (g kg-1) (mmolc kg-1) (mg kg-1) (mg kg-1) (mg kg-1)
S1 8.4 0.2 216 80 704 14 77 23 25 60 409 22 32 370 140
S2 8.5 0.1 132 193 674 10 14 5 11 50 167 35 17 623 612
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83
5.4.2. Evolution of pH.
The evolution of soil pH with time is shown in Figure 5.2. The analysis of variance (ANOVA)
for all wastes, soils and times revealed that the variations in soil pH after the incorporation
of winery and distillery wastes were affected by these three factors (Table 5.4).
Figure 5.2. Evolution of pH in soils treated with the different organic materials during the
incubation period. Vertical bars indicate standard error (n=3).
A slight decrease of the soil pH (0.12-0.16 units on average) was observed for soils
amended with those materials containing a higher proportion of exhausted grape marc
and lees cake (EGM, C1 and C3), especially at the beginning of the incubation. This
acidification could be due to the acidic character of EGM (Bustamante et al., 2007),
together to the acid effect of the decomposable products of the organic wastes (Usman
et al., 2004; Achiba et al., 2009; Roig et al., 2012). The evolution of soil pH with time
depend on the soil of study. In S1, all the organic materials produced a slight acidification
during the first weeks of incubation, significant in those soils amended with EGM and C1,
but after 28 days of incubation the pH values of the amended soils were equalized to the
control and tended to stabilize until the end of the experiment, due to the strong buffer
capacity of the calcareous soils, as observed by other authors (Gallardo-Lara and
Nogales, 1987; Bustamante et al., 2007). In S2, this initial decrease was not observed for
P availability from wine-distillery wastes
84
GS and C2, and even there was a clear increase with respect to the other treatments. In
this soil, the pH in all treatments increased to more alkaline values in the second week of
incubation, possibly as a consequence of the organic matter mineralization, reaching
the highest values in the soils amended with GS and C2. The following week, soil pH
decreased to stabilize until the end of the incubation, except in the soil amended with
EGM, in which a decreasing tendency was still observed.
Table 5.4. ANOVA of the changes in pH (ΔpH), dissolved organic C (ΔDOC) and Olsen P
(ΔOlsen P) concentrations in soils treated with the organic materials as affected by
waste, soil type and time.
Source dfa MSb Fc Pd
ΔpH
Soil 1 0.078 17.048 <0.001
Waste 4 0.257 56.557 <0.001
Time 4 0,025 5.554 <0.001
ΔDOC
Soil 1 125479.365 200.302 <0.001
Waste 4 6920.742 11.048 <0.001
Time 4 1536.070 2.452 0.051
ΔOlsen P
Soil 1 1.87069234 0.469 0.495
Waste 4 313.118996 78.583 <0.001
Time 4 160.736144 40.340 <0.001 adf: degrees of freedom; bMS: mean square; cF: variance ratio; dP: probability
5.4.3. Evolution of dissolved organic C.
The variations in DOC concentration with time subsequent to the application of organic
wastes are shown in Figure 5.3. Significant differences in the increment of DOC were
found between wastes and soil types but not along time (Table 5.4). The application of
the organic materials to soil enhanced the concentrations of DOC during all the period
of study, being this increment larger in S2 than in S1 (2.6 and 2.2 fold, respectively). Mean
values for DOC in the unamended soils over the 112 days of incubation were 52 and 76
mg C kg-1 for S1 and S2 respectively. This could indicate a higher C mineralization in S2
possibly due to its lower proportion of clay compared to S1 as various authors have found
that soil clay content protect organic matter from mineralization (Bustamante et al., 2010;
Hernández et al., 2002). In the unamended S1, there was a significant decrease of DOC
in time, while in S2 an opposite tendency was observed.
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85
Dissolved organic C in soil may serve as a ready available source of C and energy
for the microbial biomass (Sánchez-Monedero et al., 2004; Guerrero et al., 2007). After 56
days of incubation, the concentration of DOC in both soils amended with non-stabilized
materials slightly decreased towards the end of the incubation period, whereas in those
soils amended with compost, the concentration of DOC remained constant along the
incubation time or even increase. The decrease of DOC in soils may be attributed to
microbial transformation (immobilization or mineralization) (McDowell, 2003). The
application of organic materials to soil has been reported to increase soil microbial
biomass (Pascual et al., 1997, Sánchez-Monedero et al., 2004) and this enhancement
was found to be higher when less stabilized materials were added, which could explain
the decline of DOC with the application of raw materials to both soils. In addition, the
steady increase of DOC concentration in soil after the application of composted
materials may be due to the incorporation of carbohydrates more resistant to
degradation, as a result of the stabilization of the organic matter during the composting
process (Pascual et al., 1997, 1999).
Figure 5.3. Evolution of dissolved organic C in soils treated with the different organic
materials during the incubation period. Vertical bars indicate standard error (n=3).
P availability from wine-distillery wastes
86
5.4.4. Evolution of Olsen P and recovery of applied P.
Figure 5.4 shows the variations in P Olsen concentration with time subsequent to the
application of the organic materials, which were affected by waste type and time, but
no effect of soil type was found (Table 5.4). Although the average concentration of
available P (Olsen P) in the unamended soils was higher in S2 (36.6 mg kg-1) than in S1
(24.7 mg kg-1), the increment produced as consequence of the organic amendment
incorporation did not differ between soils. The evolution of Olsen P with time was different
between soil types, remaining constant in S1 and increasing along the incubation period
in S2. The higher DOC concentration in S2 could explain the upward trend in the evolution
of Olsen P as a result of mineralization of organic P. Although it is assumed that
mineralization of organic P also occurred in S1 (but in a lesser extent), the available P
released could have been adsorbed by the solid components of the soil, due to the
higher sorption capacity of this soil (Table 5.2).
Figure 5.4. Evolution of Olsen P in soils treated with the different organic materials during
the incubation period. Vertical bars indicate standard error (n=3).
The application of five different winery and distillery wastes, stabilized and non-
stabilized, to the two soils resulted in a moderate but significant increase in available P
especially from the middle of the incubation period, during which the maximum increase
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87
was observed (9.2 mg kg-1 on average), remaining stable until the end of the experiment.
The increase in the concentration of Olsen P in the amended soils from the beginning
until the end of the incubation period in S1 ranged in each treatment as EGM (6.9)>GS
(2.8)>C2 (-1.0)>C1 (-1.1)> C3 (-1.6) and in S2 as EGM (16.6)>C3 (13.2)>GS (10.3)>C1
(9.7)>C2 (7.8). At the end of the incubation period, all the wastes increased significantly
the concentration of Olsen P with respect the unamended soils. However, no differences
were found between the different wastes in S1, despite the different amounts of P
applied with each one, whereas in S2, only the application of C3 increased the
concentration of Olsen P with respect to the other wastes applied.
The recovery of applied P (based on Olsen P) was calculated as (all the units in
mg kg-1):
Recovery % = ((P amended – P control)/Applied P) x 100
As shown in Table 5.5, the recovery of applied P was higher in the non-stabilized
materials, being the mean recovery in the EGM treatment 6.1 and 6.6% for S1 and S2
respectively, and 21.1% for GS in both soils. This fact suggested us that the increase in soil
available P is not only a result of the P input to soil but also of the effect of the organic
matter addition on soil processes. The significant relationship between the concentration
of DOC and Olsen P found in this study (r=0.68, n=180, p<0.001) underlines this fact.
Table 5.5. Recovery of applied P (%) over time in soils treated with the different organic
materials.
Days of incubation
Soil
Organic
material 7 14 28 56 112
S1
EGM -0.66 a -1.13 a 4.65 a 12.22 b 15.32 b
GS 13.10 b 16.40 b 24.25 b 24.71 c 27.08 c
C1 4.08 a 2.55 a 5.97 a 4.24 a 5.32 a
C2 1.77 a 1.34 a 4.43 a 2.96 a 3.07 a
C3 2.27 a 2.20 a 2.46 a 2.46 a 2.62 a
S2
EGM -1.92 a 2.56 a 8.21 a 9.26 b 14.88 bc
GS 6.14 b 22.22 b 32.34 b 28.38 c 16.31 c
C1 1.71 a 4.78 a 5.37 a 3.19 ab 3.85 ab
C2 1.34 a 1.41 a 3.03 a 1.81 a 2.43 a
C3 1.60 a 2.00 a 3.61 a 2.62 ab 3.18 ab
Different letters within each soil and time indicate significant differences between treatments at
p≤0.05.
The increase in soil available P with time in the non-stabilized wastes coincided
with a decrease in DOC concentration, possibly due to mineralization of the organic P
contained in these wastes, being carbon consumed by microorganisms as energy
P availability from wine-distillery wastes
88
source. The application of organic materials to soil has been proven to produce changes
in the enzymatic activity that affects the availability of P (Marinari et al., 2006; Krey et al.,
2013; Requejo and Eichler-Löbermann, 2014). This enhancement of Olsen P could be also
a result of adsorption of DOC to mineral surfaces releasing sorbed P to the soil solution.
Ohno et al. (2005) stated that the application of manure increase available P in soil due
to an enhancement in DOC concentration which inhibit P sorption.
Table 5.6. Correlation coefficients between the increment in Olsen P (ΔOlsen P) and
some selected properties of the organic materials at each incubation interval in
amended soil.
Days of
incubation Pt Pi Po Olsen P WSOC C/P Ca
7 0.72** 0.74** -0.58** -0.19 -0.68** -0.39* -0.10
14 0.46* 0.49** -0.75** 0.00 -0.45* -0.20 -0.32
28 0.49** 0.52** -0.75** -0.02 -0.45* -0.04 -0.28
56 0.29 0.32 -0.68** 0.41* 0.01 0.29 -0.53**
112 0.42* 0.42* -0.30 0.52** -0.05 0.01 -0.54**
*Significant at p≤0.05
**Significant at p≤0.01
Table 5.6 presents the correlation coefficients for the linear relationship between
the increase in P Olsen (Δ) and some properties of the organic wastes that could control
the availability of P in soil. At the beginning of the incubation, total and inorganic P
concentration in the organic wastes were the aspects that influenced more the
availability of P in soil, obtaining positive relationships. On the contrary, organic P applied,
the water soluble organic C (WSOC) content and C:P ratio of the materials, influenced
negatively. One example is the negative recovery of Olsen P observed in the first weeks
after the application of the least stable material (EGM), with the highest WSOC content
and a high C:P ratio (see Table 5.3). This indicates that microbial P immobilization
occurred caused by the addition of soluble organic compounds, fact that was also
observed by López-Martínez et al. (2004) using other organic wastes. The C:P ratio of EGM
is well above the threshold of 100 established by Cheshire and Chapman (1996) for soil P
immobilization. After 56 days of incubation, a negative relationship between the total Ca
content in the different organic materials and ΔOlsen P was observed and remained until
the end of the incubation. This is in concordance with previous studies of other authors
using manures and sludge (Robinson and Sharpley, 1996; Siddique et al., 2000). The
highest input of Ca was provided by C2 (91 mg kg-1 soil) while with the other wastes it
ranged from 6 to 17 mg kg-1. Siddique and Robinson (2003) found that large inputs of Ca
in amended soils regulate the readily available pool of P due to an increase in
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89
exchangeable Ca and the formation of Ca-P precipitates in soil close to the P source.
Another hypothesis is that the higher concentration of Ca in C2 decreased the
availability of P in the waste (C2 had the lowest concentration in Olsen P) due to the
adsorption and precipitation of phosphate, as other authors stated that there is an
inverse relationship between the concentration of Ca, Fe and Al and P solubility (Shober
and Sims, 2007; García-Albacete et al., 2012).
Soil pH is another factor that could control the variations in soil P availability by
governing the sorption processes of P fixing minerals and the solubility of P-minerals
(Chaïrat et al., 2007; Devau et al., 2011). In our study, a negative relationship between
Olsen P and soil pH was found (r=-0.301; n=180; p<0.001) which reflects that soil
acidification induced by some of these wastes may contributed to increase the amount
of available P in soil.
The increase in soil available P after amendment with winery and distillery wastes
was quite slight if we compare our results with those obtained in other studies using
different by-products as manures or sludges, which means a lower risk of water P
contamination. For example, Adeli et al. (2005) reported an increase in soil available P
content of 133 mg kg-1 applying a lower dose of poultry litter to a calcareous soil. In the
same line, Almendro-Candel et al. (2003) reported increments of available P up to 75 mg
kg-1 at different depths in a calcareous soil amended with sewage sludge at a rate of 90
t ha-1. Our results are more concordant with studies using plant residues (Sharpley and
Smith, 1989; Garg and Bahl, 2008) or composts made with other vegetable and industrial
wastes (Uygur and Karabatak, 2009; Malik et al., 2013) probably as a consequence of
the lower P content and higher C:P ratios of these materials.
5.4.5. Changes in soil P status: inorganic P fractions, P sorption capacity and
degree of P saturation.
Figure 5.5 shows the different inorganic P fractions in the soils after 112 days of incubation.
The NaOH-NaCl-P, the most labile fraction of the sequential extraction, represented the
minor fraction in both soils, being this proportion higher in S1 (11% of total recovered
fractions) than in S2 (1%). The incorporation of the organic residues increased significantly
the concentration of this fraction in soil except for C2 and C3 in S1 and only for C3 in S2.
Similarly, an increase of NaOH-NaCl-P after the application of other organic wastes to
calcareous soils was found by other authors (Halajnia et al., 2009; Uygur and Karabatak,
2009). The highest increases of NaOH-NaCl-P were observed in those wastes containing
a major proportion of EGM, being this fraction 13 mg kg-1 greater than the control in S1
(C1) and 10 mg kg-1 in S2 (EGM). It is possible that the slight acidification induced by the
P availability from wine-distillery wastes
90
application of wastes containing larger amounts of EGM have influenced the higher
recovery of this fraction. Adhami et al. (2013) showed that NaOH-NaCl-P was positively
correlated with oxalate extractable Fe, which could explain the higher concentration of
this fraction in S1, the soil with a higher Fe content.
Figure 5.5. Inorganic P fractions given by the sequential fractionation scheme of Kuo
(1996) in soils treated with the different organic materials at the end of the incubation
period. Different letters within each soil columns indicate significant differences
between treatments (p≤0.05). The bars indicate standard error (n=3).
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91
Citrate-bicarbonate-dithionite-P was more abundant in the soils of study, being
the majoritary fraction in S1 (73%) and the intermediate fraction in S2 (41%). This fraction
is also related to metastable phases (i.e. dicalcium, octacalcium and tricalcium
phosphate) and possible to a small fraction of the stable phase (hydroxyapatite) which
could be extracted with citrate bicarbonate (Delgado et al., 2000). The CBD-P only
increased with the highest application of P (C3) in both soils, by 111 mg kg-1 (28%) and
139 mg kg-1 (30%) respectively. This fact may reflect the formation of organic-Fe
compounds that could have removed P from solution by formation of a cationic bridge
between organic C and P (Leytem and Westermann, 2003). This observation also would
explain the tendency to decrease NaOH-NaCl-P fraction in this treatment in S1 and was
also observed by Halajnia et al. (2009) with manure application together with inorganic
P.
Table 5.7. ANOVA of the changes in the inorganic P fractions (ΔNaOH-NaCl, ΔCBD-P,
ΔHCl-P), P sorption index (ΔPSI) and degree of P saturation (ΔDPS) in soils treated with
the organic materials as affected by waste and soil type at the end of the incubation
period.
adf: degrees of freedom; bMS: mean square; cF: variance ratio; dP: probability
Regarding HCl-P, this pool accounted for 16% of total extracted P in S1 and was
the major fraction in S2 (58%). A slight decrease of this fraction was observed for the non-
stabilized materials whereas the highest application of P (C3) caused a clear
enhancement of this fraction in both soils, especially in S1 in which the increment was
278% of P recovered in the control. This is in concordance with other studies stating that
precipitation of poorly soluble Ca-P was more marked at higher P applied (Delgado et
al. 2002) and with the application of more stabilized materials (Uygur and Karabatak,
Source dfa MSb Fc Pd
ΔNaOH-NaCl-P
Soil 1 38.669 11.745 0.003
Waste 4 249.509 75.784 <0.001
ΔCBD-P
Soil 1 8077.002 1.002 0.329
Waste 4 23548.011 2.923 0.047
ΔHCl-P
Soil 1 24655.053 67.567 <0.001
Waste 4 28422.961 77.892 <0.001
ΔPSI
Soil 1 14933.422 50.287 <0.001
Waste 4 563.915 1.899 0.150
ΔDPS
Soil 1 1302.843 11.952 0.002
Waste 4 5.795 0.053 0.994
P availability from wine-distillery wastes
92
2009). The reduction of the HCl-P fraction in the soils amended with non-stabilized
materials with respect to the unamended soil, may show that the organic acids released
during the decomposition of these materials have been enhanced the release of
sparingly soluble P from the more stable P fraction (Von Wandruszka, 2006).
The variations of each inorganic P fraction as consequence of the adittion of
winery and distillery wastes were influenced by the waste type (Table 5.7). However, the
effect of soil was observed only for NaOH-NaCl-P and HCl-P, but not for CBD-P.
The soil phosphorus sorption index (PSI) and degree of P saturation (DPS) in the
different treatments after 112 days of incubation is shown in Table 5.8. The effect of the
addition of winery and distillery wastes on soil P sorption was different between the two
types of soil studied. In S1, there was a general tendency to decrease PSI values by 7%
(C3) to 21% (EGM) whereas in S2 no significant changes were induced by the application
of organic wastes. This could be probably due to the higher saturation of P in S2 with
respect to S1, as indicated by the initial total inorganic P content (1252 and 533 mg kg -1
respectively) and PSI (409 and 167 mg kg-1), which has been found to be negatively
correlated with the decrease in P sorption (Siddique and Robinson, 2003). The factorial
ANOVA also indicated that changes in soil P sorption induced by the application of
organic wastes were only affected by the soil type (Table 5.7).
Table 5.8. Olsen P, P sorption index (PSI) and degree of P saturation (DPS) in soils treated
with the different organic materials at the end of the incubation period.
Mean ± standard error. Different letters within each soil indicate significant differences between
treatments at p≤0.05.
It is well known that the humic substances and organic acids released from the
decomposition of organic wastes can be adsorbed into soil surfaces and modify the P
sorption capacity of soils (Mokolobate and Haynes, 2002; Mkhabela and Warman, 2005;
Fuentes et al., 2008). In S1, the decrease in PSI in the EGM and GS treatments coincided
Soil Organic material Olsen P PSI DPS
(mg kg-1) %
S1
Control 23.2 ± 1.2 a 464.1 ± 10.4 b 4.8 ± 0.3 a
EGM 32.9 ± 1.8 b 365.7 ± 18.2 a 8.3 ± 0.6 b
GS 34.8 ± 1.8 b 422.0 ± 30.4 ab 7.7 ± 0.9 b
C1 32.4 ± 0.6 b 380.5 ± 8.9 ab 7.8 ± 0.1 b
C2 28.4 ± 0.3 ab 398.9 ± 5.8 ab 6.6 ± 0.0 ab
C3 35.2 ± 2.0 b 431.1 ± 21.0 ab 7.6 ± 0.7 b
S2
Control 42.2 ± 0.4 a 159.5 ± 18.9 a 21.3 ± 2.0 a
EGM 51.6 ± 0.7 bc 182.4 ± 3.7 a 22.0 ± 0.5 a
GS 49.1 ± 2.3 b 180.5 ± 14.9 a 21.6 ± 1.8 a
C1 48.9 ± 1.3 b 149.5 ± 11.2 a 24.8 ± 1.6 a
C2 46.3 ± 1.3 ab 154.1± 19.6 a 23.6 ± 2.7 a
C3 56.7 ± 0.5 c 163.4 ± 10.4 a 25.9 ± 1.1 a
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93
with a slight reduction of the HCl-P fraction with respect to the unamended soil (14 and
12% respectively). On the contrary, precipitation of Ca-P was more marked when
applying compost so that the reduction in PSI in this case might be due to complexation
of the P released from these wastes (Reddy et al., 1980; Iyamuremye et al., 1996b).
Additionally, the phenolic compounds contained in the winery and distillery wastes
(Bustamante et al., 2008a) have been also reported to prevent P adsorption in
calcareous soils (Hu et al., 2005a, 2005b). As the stabilization of organic wastes by
composting reduce the concentration of these compounds (Galli et al., 1997; Nogales
et al., 2005), this explains why this effect was not observed in soils amended with compost.
As shown in Table 5.8, DPS in the soils of study increased with the application of
organic materials, however, this increase was statistically significant only in S1. Changes
in the DPS induced by the application of the winery and distillery wastes were influenced
by the soil type but no effect was observed regarding the waste type (Table 5.7). This
index has been used as a P loss risk indicator because it has a strong relationship with P
concentration in runoff or leaching (Sims et al., 2002; Nelson et al., 2005; Casson et al.,
2006). Soil DPS values up to 25-40% are generally associated with a greater risk of P losses
(Pautler and Sims, 2000). Soil 1, which showed the highest contents of clay, Fe and CaCO3
(S1) exhibited the greatest values of PSI and consequently, the lowest DPS. On the
contrary, the higher values of DPS of S2 were very close to the threshold proposed by
these authors, therefore (although in this case no significant differences were observed)
the surplus of P inputs in this soil could represent an eutrophication hazard. Casson et al.
(2006), using the same indicator, established a DPS threshold of 27 or 44% by measuring
water extractable soil P or CaCl2-P respectively, in five manure-amended soils with
different characteristics.
5.5. CONCLUSIONS.
Available P in both soils increased due to the addition of winery and distillery wastes.
Although in the first weeks changes in soil Olsen P were mainly related to the P input, the
recovery of Olsen P along time suggested us that the incorporation of less stabilized
organic matter with the raw materials (with lower P) also influenced P availability by
enhancing the mineralization of organic P contained in these wastes or by decreasing
soil P sorption. The application of higher amounts of P with the composts (C1, C2 and C3)
resulted on the formation of Ca phosphates increasing the recovery of P in the HCl-P
fraction, due to the calcareous nature of these soils. On the contrary, the addition of
EGM and GS produced a slight release of P from the HCl-P pool which was turned into
P availability from wine-distillery wastes
94
more labile forms (NaOH-NaCl-P). This was consistent to a decrease of the soil P sorption
and an increase of the degree of P saturation that only was appreciable in the less P
saturated soil.
Acknowledgments
This work was funded by the Instituto Nacional de Investigación y Tecnología Agraria y
Alimentaria, Spanish Agency (RTA2010-00110-C03). The authors would like to thank
Agricompost S.L. for providing the organic materials used in this study.
CHAPTER 6:
ORGANIC AND INORGANIC PHOSPHORUS FORMS IN
SOIL AS AFFECTED BY LONG-TERM APPLICATION OF
ORGANIC AMENDMENTS
Requejo, M.I., Eichler-Löbermann B. (2014). Nutrient Cycling in Agroecosystems 100(2): 245-255.
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97
6.1. ABSTRACT.
Organic amendments contribute significantly to the phosphorus (P) supply in
agroecosystems. However, their long-term effects on specific P forms in soils are not
completely understood. The objective of this study was to investigate the concentration
of organic P forms and inorganic P pools in soil and the activity of enzymes involved in
the P turnover in a long-term field experiment running since 1998 in Northern Germany as
affected by P amendments. The following treatments with different P supplies were
sampled in 2012, 14 years after the establishment of the experiment: control (no P), cattle
manure (manure), biowaste compost (compost), and biowaste compost in combination
with triple-superphosphate (compost + TSP). The classification of organic P forms by using
enzyme additions to NaOH–EDTA soil extracts showed non-hydrolyzable organic P as the
dominant form in soil followed by inositol hexakisphosphate (Ins6P)-like P. Non-
hydrolyzable and total organic P concentrations in soil were highest in the combined
compost + TSP treatment, which received the highest amount of inorganic P. The values
of the bioavailable P pools (water-extractable P and double lactate-extractable P) were
in accordance with the P balance (P addition with the amendments minus P removal
with harvested crops) independently of the type of amendment. The results of this
research suggest that the distribution of soil P forms is more reliant on the turnover
processes in the soil than on the forms of P added.
Organic and inorganic P forms affected by organic amendments
98
6.2. INTRODUCTION.
Studies predict that resources of phosphate rock will be depleted within a short to
medium duration of years (Rosmarin, 2004; Cordell et al., 2009; Van Vuuren et al., 2010).
Therefore, recycling organic residues as compost and manure in agriculture is an
important practice to replenish soil phosphorus (P) pools, replacing inputs of mineral
fertilizer and to improve soil properties (Eichler-Löbermann et al., 2008; Gopinath et al.,
2008).
Organic amendments are composed of organic and inorganic P forms, but
inorganic P is usually the principal P form. In a wide variety of organic wastes of different
origins and treatments (including compost, pig slurries, sludge and digestate) organic P
forms only account between 1 and 30% of total P (Sinaj et al., 2002; García-Albacete et
al., 2012). However, studies using manure from different animal species showed that
organic P can amount to 80% of total P (Pagliari and Laboski, 2012; Pagliari and Laboski,
2013). To improve the nutrient utilization in cropping systems and thus avoiding P losses a
comprehensive understanding of the influences of organic amendments on soil fertility
and soil P forms is necessary.
It is well documented that increase in organic matter enhances microbial
biomass and microbial activity of soils (Ehlers et al., 2010; Krey et al., 2013). Thus, farming
systems with higher inputs of organic amendments result in larger microbiological activity
compared to soils receiving mainly mineral fertilizers. Higher microbial activities however,
also represent higher P turnover in soil, by microbial immobilization of inorganic P,
mineralization of organic P and microbial P synthesis (Richardson, 2001; Richardson and
Simpson, 2011). The activity of microorganisms in soil can be estimated by the activity of
intracelular enzymes, e.g. dehydrogenase (Nain et al., 2009). The hydrolysis of organic P
like esters and anhydrides of phosphoric acid is catalyzed by phosphatase in soil (Eivazi
and Tabatabai, 1977). Alkaline (AlP) and acid phosphatase (AcP) can be excreted by
both, crops and microorganisms, whereas higher crops almost exclusively excrete AcP
(Dick et al., 2000).
In the past, fertilizer recommendations were given with especial attention on soil
inorganic P pools (Steffens et al., 2010). However, organic P constitutes an important
proportion of total P in soil (30–65%; Harrison, 1987) that can be available after being
hydrolyzed to inorganic forms. The soil organic P fraction includes different chemical
forms which differ in their susceptibility to degradation by soil enzymes (Bowman and
Cole, 1978), and thus in their potential availability for plants. Several investigations have
studied the hydrolyzabiliy of organic P forms by adding combinations of phosphatase
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99
enzymes with different substrate specificities to soil extracts (He et al., 2004a; Turner et al.,
2003b), soil water suspensions (Annaheim et al., 2013), and manures and composts (He
et al., 2004b; Pagliari and Laboski, 2012; Annaheim, 2013). This technique requires
relatively little laboratory equipment and low costs compared to other techniques like
31P-NMR spectroscopy. Comparison of P classes obtained by enzymatic hydrolysis and
31P nuclear magnetic resonance (NMR) spectroscopy in soil and manure extracts
revealed similar results for both methods (He et al., 2007; Johnson and Hill, 2010).
The effect of different long-term farming systems (conventional, biodynamic and
bioorganic) on organic P forms was investigated by Keller et al. (2012) using incubations
of alkaline soil extracts with acid phosphatase from potato (Solanum tuberosum L.),
nuclease from Penicillium citrinum and phytase from Peniophora lycii. The authors
concluded that the different farming systems did not affect the organic P fractions. In a
field experiment in Switzerland Annaheim (2013) also found that forms of organic P
remained largely unchanged in soils with contrasting fertilizer histories, including a mineral
fertilizer (NPK) and different organic amendments (sludge, compost and manure). Our
study contained a control without P supply, cattle manure, biowaste compost as well as
a combined compost + TSP treatment. The objective of this study was to investigate the
concentration of organic and inorganic P pools in soil, and to quantify the activity of
enzymes involved in the P turnover after 14 years of treatment application under field
conditions in Northern Germany. To estimate the enzymatic hydrolyzability of organic P,
phosphatase enzymes with different substrate specificities were added to soil extracts.
6.3. MATERIAL AND METHODS.
6.3.1. Site characterization.
Soil samples were taken in 2012 at the long-term field experiment established in autumn
1998 to study the effect of different organic amendments and inorganic fertilizers in single
application and combination (see also Eichler-Löbermann et al., 2007). The site is located
at the experimental station of Rostock University in Northern Germany in a maritime
influenced area about 15 km south of the Baltic Sea (54°3´41.47´´ N; 12°5´5.59´´ E). The
mean annual precipitation is about 600 mm and the average annual temperature is 8.1
°C. The soil texture is loamy sand and according to the World Reference Base for Soil
Resources, the soil is classified as Stagnic Cambisol. The bulk density (0–30 cm) was 1.38
kg L-1, and the soil organic matter (SOM) concentration was 2.50%. The initial Pdl
concentration of 42.2 mg P per kg soil indicated a suboptimal P supply according to the
Organic and inorganic P forms affected by organic amendments
100
German soil-P classification (Kerschberger et al., 1997). Soil characteristics at the
beginning of the experiment are given in Table 6.2.
Although the experiment originally had 9 treatments, this study focused on 4
treatments, which were: control (no P), cattle manure (manure), biowaste compost
(compost), and biowaste compost in combination with mineral fertilizer (Triple-
superphosphate TSP) (compost + TSP). Each treatment had 5 replications. The plots had
a size of 30 m2 and were arranged in a randomized block design. Cattle manure and
biowaste compost were applied every three years beginning in September 1998 at a
rate of about 30 t ha-1. Consequently, the last application before sampling in 2012 was in
2010. The compost was produced in a compost facility near Rostock and provided as
mature and sanitized compost on the basis of green garden and landscape residues.
Mineral fertilizer was supplied annually as TSP at a rate of 21.8 kg P ha-1. The control did
not receive any P since 1998. The plots of this study were cropped to spring oilseed rape
(Brassica napus) 1999, spring wheat (Triticum aestivum) 2000, spring barley (Hordeum
vulgare L.) 2001, spring oilseed rape 2002, Winter wheat 2003, winter barley 2004, winter
oilseed rape 2005, maize (Zea mais) 2006, maize 2007, maize 2008, green winter rye
(Secale cereale) (as green manure) and sorghum (Sorghum bicolor) 2009, sorghum 2010,
sunflower (Helianthus annuus) 2011, winter rye 2012. The P supply with the treatments and
the P uptake of the crops resulted in different P balances (see Table 6.1). For oilseed rape
and the cereals (1999–2005, 2012) only the grains of the plants were harvested and the
straw remained on the field. For maize, sorghum and sunflower the whole crops were
harvested and the plants were cut 10 cm above the soil. Plant residues were mulched
into the soil after crop harvests. Afterwards the soil was ploughed at a depth of 25 cm
either in autumn or in spring. The P removal was determined by multiplying the dry weight
of harvested biomass with its P concentration. The P balances for the treatments were
calculated by subtracting the amount P applied minus the P removed with the harvested
crop parts (see Table 6.1).
Table 6.1. Total P supply, total P uptake and total P balance accumulated after 14 years
in field experiment.
Treatment Total P supply within
14 years
P removal* within 14
years
P balance
within 14 years
(kg ha-1)
Control, no P - 345 -345
Cattle manure 311 373 -62
Biowaste compost 350 374 -24
Compost + TSP 655 397 +258
Compost+ TSP= Biowaste compost + Triple-superphosphate * The P removal was estimated by
multiplying the dry weight of harvested biomass with its P concentration.
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101
6.3.2. Soil analysis.
Six soil cores per plot were collected from the upper soil layer (0–30 cm) in September
2012. Samples were air-dried, sieved to 2 mm and stored at room temperature until
analysis. Samples of the organic amendments of the last application (2010) before
sampling were kept frozen at -18 °C until analysis. Various methods were used to
characterize the different soil P pools. Water-extractable P (Pw) was measured
according to Van der Paauw (1971) by extracting 12 g soil in 200 mL of ultrapure water.
Phosphorus concentration in the water extracts was determined by the
phosphomolybdate blue method via flow-injection analysis. Double lactate-extractable
P (Pdl) was determined by extracting P from 12 g soil with 150 mL double lactate solution
(C6H10CaO6.H2O + 10 N HCl) using the method of Hoffmann (1991). Phosphorus
concentration in the dl extract was determined using the vanadate-molybdate method
with a spectral photometer. Oxalate soluble P, Al and Fe (POX, AlOX, FeOX) were extracted
according to Schwertmann (1964) by shaking 2 g of soil with 100 mL of acid oxalate in
the dark for 1 h and measured using inductively coupled plasma spectroscopy (ICP-OES,
JY-238). The P sorption capacity (PSC (mmol kg-1) = (AlOX + FeOX)/2) and the degree of P
saturation (DPS (%) = POX/PSC x 100) were calculated using the oxalate extractable P, Al
and Fe according to Schoumans (2000). Total P(Pt) was extracted using the aqua regia
method, where 6 mL HCl and 2 mL HNO3 are added to 0.5 g soil and digested in a
microwave oven (Mars Xpress, CEM GmbH, Kamp-Lintfort, Germany) followed by ICP
spectroscopy. SOM content was determined by dry ashing.
Dehydrogenase (DHA) activity was determined according to Thalmann (1968) by
suspending 1 g soil in 0.8% triphenyltetrazoliumchloride solution followed by incubation at
37 °C for 24 h. Most microorganisms are able to reduce triphenyltetrazoliumchloride to
triphenylformazan (TPF). The released TPF was extracted with acetone and measured
photometrically at 546 nm. Results were expressed as µg TPF per g soil released within 1 h
(µg TPF g-1 h-1). The activities of acid and alkaline phosphatases (phosphoric monoester
hydrolases, AcP and AlP) were measured according to Tabatabai and Bremner (1969).
The activities were expressed as µg p-nitrophenol released from a pregiven p-nitrophenyl
phosphate solution in 1 g soil after incubation at 37 °C for 1 h (µg p-nitrophenol g-1 h-1).
6.3.3. NaOH–EDTA extraction and enzymatic incubation assay.
Ten grams of fresh compost and manure were added to 100 mL of 0.25 M NaOH–0.05 M
Na2EDTA for 16 h (Cade-Menun and Preston, 1996) centrifuged and filtered (Whatman
4). The extracts were diluted 200 times before colorimetric measurement of inorganic P
with malachite green at 623 nm (Ohno and Zibilske, 1991). For total P, a persulfate
Organic and inorganic P forms affected by organic amendments
102
digestion (Tiessen and Moir, 1993) was performed with 1 mL of NaOH–EDTA extracts and
10 mL of digestion mix (6 g NH4- persulfate in 100 mL 0.9 M H2SO4). In the neutralized digests
total P was determined by colorimetric measurement. Organic P was calculated as the
difference between total P and inorganic P. Extractions were repeated 3 times.
The extraction and analysis of soil samples followed the same procedure. Ten
grams of dry soil were treated for 16 h with 0.25 M NaOH–0.05 M Na2- EDTA in a soil-to-
solution ratio of 1:10 (w:v). For inorganic P, the extracts were diluted 10 times for
colorimetric measurement. Total extractable P was determined as described for the
organic amendments above and organic P was calculated by the difference between
total P and inorganic P.
The enzymatic hydrolyzability of organic P in the NaOH–EDTA extracts from soil,
compost and manure was determined following the protocol of Keller et al. (2012) with
minor modifications. Three enzyme solutions were prepared: acid phosphatase from
potato (Sigma P1146; 50 UN diluted in 15 mL H2O); nuclease from Penicillium citrinum
(Sigma N8630; 0.167 mg dissolved in 1 mL H2O); and a comercial granulated phytase
prepared from Peniophora lycii (RONOZYME; 0.25 g dissolved in 50 mL H2O, centrifuged,
and the supernatant filtered at 0.2 µm). Aliquots of the enzyme solutions (0.5 mL) were
dispensed and stored at -18 °C in a freezer until use. The incubations were conducted in
micro-centrifuge vials achieving a final volume of 320 µL. 40 µL of NaOH–EDTA extract
were added to 60 µL of 1 M MES buffer (2-(N-morpholino) ethanesulfonic acid; at pH 5.2)
and incubated with 20 µL of acid phosphatase alone or in combination with 20 µL of
nuclease solution, or with 40 µL of phytase alone. The final volume was reached by
adding H2O. The incubations were performed at 30 °C for 24 h (Annaheim et al., 2013).
After incubation, inorganic P in the incubation mix was determined using the malachite
green method with a spectral photometer.
Figure 6.1. Detail of the enzymatic incubation assay and the incubator used.
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103
Acid phosphatase hydrolyzes simple phosphomonoesters and
phosphoanhydrides (He et al., 2004a), while phytase additionally hydrolyzes inositol
hexakisphosphate (Ins6P) and RNA (Annaheim et al., 2013). Nuclease is used to achieve
the hydrolysis of diester bonds in nucleic acids, whose organic P is then hydrolyzed by
acid phosphatase (He et al., 2004a). Three model substrates were used to verify the
effectiveness of the three enzyme preparations in the buffered solutions. Acid
phosphatase was shown to be active against glycerolphosphate, with about 90% of
hydrolized P. The combination of acid phosphatase and nuclease hydrolized 74% of P
from DNA and the phytase preparation hydrolyzed 83% of Ins6P. Vials containing only soil
extract or only enzyme in the buffered solution were run simultaneously with the samples
along with an internal calibration curve (0–1.25 µg orthophosphate). No inorganic P
release was detected in the vials incubated with the enzyme preparations alone. Enzyme
labile P was calculated as the difference of inorganic P measured in the soil extracts
incubated with the enzymes and the P measured in the vials without enzyme (4 analytical
replications each). An orthophosphate spike (0.4 µg P) was added to the soil extracts (4
analytical replicates) in parallel with the samples to check the recovery of hydrolyzed P.
Enzyme labile-P was classified into simple monoester-like P (P released by acid
phosphatase); DNA-like P (difference of P released by acid phosphatase in combination
with nuclease and acid phosphatase alone); and Ins6P-like P (difference of P released
by phytase and by acid phosphatase). Non-hydrolyzed P was calculated as the
difference between organic P and the sum of the three fractions of enzyme-labile P.
6.3.4. Statistical analyses.
Analyses of variance were performed using the General Linear Models (GLM) of PASW
Statistics 18 software (SPSS Statistics). For all statistical analyses, a p value of 0.05 was used
to determine significance. In case of significant effects, Duncan multiple range test was
performed to compare means of soil parameters. Descriptive statistics of soil and organic
amendments parameters were calculated using the same software.
Organic and inorganic P forms affected by organic amendments
104
6.4. RESULTS AND DISCUSSION.
6.4.1. Effect of long-term P amendment practice on inorganic and total soil P
pools.
The treatments applied in combination with crop P uptake affected soil bio-available P
contents (Pw and Pdl, see Table 6.2). In relation to the initial Pdl concentration of about
42 mg per kg soil in 1998 the values dropped significantly in the treatment without any P
supply and the manure treatment, remained almost stable in the compost treatment and
increased after the combined compost + TSP application. Organic residues usually
contain considerable amounts of soluble inorganic P which contribute to the fast release
of P after incorporation into soil. The control treatment without any P supply during the
last 14 years showed very low P concentration; Pdl concentration of only 29.7 mg per kg
soil correspond to strong suboptimal P status according to the German soil P classification
(Kerschberger et al., 1997). The clear responses of Pdl values to the P application
(correlation coefficient r = 0.97, p<0.01) make these parameters a suitable tool for fertilizer
recommendations. High correlations between P applied and Pdl and Pw concentrations
in the soil were also found previously in 2004 (correlation coefficient r = 0.73 and 0.77,
p<0.01), where more than 50% of the variation of Pw and Pdl values could be explained
by the P supply (Eichler-Löbermann et al., 2007). No differences were found between
treatments regarding the P-sorption capacity (PSC) in 2012 (Table 6.2). However, the
compost + TSP treatment resulted in the highest degree of P saturation (DPS) with values
of more than 50%. DPS values higher than 40% are critical with respect to P losses from
sandy soils (Schoumans, 2000).
The differences of the Pt contents in soil from the beginning of the experiment and
after 14 years (Δ Pt, kg ha-1) were positively correlated with the P balances given in Table
6.1 (r = 0.75**, n = 20). However, Δ Pt was not identical with the P balances. In the
treatments which have received P the Pt contents in 0 to 30 cm were lower than
expected. The comparison of Δ Pt and the P balances (kg ha-1) showed for the manure
treatment: -112 versus -62, for compost treatment: -66 versus -24, and for the compost +
TSP treatment: +50 vs. +258. This could be explained by a downward movement of P into
deeper soil layers, which was also described by Oehl et al. (2002). Whereas for the
manure and compost treatment the differences between Δ Pt and the P balances were
moderate, much higher Pt values were expected in the compost + TSP treatment and
about 200 kg ha-1 is mathematically missing in the topsoil. This result however, confirms
other studies which showed that P movement with percolating water or in preferential
flow paths increased with higher fertilizer inputs and following higher bioavailable P
contents in the topsoil (Godlinski et al., 2004). On the other hand, in the control the Pt
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105
contents in 0–30 cm were higher than expected (Δ Pt vs. P balances (kg ha-1): -232 versus
-345). This can be explained by a upwards movement of P from the subsoil to the topsoil
under P deficient conditions which was also shown by previous long-term studies by Oehl
et al. (2002) and Gransee and Merbach (2000).
Organic and inorganic P forms affected by organic amendments
106
Table 6.2. Properties of the topsoil collected in 2012 after 14 years of different P amendment management and at the beginning of the
experiment in 1998.
Treatment pH Pdl Pw Pt Pox PSC DPS SOM
(mg kg-1) (mg kg-1) (mg kg-1) (mmol kg-1) (mmol kg-1) (%) (%)
Control, no P 5.30 ± 0.18a 29.7 ± 3.3a 7.4 ± 1.6a 426 ± 38a 9.8 ± 1.7a 22.3 ± 3.9a 43.9 ± 3.2a 2.27 ± 0.14a
Cattle manure 5.44 ± 0.15b 36.1 ± 2.2b 10.8 ± 1.3b 455 ± 24ab 12.4 ± 0.5bc 26.9± 2.6a 46.2 ± 3.2ab 2.60 ± 0.16b
Biowaste compost 5.53 ± 0.18c 44.4 ± 3.7c 11.8 ± 2.0b 466 ± 42ab 11.3 ± 1.8ab 23.8 ± 3.9a 47.6 ± 2.1b 2.63 ± 0.11b
Compost + TSP 5.52 ± 0.13bc 53.8 ± 5.6d 15.6 ± 1.7c 494 ± 43b 13.6 ± 0.6c 26.3 ± 1.3a 51.7 ± 1.6c 2.64 ± 0.14b
p value <0.001 <0.001 <0.001 0.015 0.005 0.161 0.004 0.002
Beginning (1998) 5.75 42.2 - 482 12.4 27.9 44.4 2.50
Mean ± standard deviation of five field replicates. Letters in the same column indicate significant differences between treatments (p<0.05); Pdl double lactate-
extractable P; Pw water extractable P; Pox oxalate-extractable P; PSC P-sorption capacity; DPS degree of P saturation. Compost + TSP= Biowaste compost + Triple-
superphosphate.
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107
6.4.2. Effect of long-term P amendment practice on activities of enzymes.
The long-term P amendment practice also affected the activity of enzymes involved in
the soil nutrient turnover. In the amended plots (which also showed higher SOM contents)
we found higher microbial activities (DHA) in comparison to the control. The availability
of organic substrates usually promotes the growth of microbial populations and microbial
P (Bünemann et al., 2006; Fliessbach et al., 2007). Furthermore, the application of
compost alone or in combination with TSP enhanced the activity of AlP (about 70 µg g-1
h-1, p-np) in comparison to the control (about 40 µg g-1 h-1, p-np). This may be explained
with higher microbial activity (since the synthesis and excretion of AlP is coupled to
microbial activity or population size) and the slight increase of soil pH (from 5.30 in the
control to 5.44 in the manure plots and 5.53 in the compost plots). However, the
application of compost or manure had no significant effect on the activity of AcP. This
was also observed after 10 years experimental time in this field experiment (Krey et al.,
2013) and confirms findings from another field experiment where the long-term
application of manure also enhanced the activity of AlP, but not of AcP in the soil
(Parham et al., 2002). Phosphatases contribute to mineralization of hydrolyzable organic
P in soil, whereas considerably parts of total organic P may be non-hydrolyzable, as
shown in this study (see chapter `Effect of long-term P amendment practice on inorganic
and total soil P pools’) (Table 6.3).
Table 6.3. Soil enzyme activities in soil samples collected in autumn 2012 after 14 years
of different P amendment management.
Mean ± standard deviation of five field replicates. Letters in the same column indicate significant
differences between treatments (p<0.05); AcP, acid phosphatase activity; AlP, alkaline
phosphatase activity; DHA, dehydrogenase activity; p-np, p-nitrophenol; TPF, triphenylformazan;
Compost + TSP= Biowaste compost+Triple-superphosphate.
AcP AlP DHA
Treatment p-np p-np TPF
(µg g-1 h-1)
Control, no P 121± 16 a 39.9 ± 18.9 a 1.08 ± 0.19 a
Cattle manure 133 ± 13 a 58.3 ± 23.4 ab 1.66 ± 0.36 b
Biowaste compost 134 ± 14 a 72.2 ± 17.6 b 1.51 ± 0.18 b
Compost + TSP 130 ± 14 a 69.9 ± 16.3 b 1.48 ± 0.22 b
p value 0.519 0.032 0.032
Organic and inorganic P forms affected by organic amendments
108
6.4.3. NaOH–EDTA extractable P forms in manure and compost.
Inorganic P was the major fraction of P present in the NaOH–EDTA extracts of both
amendments (Table 6.4). Organic P accounted only for 12 and 18% of total P extracted
with NaOH–EDTA for compost and manure respectively. These proportions of organic P
are within the range proposed by Sinaj et al. (2002) for compost (12–33%) and by Pagliari
and Laboski (2012) for manure (12–24%). The concentration of both inorganic and
organic P in NaOH–EDTA extracts was higher in manure than in compost. All organic P in
the NaOH–EDTA extracts compost was enzymatically hydrolyzable, with DNA-like P being
the main form detected (9%). The concentration of Ins6P-like P was lower (7%) and
monoester-like P was not detected. Annaheim (2013) also reported a complete hydrolysis
of all organic P in compost obtained from green waste, whereas the concentration of
Ins6P-like P was 10%, of DNA-like P 6% and of monoester-like P 3%. In the manure extract,
60% of organic P was hydrolyzed by enzyme addition. Monoester-like P was the major
form of the total NaOH–EDTA extracted P (10.7%), the DNA-like P fraction was only 0.2%
and Ins6Plike P was not detected. Annaheim (2013) also obtained a dominance of the
monoester-like P fraction when analysing manure (13.3% monoester-like P). The high
proportion of monoester-like P was also shown for dairy manure with no bedding in water
(He et al., 2006) and NaOH–EDTA (He et al., 2007) extracts. However, He et al. (2007)
reported higher proportions of hydrolyzable organic P and found higher proportions of
DNA-like P and Ins6P-like P (6.4 and 10.1% of total NaOH–EDTA extracted P) as compared
to our results. These differences could be due to the presence of straw in the manure
used in our study. Furthermore, it is important to point out that the composition of manures
and composts may vary from one sampling date to another even if the material
originates from the same facility (Schoumans et al., 2010).
For this study the characterization of the organic P forms in compost and manure
was performed using the amendments applied in 2010. Although the main nutrient
concentrations of manure and compost were similar in all application years (both
materials were obtained from the same facilities since the start of the experiment in
1998—see also Krey et al., 2013), the concentration of organic P forms can be different
and these results should be considered as approximations when interpreting the results.
CHAPTER 6
109
Table 6.4. Characteristics of the organic amendments applied in 2010.
Means of three analytical replicates ± standard deviation expressed per kg moist amendment.
Values in parenthesis below each form of organic P are the proportion of total P extracted with
NaOH-EDTA ± standard deviation.*Aqua regia extractable; **NaOH-EDTA extractable P; ND: non-
detected.
6.4.4. Effect of long-term P amendment practice on organic and inorganic P
forms in the NaOH–EDTA soil extracts.
The extraction of soil with NaOH–EDTA showed a complete recovery of the total P in the
topsoil. The concentration of the different P forms in the NaOH–EDTA soil extracts differed
with respect to the amendment used (Table 6.5). The compost + TSP treatment was found
to have the highest inorganic and total organic P contents. Generally, in this study the
ratio of inorganic P:organic P was higher in the organic amendments than in the soil
(Table 6.4). Organic P applied to fields can be mineralized, but inorganic P can also be
immobilized into microbial and plant biomass and thus converted into organic P
(Bünemann et al., 2008; Richardson and Simpson, 2011). Soil analyses of the combined
compost + TSP treatment, which showed the highest total organic P contents and the
lowest inorganic P:organic P ratio (1.14) in soil underline this process. A conversion of
inorganic into organic soil P forms in this experiment is also supported by the fact, that
during the experimental time only about 50 kg ha-1 organic P were applied with the
organic amendments (calculation on the basis of the characteristic of the organic
amendments in 2010). The rest of total P supply (see Table 6.1) was in inorganic form. The
50 kg ha-1 corresponds to a soil P concentration of about 12 mg kg-1 (soil bulk density 1.35
g cm-3, 0–30 cm), however, the difference between the concentration of organic P in soil
in the compost + TSP treatment and the control treatment in 2012 was more than 60 mg
kg-1. Nutrients are also provided with crop residues (Eichler et al., 2004; Noack et al., 2012).
Cattle manure Biowaste compost
(mg P kg-1)
Total P* 1802 2010
Inorganic P** 1510 ± 272 1834 ± 182
Organic P** 327 ± 25 249 ± 37
Simple monoester-like P 197 ± 107
(11 ± 6)
ND
-
DNA-like P 3 ± 4
(0.2 ± 0.2)
146 ± 28
(9 ± 2)
Ins6P-like P ND
-
109 ± 77
(7 ± 5)
Non-hydrolyzable P 127±111
(7 ± 6)
ND
-
pH
Dry matter (%)
8.9
30.2
8.6
50.1
Organic and inorganic P forms affected by organic amendments
110
However, in our experiment the amount of organic P applied can be expected to be
low, since straw remained at the field only from 1999 until 2005 and in 2012 when cereals
were cultivated, whereas in the other years the whole plants were removed from the field
(see chapter `Site characterization´). Thus, organic P applied with residues can explain
differences in organic soil P contents between the treatments only partly. In addition,
Randhawa et al. (2005) showed that plant residues can increase the P mineralization in
soil, and plant residues must not result in higher amount of organic P in soil.
The amendments did not increase the enzyme hydrolyzable organic P pools
compared with the control (Table 6.5). The proportion of hydrolyzed (the sum of simple
monoester, DNA and Ins6P-like P) to total organic P extracted by NaOH–EDTA ranged
from 26% (compost + TSP) to 50% (compost). The maximum release achieved was similar
to that observed by Annaheim (2013) (37–50%) applying an inorganic fertilizer and
different organic amendments, and higher than that reported by Johnson and Hill (2010)
(25–36%) for soils amended with poultry manure. He et al. (2004a) found that as little as
29% and as much as 49% of organic P was enzymatically hydrolyzable in soils with and
without long-term swine manure application extracted with NaOH (previously extracted
with water and NaHCO3).
Ins6P-like P was the main form of enzymatically hydrolyzable P in all treatments
ranging from 19% (compost + TSP) to 38% (manure) of organic P extracted with NaOH–
EDTA. Several studies reported that Ins6P-like P is the major hydrolyzable form in most soils
using different extractants like water (Turner et al., 2002b), NaOH (He et al., 2004a, 2008)
or NaOH–EDTA (Keller et al., 2012). This abundance can be explained by a strong affinity
of Ins6P to soil surfaces which protect it from degradation, causing its accumulation in
soil (Celi et al., 1999, Steffens et al., 2010). Based on the characterization of the organic
amendments applied in 2010, plots amended with compost have received the highest
amounts of Ins6Plike P, while plots amended with manure have received the lowest, as
Ins6P-like P was not detected in the manure NaOH–EDTA extract (Table 6.4). However,
the concentration of Ins6P-like P in soil in the plots with compost addition was not higher
than in the plots with manure addition. This suggests that inputs of Ins6P-like P via organic
amendment may have no or only little influence on the amount of this form in the topsoil.
Crop residues and harvest losses can be also a source of Ins6P in soil (Noack et al., 2012).
However, as stated before, the amount of organic P provided with plant residues must
have been low, and Ins6P-like inputs by plant residues should be proportional to crop
yield and P uptake, so this form of organic P also would be expected to be higher in the
compost + TSP plots than in other treatments. Ins6P-like P accumulation in soil could be
also from microbial origin (Makarov et al., 2002). In our study we found a positive
correlation between the activity of DHA and the Ins6P-like P concentration in soil (r =
CHAPTER 6
111
0.588*, n = 19) which can support this fact. Further analyses (e.g. 31P NMR) of the NaOH–
EDTA extract could help to interpret the high amount of not hydrolyzable P in the
compost + TSP treatment, together with the decrease in Ins6P.
Monoester-like P and DNA-like P forms were also detected in NaOH–EDTA soil
extracts but in a lower concentration and no differences between treatments were
found. The proportion of simple monoester-like P to organic P ranged from about 2%
(control) to 8% (compost) while the proportion of DNA-like P to organic P ranged from 4%
(compost + TSP) to 11% (manure). Compost addition represents an important input of
DNA-like P to soil, whereas cattle manure contains important amounts of simple
monoester-like P. However, the low percentage of these forms of P in the topsoil may
reflect the previous degradation (Dick and Tabatabai, 1978) or the leaching to deeper
soil layers (Anderson and Magdoff, 2005). In addition, some orthophosphate diesters
could have been degraded rapidly in NaOH–EDTA extracts, which may underestimate
the concentration of DNA-like P and limit the characterization of soil organic P with this
method applied (Makarov et al., 2002; Turner et al., 2003b).
A large fraction of soil organic P remained non-hydrolyzed in NaOH–EDTA extracts
(47–73%). This fraction includes more complex forms associated with humic substances
(Brannon and Sommers, 1985) not accessible for hydrolytic enzymes. These proportions
are in accordance with other similar studies (He et al., 2004a; Keller et al., 2012).
Organic and inorganic P forms affected by organic amendments
112
Table 6.5. P classes in NaOH-EDTA soil extracts (0-30 cm) collected in autumn 2012 after 14 years of different P amendment management.
Treatment Inorganic P
Organic P
Simple monoester-
like Po DNA-like Po Ins6P-like Po
Non-hydrolyzable
Po Total*
(mg P kg-1 )
Control, no P 249 ± 26 a 4.5 ± 6.9 a 13.4 ± 6.3 a 59.2 ± 16.4 ab 108.7 ± 26.6 a 186 ± 9.6 a
Cattle manure 272 ± 15 ab 8.8 ± 5.9 a 20.6 ± 11.8 a 78.4 ± 16.3 b 108.5 ± 64.3 a 216 ± 59 b
Biowaste compost 275 ± 14 ab 14.1 ± 10.0 a 12.5 ± 13.0 a 69.1 ± 10.1 b 95.5 ± 31.6 a 191± 34 a
Compost + TSP 289 ± 16 b 7.6 ± 9.7 a 10.5 ± 12.3 a 47.9 ± 10.8 a 186.3 ± 59.9 b 252 ± 71 c
p value 0.029 0.377 0.586 0.023 0.033 0.048
(% of total organic P)
Control, no P - 2.4 ± 3.6 a 7.3 ± 3.6 a 32.0 ± 9.5 b 58.3 ± 13.0 ab -
Cattle manure - 4.8 ± 3.8 a 11.0 ± 8.3 a 37.5 ± 7.7 b 46.7 ± 17.5 a -
Biowaste compost - 7.9 ± 6.3 a 6.5 ± 6.5 a 36.7 ± 6.7 b 48.9 ± 9.0 a -
Compost + TSP - 3.6 ± 4.9 a 3.8 ± 4.3 a 19.2 ± 2.4 a 73.4 ± 5.8 b -
p value - 0.377 0.362 0.003 0.013 -
Mean ± standard deviation of five field replicates. Letters in the same column indicate significant differences between treatments (p<0.05); Compost + TSP=
Biowaste compost + Triple-superphosphate.
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113
6.5. CONCLUSIONS.
The long-term amendment practice affected the organic and inorganic P forms in soil.
However, the amount of organic P applied with the amendments cannot explain the
concentration of organic P in soil. Likely, the organic P forms in soil are a result of the
multifaceted P turnover processes in soil which are affected by plants, microorganisms
and abiotic factors. Higher ratio of P organic:P inorganic in soil than in the organic
amendments supported this. To include further treatments and field experiments can
help to gain a better understanding of factors affecting organic P classes in soil. The
results showed that the application of P with organic amendments can be considered
to be an adequate P source since the high-soluble contents of inorganic P in soil (Pw and
Pdl) could be warranted for a period of 14 years, when the P supplied is balanced with
the P removed by harvested crop products.
Acknowledgments.
M.I. Requejo is thankful for the grant conceded by the Instituto Nacional de Investigación
y Tecnología Agraria y Alimentaria, Spanish Agency.
CHAPTER 7:
AGRONOMIC EVALUATION OF WINE-DISTILLERY
WASTE COMPOST ADDITION TO A MELON
CROP: NITROGEN AND PHOSPHORUS
CHAPTER 7
117
7.1. ABSTRACT.
The high generation of organic wastes together with the increasing interest in developing
a sustainable agriculture convert the recycling of these materials as source of organic
matter and nutrients in a good option of management. A field experiment was
established during 2011 and 2012 to evaluate the use of a compost made from wastes
from the winery and distillery industry in an irrigated melon crop traditionally grown in the
area where these wastes are generated. A randomized complete-block design was used
with four treatments consisting on three doses of compost: 1 (D1), 2 (D2) or 3 (D3) kg
compost per linear meter of plantation and a control (D0) without compost application;
and the effect on plant growth, nitrogen (N) and phosphorus (P) accumulation and fruit
yield and quality was studied. The application of compost produced a slight increase in
plant biomass accompanied by a significant improvement of fruit yield with respect to
the control treatment, obtaining the maximum yield with D2. Although the potential
effects of N and P derived from compost were partially masked by other inputs of these
nutrients to the system (N in irrigation water, P supplied through fertigation), an effect of
P was observed resulting in a higher number of fruits in the plots amended with compost.
An increase of N and P availability in soil with compost addition was also observed at the
end of the crop season.
Agronomic evaluation of wine-distillery waste compost
118
7.2. INTRODUCTION.
The interest in developing a sustainable agriculture is becoming more important day by
day. Consumers are increasingly concerned about the type, quality and origin of food,
while there is greater awareness about the production processes and the ecological
impact they cause. The utilization of mature compost to improve crop yield and nutrient
utilization in an eco-friendly way has been widely studied in several horticultural crops
(Montemurro et al., 2005; Pérez-Murcia et al., 2006; Miglierina et al., 2013).
Spain is the third world producer of wine with a total production up to 3 million
tonnes per year (FAO, 2013). The winery and distillery industries generate large amounts
of residues within a short period of the year (from August to October) which have
polluting characteristics such as low pH and phytotoxic and antibacterial phenolic
substances (Bustamante et al., 2008a). During the winemaking process, solid residues like
grape stalks are generated, as well as grape marc (composed of the skin, pulp and seeds
of the grape) and wine lees as by-products. Grape marc and wine lees are usually sent
to alcohol-distilleries, generating exhausted grape marc, lees cake and vinasses. The
composting of these wastes has been shown to be a feasible option for the use of these
wastes in agriculture soils, as source of nutrients and organic matter without causing
environmental damage (Bustamante et al., 2008b).
Several authors have reported positive effects after the application to soil of
winery and distillery waste composts. These include the enhancement of carbon stocks
and soil microbial activity (Bustamante et al., 2010; Mosetti et al., 2010; Paradelo et al.,
2011), the availability of macronutrients (Bustamante et al., 2011) and the suppression of
plant pathogens (Borrero et al., 2004; Diánez et al., 2007; Segarra et al., 2007).
The winery industry is located mainly in Mediterranean agroecosystems which are
characterized by semiarid conditions and degraded soils with a deficit of organic matter.
Horticultural crops have been also traditionally grown in these areas with high inputs of
water and fertilizers. Thus, the utilization of compost derived from the winery and distillery
residues to improve the yield and quality of crops within the area where they are
generated could be a sustainable option of agricultural management.
Spain is the main European producer of melon, occupying almost 30000 ha of the
territory and with a total production up to 8.7 million tonnes, mostly of Piel de sapo type
(MAGRAMA, 2012b). The largest areas are located in the southern half of the Iberian
Peninsula, especially in Castilla-La Mancha region (central Spain) with more than 10000
ha dedicated to this crop and where more than the half of the total wine production of
the country is concentrated (MAGRAMA, 2012a).
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119
Previous studies have demonstrated positive effects in plant growth after the
application of wine and distillery waste composts to horticultural crops for transplant
production (Reis et al., 1997; Bustamante et al., 2008c; Manenoi et al., 2009; Carmona et
al., 2012); in vegetable production under greenhouse (Inácio et al., 2000; Moral et al.,
2006); or as substrate in hydroponic production (García-Martínez et al., 2009). However,
there is a lack of information regarding the use of this kind of wastes in horticulture under
open-field conditions.
A field study was carried out in central Spain to evaluate the effect of the use of
exhausted grape marc compost on the growth, yield, quality and uptake of nitrogen (N)
and phosphorus (P) of melon, a crop cultivated in this area predominantly in field
conditions.
7.3. MATERIAL AND METHODS.
7.3.1. Experimental design.
Field trials were conducted in different and adjacent plots at La Entresierra field station
in Ciudad Real, central Spain (3° 56´W; 39° 0´N; 640 m altitude) during the growing
seasons of 2011 and 2012. The soil of the experimental site is a shallow sandy clay loam,
classified as Petrocalcic Palexeralfs (Soil Survey Staff, 2010) with a depth of 0.6 m and a
discotinous petrocalcic horizon between 0.6 and 0.7 m. Soil characteristics of the trial
plots are showed in Table 7.1. For the previous three years, the plots had been cultivated
with non-irrigated wheat (Triticum aestivum L.) that did not receive organic amendment
nor fertilizers. The area is characterized by a Mediterranean climate, with a strong
continental character and contrasting daily temperatures.
The compost used in this experiment was provided by a composting plant
located in Socuéllamos (Ciudad Real) which collects residues from the wineries and
distilleries throughout this area and through an aerobic transformation produces a
commercially compost composed of grape stalks, exhausted grape marc and lees cake.
The main properties of the compost used are shown in Table 7.2. This compost fulfilled the
criteria established by the Spanish legislation concerning the use of organic materials in
agriculture.
Agronomic evaluation of wine-distillery waste compost
120
Table 7.1. The physicochemical characteristics of the soil (Ap horizon) in the plots used
in 2011 and 2012 at the field experimental site.
Parameters 2011 2012
pH 8.4 8.4
EC (mS cm-1) 0.1 0.1
Organic matter (g kg-1) 22.2 22.6
N (Kjeldahl)(g kg-1) 0.9 1.0
NO3--N (mg kg-1) 14.1 11.2
NH4+-N (mg kg-1) 0.8 0.9
Olsen P (g kg-1) 28.3 16.9
Available K (mg kg-1) 479.3 325.6
Available Ca (mg kg-1) 1648.4 1343.3
Available Mg (mg kg-1) 396.0 622.3
The experimental design was a randomized complete block with the dose of
compost as factor of variation. Four treatments with four replications each were
established both years: no compost addition (D0); and 1 (D1), 2 (D2) or 3 (D3) kg of
compost per linear meter. The compost application took place on 20 April 2011 and 19
April 2012 using a tow-behind spreader. The compost was incorporated into the soil at
around 10-20 cm depth localized in the crop row, and immediately covered with soil.
Table 7.2. Characteristics of the compost used in 2011 and 2012.
Parameters 2011 2012
pH 9.2 9.8
EC (mS cm-1) 1.2 1.0
N Kjeldahl (g kg-1) 32.9 33.0
Organic matter (g kg-1) 545 600
C:N ratio 10.1 11.2
NH4+-N (g kg-1) 1.5 1.8
NO3--N (g kg-1) 0.2 0.3
Total P (g kg-1) 6.8 7.6
Total K (g kg-1) 33.7 21.1
Total Ca (g kg-1) 0.10 0.28
Total Mg (g kg-1) 0.67 0.64
Each year of experimentation, melon seeds of `Piel de Sapo´ melon (Cucumis
melo L. cv. Trujillo) were were germinated under greenhouse conditions from late March
until they had two or three real leaves. The seedlings were transplanted to soil covered
CHAPTER 7
121
with transparent plastic mulch on 11 May 2011 and 9 May 2012 at a density of 4444 plants
ha-1 (1.5 by 1.5 m), approximately 20 days after compost application.
The plots (12 x 15 m) had ten rows of eight plants each. Each row was irrigated by
a drip line and emitters of 2 L h-1, 0.5 m apart. The total irrigation applied was 341 mm in
the first year and 461 mm in the second.
All treatments received 120 kg ha-1 of P2O5 (phosphoric acid) for the whole season,
injected daily through the drip-irrigation system, starting with the irrigation scheduling. In
2011 it was applied from 27 days after transplanting (DAT), after the female flowering
phase. However, in 2012, the phosphorus application started from 21 DAT, before the
female flowering period. Nitrogen fertilizers were not applied due to the high
concentration of nitrates dissolved in the irrigation water. Due to the high content of
potassium in soil, this element was not applied. Phytosanitary treatments according to
standard management practices were applied throughout the growing season.
Meteorological information as the evolution of mean temperature and the rainfall
amount during each growing season together are graphically presented in Figure 7.2.
7.3.2. Plant growth and leaf area index.
Four plants per treatment were sampled at 48 and 90 DAT in 2011 and at 47 and 89 DAT
in 2012, coinciding with the formation of the first fruits and at the end of the crop
development respectively. Leaves, stems and fruits were separated and weighted to
obtain the fresh weight and the total leaf area was measured with a leaf area meter (LI-
3100C, LI-COR, Lincoln, NE). The dry weight of the different aboveground plant organs
was determined following oven drying at 80 °C to constant weight. The dry weight of the
melon plant was the sum of the different plant organs. The leaf area index (LAI) was
calculated by dividing the total leaf area by the ground area available for one plant
(2.25 m2).
Figure 7.1. Melon crop at 40 DAT (left) and detail of the fruit set (right).
Agronomic evaluation of wine-distillery waste compost
122
7.3.3. Nitrogen and phosphorus contents in aboveground organs.
Sub-samples of the dried plants were ground to a fine powder to determine N and P
concentration. Total N was determined using Kjeldahl method and total P was
determined after digestion with HNO3 and HCl and measured by ICP-OES. The N and P
accumulation in every organ was obtained as the product of the concentration and the
dry biomass. The total plant uptake was determined as the sum of N or P accumulated
in each aboveground organ of the plant.
7.3.4. Crop yield and quality.
Melons were harvested when a significant number of fully ripe fruits were observed in the
field and harvesting was carried out weekly. Each melon was weighed at harvest to
determine the total fruit yield (FY). For each harvest, four representative marketable fruits
from each replicate were analyzed for fruit quality. Flesh firmness was measured at four
zones of the cut equatorial surface using a Penefel (Agro-Technologie, Tarascon,
France), with an 8 mm diameter tip measuring the force necessary to penetrate into the
flesh. From the liquid obtained by liquefying the mescarp of each fruit, the total soluble
solids content were determined by handheld Atago refractometer.
7.3.5. Available N and P in soil.
Soil samples from the arable layer (0-30 cm) were collected before compost addition
and at the end of the growing season. The samples were extracted with deionized water
and 1 M KCl for the determination of NO3 and NH4 respectively. The NO3 concentration
in the extracts was determined with an ion-selective electrode and the NH4
concentration using spectrophotometric techniques. Available P was measured by the
Olsen method (Olsen et al., 1954), using an extraction with 0.5 M sodium bicarbonate
and measured by spectrophotometric techniques.
7.3.6. Statistical analysis.
Statistical analysis were carried out for each year separately. The effect of the compost
addition was evaluated by analysis of variance (ANOVA) on the measured parameters
and the Tukey test (p≤0.05) was carried out to establish statistical differences between
treatments.
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123
Figure 7.2. Evolution of daily mean temperature and rainfall during the crop seasons of 2011 and 2012. Continuous vertical lines represent the
beginning of different phenological stages of the melon plant. The discontinuous yellow line represents the beginning of the irrigation schedule
together with P fertilization.
Agronomic evaluation of wine-distillery waste compost
124
7.4. RESULTS AND DISCUSSION.
7.4.1. Plant growth and leaf area index.
Dry matter accumulation in the aboveground melon plants during the two samplings
made during the growing seasons of 2011 and 2012 is shown in Table 7.3 and Table 7.4.
Table 7.3. Dry matter production by the aboveground melon plant organs and leaf area
index (LAI) at the middle of the growth cycle on 2011 and 2012 (48 and 47 DAT
respectively).
Within each column and year, means followed by the same letter are not significantly different at
p≤0.05.
In 2011, at 48 DAT, a significant increase (p ≤ 0.05) of dry biomass in stem and fruit
as well as the whole plant was observed when the higher doses of compost were
applied, coinciding with the first fruit setting (Table 7.3). Leaf dry biomass increased by
34% in D2 with respect to D0, stem by 42% and the increase of the whole plant was 38%.
This increase of plant biomass was accompanied by an enhancement of the LAI by 29%
(D3) and 36% (D2) with respect to D0. At the final stages of the crop growth, no significant
differences in the accumulation of dry biomass were observed in any of the organs.
However, a positive trend to increase the plant dry weight and LAI with the application
of compost was noticed.
On the contrary, in 2012 the major differences between treatments were
observed at the end of the crop cycle (Table 7.4). A significant increase of 14% on
average was obtained in the fruit at 90 DAT with the application of compost, with no
differences between the different application rates. In addition, a tendency to increase
the stem dry weight was noticed in D2 and D3 (p ≤ 0.18). At the end of the plant
development, the application of the three doses of compost produced an average
Year Treatment Leaf Stem Fruit Whole plant
LAI
(g m-2)
2011
D0 80.2 a 31.9 a 27.8 a 139.8 a 1.09 a
D1 77.8 a 27.7 a 52.2 a 157.7 ab 1.20 ab
D2 107.5 b 45.3 b 39.5 a 192.3 b 1.49 b
D3 95.4 ab 35.9 ab 53.4 a 184.6 b 1.41 ab
2012
D0 136.6 a 54.3 a 19.6 a 210.4 a 1.84 a
D1 107.7 a 42.3 a 11.7 a 161.7 a 1.43 a
D2 122.9 a 47.2 a 33.3 a 203.4 a 1.63 a
D3 118.4 a 47.1 a 22.4 a 187.9 a 1.60 a
CHAPTER 7
125
increase of 10% in the whole plant with respect to the control. At this stage, the LAI was
lower in the control plants compared with the treatments using compost (p ≤ 0.12).
Table 7.4. Dry matter production by the aboveground melon plant organs and leaf area
index (LAI) at the end of the growth cycle of 2011 and 2012 (90 and 89 DAT
respectively).
Within each column and year, means followed by the same letter are not significantly different at
p≤0.05.
The enhancement obtained in plant growth with compost addition was slight and
sometimes it even did not occur. This was probably due to the large input of N from the
irrigation water, element that contributes primarily to the plant growth. Castellanos et al.
(2011) studied the effect of applying different N doses (from 30 to 393 kg N ha-1) to a
melon crop under fertigation and stated that leaf and stem biomass increased with the
dose of N while the fruit biomass reached a maximum at about 90 kg N ha-1. Our control
was supplied with 89 (2011) and 132 (2012) kg N ha-1 with the irrigation water so it did not
show a lack of this element (Chapter 4). Additionally, the low mineralization rate of
organic N after the application of this compost (from 5 to 15% of total N applied during
the growing season, Chapter 3) explained why we did not observe a large effect on leaf
and stem biomass with compost addition.
Other authors have also found positive effects on plant growth with the
combination of inorganic fertilization and organic matter. Ahmed and Shalaby (2013)
found that the application of manure compost or seaweeds improved the vegetative
growth of cucumber. Biochar have been also found to increase the dry weight of pepper
and tomato plants grown under fertigation (Graber et al., 2010). In the same line,
Tzortzakis et al. (2012) reported increases in pepper growth combining fertigation with the
application of municipal solid waste compost.
Year Treatment Leaf Stem Fruit Whole plant
LAI
(g m-2)
2011
D0 173.2 a 47.0 a 392.1 a 612.3 a 1.35 a
D1 211.3 a 54.1 a 381.8 a 647.2 a 1.54 a
D2 171.2 a 48.7 a 427.9 a 647.7 a 1.46 a
D3 200.2 a 53.0 a 438.5 a 691.7 a 1.65 a
2012
D0 276.7 a 76.7 a 417.6 a 770.9 a 1.95 a
D1 279.4 a 79.9 a 502.7 b 862.0 b 2.29 a
D2 285.5 a 87.0 a 460.9 ab 833.4 b 2.61 a
D3 291.6 a 90.7 a 468.9 ab 851.1 b 2.30 a
Agronomic evaluation of wine-distillery waste compost
126
7.4.2. Nitrogen and phosphorus plant accumulation.
Nitrogen plant accumulation at the middle and at the end of the plant growth cycle is
represented in Figure 7.3.
In 2011, the total N uptake by the melon plants was significantly affected by the
compost application. At 48 DAT, N leaf accumulation tended to increase with respect to
the control (maximum increase of 30% with D2) being this increase statistically significant
in the stem (41% increase with D2) and consequently in the whole plant. Nevertheless, at
the end of the growing cycle this effect was less pronounced and non-significant
differences with compost addition were observed. In 2012, no differences were observed
in N uptake for the whole plant in the two samplings, but N accumulation in the fruit
increased significantly for all the compost doses at the end of the crop cycle (9% on
average).
Regarding P, the uptake of this element followed a similar trend as N (Figure 7.4).
In 2011, the total P uptake by the melon plants was significantly affected by the compost
application. At 48 DAT, P leaf and P stem accumulation increased with respect to the
control when the higher doses of compost were applied (D2 and D3), obtaining the
largest increase with D2 (about 37% in both plant organs), whereas an increasing
tendency was also observed in the fruit (p≤0.16). At the end of the crop cycle, differences
in plant P uptake were due to a higher content of P in the fruits obtained in the plots
amended with D2 (58% increase)and D3 (35% increase). On the contrary, no differences
regarding plant P uptake were observed in the second year of study. Surprisingly, in the
first half of the growing season, a negative effect of compost application was observed
in leaf P accumulation, which decreased with respect to the dose of compost applied.
No differences on total plant uptake where observed neither at 48 nor 90 DAT in 2012.
Several studies have reported that exists a synergistic effect between N and P in
other crops as pepper (Qawasmi et al., 1999) or maize (Hussaini et al., 2008). Nitrogen
supply may induce changes in the rhizosphere which include the growth of root hairs and
the acidification of the soil-root interface (Riley and Barber, 1971) which could facilitate
the absorption of other nutrients as P.
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Figure 7.3. Nitrogen accumulation in leaf, stem, fruit and in the whole plant at the
middle and at the end of the growing cycle. For each year, means followed by the
same letter are not significantly different at p≤0.05. The bars indicate standard deviation
at n=4.
Agronomic evaluation of wine-distillery waste compost
128
Figure 7.4. Phosphorus accumulation in leaf, stem, fruit and in the whole plant at the
middle and at the end of the growing cycle. For each year, means followed by the
same letter are not significantly different at p≤0.05. The bars indicate standard deviation
at n=4.
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The two different patterns observed in both years of study regarding plant N and
P uptake could be explained attending to the moment of N and P application via
fertigation (see Figure 7.2). The irrigation schedule usually starts approximately two weeks
after the transplanting (Ribas et al., 2003) coinciding with the beginning of the plant
development. However, in 2011, due to the rainfall occurred just before this period, the
start of the irrigation schedule was delayed, starting at 28 DAT. During this non-irrigated
period, the male and female flowering occurred and the plant only took P supplied by
the soil or that mineralized from the compost. Phosphorus plays an important role in the
flowering and in the set of the fruits (Maroto, 2003). This could explain the higher uptake
of P in leaves and stems observed at 48 DAT in the amended treatments which could
have been mobilized to the fruits during the fruit development. On the contrary, in 2012
the irrigation schedule started before flowering, at 21 DAT. In this case, there was an
important supply of N and P during the flowering stage which suppressed the compost
effect so that no differences were observed between treatments. This could indicate that
in these conditions the plant P needs are covered with P applied via fertigation so that
no additional effect with compost addition is produced.
7.4.3. Fruit yield and quality.
The fruit yield, yield components and fruit quality data are shown in Table 7.5.
Table 7.5. Fruit yield and quality data of the growing seasons of 2011 and 2012.
Year
Treatment
Fruit yield Fruit weight Fruit number plant-1 Flesh firmness ° Brix
(t ha-1) (kg) (N)
2011
D0 29.6 a 3.07 a 2.20 a 2.64 a 11.6 a
D1 31.9 ab 2.91 a 2.48 ab 2.83 a 12.5 a
D2 36.4 b 3.08 a 2.66 b 2.85 a 12.3 a
D3 36.1 b 3.04 a 2.59 b 2.87 a 12.5 a
2012
D0 49.9 a 3.57 a 3.15 a 2.91 a 13.3 a
D1 55.2 b 3.76 b 3.30 a 2.88 a 13.9 ab
D2 55.0 b 3.79 b 3.12 a 2.90 a 13.8 ab
D3 56.1 b 3.82 b 3.22 a 2.86 a 14.2 b
Within each column and year, means followed by the same letter are not significantly different at
p≤0.05.
In both years of study the incorporation of compost increased significantly melon
yield. In 2011, only the application of the higher doses (D2 and D3) resulted in a higher
Agronomic evaluation of wine-distillery waste compost
130
yield with respect to the control (23% and 22% increase, respectively). However, in 2012,
the improvement of crop yield with compost was observed regardless of the dose
applied (11% increase on average). Similarly, Hossain et al. (2010) reported an increase
of 20% when combining biochar together with inorganic fertilization in a tomato crop in
comparison with inorganic fertilization alone.
In 2011, a higher number of fruits per plant was obtained with compost addition,
following the same pattern than fruit yield, while compost addition did not have a
significant effect on fruit weight. On the contrary, in 2012 the enhancement of melon
yield was attributed to a greater fruit weight obtained in the amended plots. The higher
number of fruits per plant observed in 2011 could be attributed to a higher P uptake in
the amended treatments. The number of fruits produced per plant depend on the
number of produced flowers and the fruit set and according to Maroto (2003)
phosphorus plays an important role in the flowering and fruit formation. However, the
effect of compost on melon fruit yield should not be only attributed to the higher input of
nutrients, as the presence of plant growth regulators and humic acids in compost could
also influence crop yield and quality. Several authors have reported positive effects in
yield with the application of humic substances to several horticultural crops (Salman et
al., 2005; Azcona et al., 2011; Shehata et al., 2011).
According to the results, the dose corresponding to 2 kg of compost per linear
meter (D2, 13 t ha-1) would be enough to achieve the highest improvement on melon
fruit yield, and no additional effect was observed for the highest dose (D3). This effect
was previously observed by Xue and Huang (2013) using a sewage sludge compost and
they attributed it to an increase of soil EC or in the content of heavy metals with the
application of large amounts of sludge compost. Bustamante et al. (2011), applying
different compost mixtures containing winery and distillery wastes to a calcareous
vineyard soil, reported an increase in the Zn, Mn and Cu concentrations in soil (in all cases
below the toxicity threshols) but no differences in soil EC with respect to the control. The
higher accumulation of these metals in soil with the application with the largest dose may
explain the absence of additional response with D3.
With respect to the quality parameters measured, a tendency to improve flesh
firmness was observed only in 2011 for all the compost treatments (p≤0.14). Attending to
previous studies, this parameter seems not to be affected by compost application, as
found for melon (Naidu et al., 2013) and watermelon (Liguori et al., 2015). The application
of compost tended to increase the Brix degrees in melon, being this increase non-
significant in 2011, but obtaining a significant increase of the Brix degrees in 2012 of 6.7%
with the application of the highest dose of compost. Other authors have also obtained
higher soluble solid contents in melon under organic fertilization (de Faria et al., 1994;
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131
Fernandes et al. 2003). Some studies have shown that nitrogen influence the soluble solids
content (Bhella and Wilkox, 1989; de Faria et al., 2000). However, previous studies carried
out in the same experimental conditions stated that total soluble solids were non-
affected by the N application (Castellanos et al., 2011, 2012). Welles and Buitelaar (1998)
reported that LAI has an effect on fruit soluble solids by extending the fruit developing
period resulting in an increase in fruit TSS.
There was a strong effect of the different years in the yield and growth of melon,
obtaining higher biomass and yield in 2012. Among the factors that could have
influenced these differences the temperature exposure is included (de Köning, 1990)
which was higher in 2012 (see Chapter 3). Additionally, temperature fluctuations after the
fruit set in 2011 could result in smaller fruits due to a lower cell proliferation during the early
stage of fruit development (Higashi et al., 1999). The accumulated biomass in all the plant
organs was higher in 2012 than in 2011 due to a longer duration of the plant growing
cycle. In 2011, due to the earlier appearance of powdery mildew, the plant senescence
started before than in 2012, which could also explain the different yield obtained in both
growing seasons.
7.4.4. N and P availability in soil for the subsequent crops.
Figure 7.5 shows available P (Olsen P) and mineral N (sum of ammonium-N and nitrate-
N) in soil before the application of compost (Initial) and at the end of the growing cycles
of 2011 and 2012.
As observed in the figure, available P in soil increased at the end of both growing
seasons in all the treatments. In the case of the control treatment, this enhancement was
of 20 mg kg-1 soil both years of study and was mainly a result of the application of
phosphoric acid via fertigation during the irrigation period. When applying compost, a
tendency to improve the availability of P in soil with respect to the unamended soil was
appreciated, but the differences observed were too slight and non-significant. Although
this compost is an important source of inorganic and available P (see Table 7.2), the
additional P supplied by the compost must have been sorbed by the soil particles, due
to the calcareous nature of this soil. In Chapter 5, we concluded that an excess of P
applied or mineralized from the compost accumulated in this soil in form of Ca
phosphates due to its high P sorption capacity.
Regarding mineral N, only the application of compost increased the
concentration of N min in soil. In the control treatment, although important amounts of
nitrate-N were added with the irrigation water during the whole irrigation period, this N
was taken up by the plants or leached to deeper soil layers so that it is not highly
Agronomic evaluation of wine-distillery waste compost
132
accumulated in soil. On the other hand, in the treatments with compost application,
mineral N in soil increased according to the dose applied as consequence of
mineralization of organic N as shown in Chapter 3. In this previous chapter, we
demonstrated that the incorporation of this compost as soil amendment in our field
experimental conditions did not contribute to increase N losses by leaching when
irrigating at 100% ETc, so that N mineralized from the compost remained at the end of
the growing season.
Figure 7.5. Available nitrogen and phosphorus in soil before compost addition (Initial)
and at the end of the melon growing season for the different compost doses. For each
year, means followed by the same letter are not significantly different at p≤0.05. The
bars indicate standard deviation at n=4.
7.5. CONCLUSIONS.
The application of compost derived from winery and distillery wastes resulted in slight
increases in plant biomass and leaf area index and in an improvement of melon fruit
yield. The dose of compost corresponding to 2 kg of compost per linear meter (13 t ha-1)
was enough to achieve the highest yield.
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133
The changes induced by the incorporation of compost were partially masked by
some aspects of this agricultural system, as the high concentration of nitrates in the
irrigation water and the P application through fertigation. However, an effect of P
derived from the compost in the plant uptake and crop yield could be observed the first
year of study, due to a delay in the beginning of the fertigation schedule which resulted
in a higher number of fruits in the amended plots. Some other beneficial aspects of this
compost not studied in this work (other nutrients, humic acids) could have also influenced
this crop improvement.
The higher availability of N and P in soil at the end of the crop season indicated
that this compost is a potential source of these nutrients for plant growth, however, further
experiments isolating the inputs of these elements to the system would be required.
CHAPTER 8:
GENERAL DISCUSSION
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137
Along the individual Chapters of this Thesis it have been addressed different aspects of
organic wastes recycling in agriculture, regarding N and P dynamics in the soil-plant
systems, that remained still poorly understood. In this Chapter, the results obtained from
the different experiments established are discussed jointly according to the objectives
proposed.
8.1. EFFECT OF THE APPLICATION OF WASTES FROM THE WINERY AND
DISTILLERY INDUSTRY ON NITROGEN DYNAMICS.
8.1.1. Effect on soil N mineralization rate.
In Chapter 3, soil N mineralization after addition of compost made from winery and
distillery wastes was evaluated under laboratory and field conditions. Both experiments
showed a slow N mineralization pattern when this material was incorporated to soil,
obtaining constants of mineralization (N0 from 42 to 71 mg kg-1; k from 0.009 to 0.015) in
the range of those reported by other authors using compost obtained from different
sources (Cabrera et al., 2005; Gil et al., 2011). In fact, the composting process is primarily
used in order to stabilize the organic matter so that microbial processes slow down (Senesi
and Brunetti, 1996; Pascual et al., 1999). This results not only in a greater resistance to
mineralization, but also prevents soil N immobilization which usually occurs when non-
stabilized winery and distillery wastes are applied (Flavel et al., 2005; Bustamante et al.,
2007). Our results showed that the composting of winery and distillery wastes resulted in
a well stabilized material which avoid the immobilization of N in soil. In Chapter 5 it can
be appreciated the effect of composting on the organic matter stabilization, where a
reduction of the C:N of some wastes derived from the wineries and distilleries (C:N from
13 to 21) was observed after the composting of these materials (C:N 7-8).
The stability of the compost used in these experiments is also reflected in Chapter
4, during the first stages of the field leaching experiment, when no differences between
the control and the compost treatments were observed regarding the concentration of
nitrates in the soil solution and N leaching.
The adjustment of the experimental data obtained in Chapter 3, to a widely
accepted model of soil N mineralization (Standford and Smith, 1972) was better if
considering mineral N in the soil-compost mixture than when subtracting N mineral from
the unamended soil, isolating the N mineral from the compost. This indicated that the soil
type is an important factor which influence N mineralization from the application of
organic wastes. As the experiments were carried out in a sandy clay loam soil, a lower N
General discussion
138
mineralization might be expected in fine-texture clayey soils (Hernández et al., 2002;
Bustamante et al., 2007). It is important to point out that under field conditions, other
factors as the air temperature or the N inputs apart from compost (e.g. N in irrigation
water) have influenced soil N mineralization. For this reason, an overestimation of N
mineralized in the field was obtained when this model is just adjusted for air field
temperature.
8.1.2. Effect of the application dose.
Mineral N in the soil of study increased with respect to the dose of compost applied as it
implies the incorporation of more mineral N and organic N susceptible to be mineralized
(Chapters 3, 7). However, the application of higher doses of compost resulted in a lower
proportion of total N that has been mineralized (1.61, 1.33 and 1.21% with the increasing
doses), and the mineralization constants did not increase proportionally to the dose
applied, in agreement with other studies found in the literature using other organic wastes
(Lindemann and Cárdenas, 1984; Hernández et al., 2002; Mantovani et al., 2006). The
application of large amounts of compost has been reported also to diminish soil microbial
activity probably due to the accumulation of heavy metals or the increase of soil EC that
can cause adverse effects on soil microorganisms (Xue and Huang, 2013).
This effect was also observed in the plant growth and crop yield of the melon crop
established in the field experiment in Chapter 7, in which an enhancement was
produced with the application of compost, but no differences were found between the
lower and the highest doses (see section 8.1.4). Bustamante et al. (2011), reported a
higher concentration of Zn, Mn and Cu in a calcareous soil amended with different
compost mixtures containing winery and distillery wastes with respect to an unamended
soil. Although the concentrations of these metals in soil reported in the mentioned study
were always below the toxicity thresholds, it is possible that the larger amounts applied
with the highest dose of compost in our study has limited the growth and yield of the
melon crop. For example, excessive levels of Cu has been reported to cause
morphological and physiological disorders in cucumber (Rouphael et al., 2008).
8.1.3. Effect of irrigation practices.
Irrigation in arid and semi-arid regions is a necessity for horticultural crops to achieve an
optimal yield and quality. However, nitrate pollution of aquifers is an undesirable effect
in these areas as consequence of the excessive amounts of water and N fertilizers added.
For this reason, in Chapter 4, the use of wine-distillery waste compost under different
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139
irrigation rates was evaluated since and agronomic and environmental viewpoint in a
melon crop. As occurred with the application of mineral N fertilizer (Kirnak et al., 2005;
Cabello et al., 2009), adding water in excess did not result in a higher melon crop yield
with compost application. Although other authors using organic fertilizers have found
higher yields when comparing irrigation supplying 100% of the crop evapotranspiration
with deficit irrigation (Abdel-Mawgoud, 2006; Lopedota et al., 2013), our results showed
that applying water above 100% of the crop evapotranspiration (e.g. 120% ETc) does not
mean higher yield profits. Numerous works stated that a the highest C decomposition
and N mineralization are produced close to the field capacity (Orchard and Cook, 1983;
Rodrigo et al., 1997), therefore applying water above the soil field capacity just increases
the losses of this resource.
On the contrary, a remarkable effect of the irrigation strategy was found
regarding N losses by leaching when applying the compost of study. If the irrigation
strategy is based on adjusting the water applied to the crop needs (100% ETc), the
application of compost (even at higher doses) did not increase N leaching during the
crop cycle. However, the higher drainage produced as a consequence of an excess of
irrigation (120% ETc) involved a higher amount of N leaching with the application of
compost which increased with the dose applied (16, 24 and 30 kg N ha-1 more than in
the unamended plots). This different pattern of N leaching depending on the irrigation
efficiency was also reported by Díez et al. (1997) using municipal solid waste compost in
a maize-wheat crop rotation.
8.1.4. Effect on the crop response.
In Chapter 7, the application of a wine-distillery waste compost was evaluated since and
agronomic viewpoint in a crop traditionally cultivated in central Spain, a melon crop
under drip-irrigation. The compost addition derived in a slight increase of plant biomass
(8-10% on average) and leaf area index (15 or 23% depending on the year of study). The
high concentration of N in the irrigation water (Castellanos et al., 2011) together with the
low mineralization of organic N after the application of this compost (Chapter 3) may
explain this low response. In the same line, the effect of compost addition on N
accumulation in the plant organs was masked by the N applied with the irrigation water,
obtaining only slight differences in the first stages of the crop development before starting
the irrigation schedule (year 2011). The incorporation of this compost enhanced melon
fruit yield both years of study, and D2 was enough to achieve the highest improvement
with respect to the unamended plots (23% increase in 2011, 10% in 2012). The total soluble
solids (°Brix), a quality parameter, increased as well with compost addition. However,
General discussion
140
based on the results obtained by Castellanos et al. (2012) and accordingly with the N
accumulation in the plant observed in our study, we could not attribute this effect to N,
because all the treatments (including the control) were supplied with sufficient amounts
of N from the irrigation water. Other elements as P (Maroto, 2003), interactions between
nutrients or even the presence of growth regulators and humic substances in compost
(Salman et al., 2005; Azcona et al., 2011; Shehata et al., 2011) could have influenced
crop yield and quality.
8.1.5. Risk associated.
The main risk associated with N losses in Nitrate Vulnerable Zones is the contamination of
aquifers by N leaching, which is an important concern especially when the aquifers are
the main source of water for human consumption. Attending to the low N mineralization
constants obtained after the incubation of soil with the wine-distillery waste compost
(Chapter 3), the nitrate leaching risk expected would be low. However, in drip-irrigated
systems, in which a fraction of the water applied is lost in drainage, N leaching is quite
probable. In Chapter 4, the risk of N leaching from the use of compost derived from the
winery and distillery wastes was evaluated under real field conditions. Two environmental
indexes developed by Castellanos et al. (2013) were used in order to achieve this
objective. The Impact Index (II) allowed us to determine if the N leachate along the
melon crop season exceed 50 mg L-1, the maximum concentration for drinking water
considered by the European Drinking Water Directive; whereas the Environmental
Impact Index (EII) determine if the leachate concentration exceed the nitrate
concentration in groundwater.
Considering both indices, the application of compost does not represent a risk to
increase the contamination of the aquifers whenever an efficient irrigation systems is
used which limits water losses via drainage, obtaining values of EII always below the limit
established at 1. On the contrary, if water is applied in excess to the plant needs, the
addition of large amounts of compost poses a higher risk of groundwater contamination,
presenting EII>1.
In Chapter 4, in order to evaluate the different doses of compost applied from an
economic and environmental viewpoint, the Management Effiency (ME) was
calculated. This index, developed by Castellanos et al. (2013), relates the crop yield and
the amount of N leached. The results showed that with an excess of irrigation ME
decreases as the dose of compost increases, as a result of the higher N leaching. On the
contrary, with irrigation adjusted to the plant needs, an adequate ME was observed for
the optimum dose (D2).
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141
8.2. FACTORS INFLUENCING MOSTLY ON THE AVAILABILITY OF PHOSPHORUS
IN SOILS AFTER THE APPLICATION OF ORGANIC WASTES.
8.2.1. Phosphorus forms in soil after the application of organic wastes.
The incorporation of organic wastes to soil influenced the inorganic and organic P forms
in soil. The application of organic wastes increased soil available P level (Olsen P, Pw, Pdl)
(Chapters 5, 6, 7). This enhancement of soil available P could be due to the release of
soluble inorganic P contained in the organic residues after its incorporation to soil
(Chapter 6), but also because of the effect of the organic matter addition on soil
processes (Chapters 5,6). In Chapter 5, we found that the application of organic
materials to soil enhanced the concentration of dissolved organic C which may serve as
source of readily available C and energy for soil microorganisms (Sánchez-Monedero et
al., 2004; Guerrero et al., 2007); so they may stimulate soil P mineralization. Moreover, in
Chapter 6, we observed a higher soil organic matter content, higher microbial activity
(DHA) and alkaline phosphatase activity in the amended plots in comparison with the
control; parameters which are associated with the growth of microbial populations and
microbial P and with the mineralization of organic P (Bünemann et al., 2006; Fliessbach
et al., 2007). In soils with a high P fixing capacity, organic wastes can modify the P sorption
capacity and release P to the soil solution (Chapter 5). In calcareous soils, although the
incorporation of organic wastes generally increase the concentration of the more labile
P pools (e.g. Olsen P, NaOH-NaCl-P, Chapter 5), it also contributes to enhance the most
stable pools (e.g. poorly soluble Ca-P), accumulating in soil, except with the application
of non-stabilized materials which contribute to release P from this fraction.
Regarding soil organic P, this pool also increased with the long-term use of
organic amendments, and even an immobilization of inorganic P is suggested in Chapter
6. However, the use of enzyme additions to soil extracts (Chapter 6) showed us that the
long-term application of organic wastes did not modify hydrolyzable P in soil, as found
previously by Annaheim et al. (2015) in a different soil. As in most soils, Ins6P-like P was the
main form of hydrolyzable organic P detected in the soil extracts (Turner et al., 2002a; He
et al., 2004a; Keller et al., 2012; Annaheim et al., 2015) due to the affinity of Ins6P to soil
surfaces (Celi et al., 1999; Steffens et al., 2010). The proportion of the other forms of
organic P detected (monoester-like P and DNA-like P) was much lower probably due to
previous degradation of these compounds (Dick and Tabatabai, 1978) or leaching
(Anderson and Magdoff, 2005). However, a major proportion of organic P was not
accessible for hydrolytic enzymes, in concordance with other studies (He et al., 2004a;
Keller et al., 2012).
General discussion
142
8.2.2. Effect of the soil type.
The effect of the addition of organic wastes was studied in soils with different
characteristics. In Chapter 5, the availability of inorganic P after the incorporation of
winery and distillery wastes was evaluated in two calcareous soils differing in the soil P
test (Olsen P), the CaCO3 and Fe content and in the size of the different textural fractions.
Different trends were found in the concentration of Olsen P along time, which remained
stable in the soil with a finer texture and higher CaCO3 and Fe content (S1) and increased
along time in the other one (S2). This allowed us to state that mineralization of organic P
might be higher in soils with coarse-textures and that the retention capacity of each soil
(influenced by P status, Fe and Al oxides, clay or CaCO3 content) also plays an important
role in controlling the availability of P in these soils.
In the acid loamy sand soil of Rostock (Chapter 6), a correlation between the
total P applied and the bioavailable P pools in soil was found after long-term application
of organic wastes. On the contrary, in calcareous soils this correlation only existed during
the first days after the incorporation of organic wastes (Chapter 5). This could be
explained by the presence of CaCO3 in these soils which resulted in the precipitation of
insoluble hydroxyapatite which reduced the soluble P concentration (Torbert et al., 2002),
together with the clay and iron content which determine P retention in soils (Sample et
al., 1980).
8.2.3. Effect of the waste type.
Phosphorus forms in soil were studied from the application of a wide range of organic
materials in which were included wastes from the winery and distillery industry (fresh and
composted) (Chapter 5), cattle manure and biowaste compost (Chapter 6). In Chapter
6, the effect of the combination of biowaste compost with an inorganic fertilizer (triple-
superphosphate, TSP) was evaluated as well. The inorganic and organic P forms in the
different organic wastes studied varied widely (Chapters 5, 6), but inorganic was the
major fraction of P present in the materials studied ranging from 50% to 96% of total P,
according to other studies (Sinaj et al., 2002; García-Albacete et al., 2012). In Chapter 5,
an increase of the proportion of inorganic P was observed in the residues subjected to a
composting treatment, as a result of P mineralization during the composting process.
In Chapter 5 we found that the availability of P in calcareous soils could be
influenced by the total P concentration, the proportion of inorganic P: organic P, the C:P
ratio or the water soluble organic carbon of the organic materials applied. We observed
that, in the first weeks after the incorporation of the organic wastes, a higher application
of inorganic P increased Olsen P in soils, whereas the application of organic wastes with
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143
a higher proportion of organic P, a high C:P ratio or water soluble organic carbon can
cause microbial P inmobilization in soil (Cheshire and Chapman, 1996; López-Martínez et
al., 2004). The Ca content of the organic wastes also was found to influence negatively
the recovery of Olsen P in soil, in concordance with previous studies using different
organic wastes (Robinson and Sharpley, 1996; Siddique et al., 2000).
In Chapter 6, we observed that the organic P forms applied with different organic
materials (cattle manure and biowaste compost) did not correspond with the
concentration of the different organic P forms in soil (monoester-like P; DNA-like P and
Ins6P-like P) and concluded that these are a result of a variety of P turnover processes in
soil.
8.2.4. Effect on the crop response.
Different crops and agricultural managements were used in the experiments developed
in Chapters 6 and 7. In Chapter 6, organic wastes were applied in a crop rotation
including oilseed rape, wheat, barley, maize, rye, sorghum and sunflower crops and
differences in the crop P removal within 14 years were observed. All the organic
amendments increased P removal when comparing with the control plots, increasing
according to the P applied.
In Chapter 7, the use of a wine-distillery waste compost was evaluated in a melon
crop traditionally cultivated in the same area where these wastes are generated (see
also section 8.1.4). The combination of organic and inorganic fertilization was evaluated
in this experiment, as phosphoric acid was applied by drip irrigation since the irrigation
schedule started. An effect of P derived from the compost application was observed in
one of the years of study, when more P was accumulated in the leaves and stems with
the higher doses of compost at the middle of the crop growing cycle. This derived in a
higher yield observed in the plots amended with compost, which showed a higher
number of fruits per plant. Phosphorus has been reported to improve the fruit set, due to
its important role in flowering and fruit formation (Maroto, 2003). Probably, this effect of
compost was only observed in the first year of study because the irrigation schedule was
delayed due to weather conditions, and the inorganic P application started later. This
indicated us that the P application via fertigation masks the potential effects of the
compost regarding this element in the second year of study, because for this crop and
in these conditions the 100 kg P2O5 per ha applied are enough for the optimal crop
growth.
General discussion
144
8.2.5. Risk associated.
Soils with application of inorganic or organic fertilizers have the potential to release P from
the soil to adjacent water bodies through surface or sub-surface movement, depending
on its capacity to retain this element. The degree of P saturation (DPS), has been used as
a tool to predict environmental limits for soil P, is usually expressed as a percentage and
calculated as the ratio of acid ammonium oxalate-extractable [P] to [Al+Fe] (van der
Zee et al., 1987; Breeuwsma and Silva, 1992; Schoumans et al., 2000). This index allowed
us to evaluate the environmental risk derived from the long-term application of organic
amendments in an acid sandy soil in Chapter 6. Schoumans (2000) reported that values
higher than 40% are critical with respect to P losses in non-calcareous soils. According to
this study, the application of organic wastes in this soil would represent a risk for P losses,
as values above this limit were observed both in the control and in the amended plots,
reaching values of more than 50% with the combination of biowaste compost and TSP.
Confirming this, the variation in the total P content in the topsoil was lower than expected
if comparing the values with those obtained from the P balance, which could be
associated with a downward movement of P into deeper soil layers (Oehl et al., 2002).
These differences were much higher with the highest fertilizer inputs (biowaste compost
+ TSP) that resulted in higher bioavailable contents in the topsoil, in agreement with
Godlinski et al. (2004).
To determine the environmental risks in the calcareous soils studied in Chapter 5,
we calculated DPS after 16 weeks of incubation with the different organic materials,
according to Pautler and Sims (2000): DPS (%)= STP * 100/(STP+PSI); where STP (soil test P,
Olsen P) and PSI (P sorption index) are expressed as mg kg-1. The soil with the highest
contents of clay, Fe and CaCO3 (S1) exhibited the greatest values of PSI and
consequently, the lowest DPS (ranging from 4.8 to 8.3%). The other soil of study (S2),
however, showed higher values of DPS, ranging from 21.3 to 25.9%, so this soil would be
more susceptible to soluble P losses. According to this index, soil DPS values up to 25-40%
are generally associated with a greater risk of P losses (Pautler and Sims, 2000).
Although the comparison of the DPS results calculated by the different methods
must be made with caution, the results suggested that the environmental risk derived
from the use of organic amendments is higher in the acidic soil (Chapter 6) rather than
in the calcareous soils (Chapter 5). Accordingly, dissolved P concentrations in runoff or
leaching has been found to be lower in calcareous than in non-calcareous soils (Torbert
et al., 2002; García-Albacete et al., 2014).
CHAPTER 9:
CONCLUSIONS
CHAPTER 9
147
From the results obtained in the present work, the following conclusions have been
drawn.
Regarding the availability and risk of nitrogen released from organic
wastes: wine-distillery waste compost addition in vulnerable zones.
1. The compost derived from winery and distillery wastes is a mature compost very
resistant to mineralization. Therefore the application of this compost to soil results in a slow
release of available N for the crop or susceptible for being leached to deeper soil layers.
2. The application of compost derived from these wastes enhances plant growth and
yield of melon grown under irrigation in a vulnerable zone, corresponding the optimum
dose to 2 kg compost per linear meter of plantation (13 t ha-1). However, this effect
cannot be attributed to the N released from the compost due to the high concentration
of N in the irrigation water in the conditions of this study. An irrigation schedule adjusted
to the plant needs or even a slight deficit of water is enough to achieve the best yield
improvements and an excess of irrigation does not result in additional yield profits.
3. According to the environmental indices proposed (Impact Index and Environmental
Impact Index), the application of compost does not represent a risk to increase the
contamination of the aquifers by N leaching whenever an efficient irrigation systems is
used. However, irrigation water applied in excess results in higher drainage losses
together with higher N leaching losses with the application of compost.
4. The application of large amounts of wine-distillery waste compost slows down N
mineralization. This also affects crop growth and yield of melon crop obtaining no
additional profits when adding a large dose of compost.
Conclusions
148
Regarding the availability and risk of phosphorus released from organic
wastes: effects in soil.
1. The incorporation of organic wastes increases both inorganic and organic P pools in
soils and enhances the availability of P. However, the distribution of P forms in soil is more
reliant on the soil P turnover processes than on the forms of P added with the organic
wastes.
2. Soil characteristics as the texture or the P retention capacity (influenced by P status,
Fe and Al oxides, or CaCO3 content) play an important role in controlling the availability
of P in soils after the application of organic wastes. Mineralization of organic P is higher in
soils with coarse textures and P released from organic wastes is less available in soils with
a higher P sorption capacity.
3. Waste characteristics as the total P, inorganic P, organic P, C:P ratio, the water soluble
organic C or the Ca content may influence the availability of P in soils. However, the
correlation between waste characteristics and soil available P depends on the soil of
study, being the effect of P added with organic wastes on soil available P less
pronounced in calcareous soils where P precipitate as insoluble hydroxyapatite. The
organic matter applied also influences soil P processes controlling the availability of P
(mineralization, sorption-desorption, solubilization). In calcareous soils, the application of
stable organic matter (e.g. compost) contributes to increase the more stable and less
available pools of P (e.g. Ca-P) whereas the application of less-stabilized organic
materials releases P from these pools.
4. The incorporation of organic wastes increases the availability of P in soils (measured as
Olsen P, Pdl or Pw). The long-term application of organic wastes contributes to increase
soil organic P, but mainly in forms not accessible to hydrolytic enzymes. The hydrolyzable
organic P forms (monoester-like P, DNA-like P and Ins6P-like P) are little affected by the
organic waste addition.
5. Phosphorus derived from the application of a wine-distillery waste compost may
influence fruit yield by increasing the number of fruits per plant of an irrigated melon crop.
6. The application of organic wastes increases the degree of P saturation (DPS) of soils
and consequently the environmental P risk. This index highly depends on the soil P
retention capacity of the soil and changes induced by the application of organic wastes
CHAPTER 9
149
are more influenced by the soil characteristics rather than on the soil type. Phosphorus
released to water is expected to be lower in calcareous soils with a higher P sorption
capacity.
CHAPTER 10:
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