university of warmia and mazury in olsztyn

Transcription

CONTEMPORARY PROBLEMS OF
MANAGEMENT AND ENVIRONMENTAL
PROTECTION
In series issued:
I.
II.
III.
IV.
V.
VI.
Soils of chosen landscapes
- edited by Prof. Dr. Bolesław Bieniek
Marsh – its function and protection
- edited by Prof. Dr. Andrzej Łachacz
Natural and cultural transformation of landscape
- edited by Prof. Dr. Krzysztof Młynarczyk and Prof. Dr. Marek Marks
Sewages and waste materials in environment
- edited by Prof. Dr. Wiera Sądej
Chemical transformation of environment
- edited by Prof. Dr. Krystyna A. Skibniewska
Environmental aspects of climate changes
- edited by Prof. Dr. Zbigniew Szwejkowski
2
UNIVERSITY OF WARMIA AND MAZURY IN OLSZTYN
FACULTY OF ENVIRONMENTAL MANAGEMENT AND AGRICULTURE
SEWAGES AND WASTE MATERIALS
IN ENVIRONMENT
Monograph
Edited by Wiera Sdej
Olsztyn 2009
Chapter Authors:
Mgr Janusz Augustynowicz, Dr. Agata Bartkowiak, Dr. Katarzyna Budziska, Mgr
Maciej Cieluk, Dr. hab. Boena Cwalina-Ambroziak, Prof. Dr. hab. Halina
Dbkowska-Naskrt, Dr. Wojciech Dbrowski, Dr. hab. Jacek Długosz, Prof. Dr.
hab. Danuta Domska, Mgr Marcin Duda, Mgr Beata Gałzewska, Dr. Helena Iglik,
Dr. Krzysztof Jówiakowski, Dr. Mohamed Hazem Kalaji, Dr. Marek Kalenik, Prof.
Dr. hab. Józef Koc, Dr. Justyna Koc – Jurczyk, Prof. Dr. hab. Teresa KorniłłowiczKowalska, Dr. hab. Joanna Kostecka, Mgr Łukasz Kubisz, Mgr Zbigniew Luliski,
Dr. hab. Stefan Pietkiewicz, Mgr Janusz Posłuszny, Dr. Szymon Róaski, Prof. Dr
hab. Stefan Russel, Dr. hab. Wiera Sdej, Dr. Paweł Skonieczek, Dr. hab. Boena
Szejniuk, Dr. Małgorzata Warechowska, Dr. Piotr Wasilewski, Dr. hab. Jadwiga
Wierzbowska
Edited by Dr hab. Wiera Sdej
Reviewer: Prof. dr hab. Józefa Wiater
Program board:
Prof. Dr. Józef Koc – chairman
Prof. Dr. Bolesław Bieniek
Prof. Dr. Andrzej Łachacz
Prof. Dr. Marek Marks
Prof. Dr. Krzysztof Młynarczyk
Dr. hab. Wiera Sdej
Prof. Dr. Krystyna Skibniewska
Prof. Dr. Zbigniew Szwejkowski
Technical editor: Andrzej C. ołnowski
Cover design: Wiera Sdej
Authors of photography: Wiera Sdej
Publishing co-financed by The Voivodship’s Found of Environmental Protection in
Olsztyn
ISBN 978-83-929462-1-2
© Copyright by Department of Land Reclamation and Environmental Management,
University of Warmia and Mazury in Olsztyn
Printing: Warmia and Mazury Center of Agriculture Consulting Service in Olsztyn
Edit. quire Print quire Copy of edition 2
Contents
PREFACE................................................................................................................. 5
CHAPTER I.............................................................................................................. 7
Krzysztof JóĨwiakowski, Teresa Korniłłowicz-Kowalska, Helena Iglik
Estimation of sanitary status of sewage treated in constructed wetland systems
CHAPTER II........................................................................................................... 23
Marek Kalenik, Maciej CieĞluk
Sewage treatment in gravel with assisting dolomite layer
CHAPTER III ......................................................................................................... 35
Józef Koc, Paweł Skonieczek, Marcin Duda
Potential for sewage water purification in an aqueous environment by a constructed
wetland
CHAPTER IV.......................................................................................................... 59
Justyna Koc-Jurczyk
Treatment technologies of municipal waste landfill leachates
CHAPTER V .......................................................................................................... 71
Wiera Sądej, Zbigniew LuliĔski, Janusz Posłuszny
Impact of municipal landfills on quality of ground and surface waters
CHAPTER VI ......................................................................................................... 95
Danuta Domska, Małgorzata Warechowska
The effect of the municipal waste landfill on the heavy metals content in soil
CHAPTER VII.......................................................................................................107
BoĪena Cwalina-Ambroziak, Jadwiga Wierzbowska
Effect of fertilization on the composition of soil fungi community
CHAPTER VIII .....................................................................................................119
Szejniuk BoĪena, Wasilewski Piotr, BudziĔska Katarzyna, GałĊzewska Beata, Kubisz Łukasz
Effect of compost from sewage sludge on plant development
CHAPTER IX ........................................................................................................129
Janusz Augustynowicz, Stefan Pietkiewicz, Mohamed Hazem Kalaji, Stefan Russel
The effect of sludge fertilization on choosen parameters of chlorophyll fluorescence and
biomass yield of jerusalem artichoke (Helianthus tuberosus L.)
CHAPTER X .........................................................................................................141
Wojciech Dąbrowski
Treatment and final utilization of sewage sludge from dairy waste water treatment plants
located in podlaskie province
3
CHAPTER XI ....................................................................................................... 153
Joanna Kostecka
Selected aspects of the significance of earthworms in the context of sustainable waste
management
CHAPTER XIII..................................................................................................... 171
Halina Dąbkowska-NaskrĊt, Agata Bartkowiak, Jacek Długosz, Szymon RóĪaĔski
The quality of soil tare from the sugar plant with regard to its utilization for soil
fertilization
4
PREFACE
Generation of liquid and solid waste is an unavoidable consequence of any
human life-sustaining or economic activity. The problems that waste causes, and
especially economic utilisation of waste or reduction of its negative impact on the
environment, have become an important issue all over the world. According to the
regulations binding in the European Union, waste should be mainly recycled or
utilised. The least desirable solution is the disposal of waste on landfills. When
waste facilities (landfills or wastewater treatment plants) are inappropriately located,
constructed or maintained, they often have an adverse influence on the environment.
Among the most severe problems is the migration of pollutants from waste to
surface and subsurface water, which is the main source of potable water for people.
The problem of water contamination should be treated as a priority because water
resources, once contaminated, will take many years to be purified and, in many
cases, cannot be successfully treated.
The rapidly increasing amounts of produced waste and stricter regulations on
the environmental conservation mean that the processes of wastewater treatment or
waste utilisation need to be improved. It is now a general tendency all over Europe
to produce composts from sewage sludge and municipal waste, because these two
types of waste are a source of mineral and organic substances, which are valuable
for soil fertility. It is obvious that this tendecy will also feature more strongly in
Poland. However, both types of biowaste create a series of problems related to
potential contamination of the environment. Sewage sludge from WTPs in small
towns and villages has better properties as fertilizer and is much safer for the
environment than waste treated in large urban agglomerations, especially the ones in
industrial regions. Similar problems arise with respect to composted unsegregated
municipal waste. Although both types of waste have many positive attributes, in
Poland composted waste is not readily used for soil fertilization. The concern such
waste use raises is to some extent justifiable because this type of waste is often
characterised by levels of toxic substances that exceed the norms. The value of
sewage sludge or soil waste composts is deteriorated mainly by the presence of
heavy metals and such xenobiotics as PAHs, PCBs or alcylophenol derivatives. The
permissible level of some xenobiotics in soil, e.g. trace elements, is strictly defined
and any excess over the threshold limit is dangerous to biological life. Therefore, it
is most recommendable to undertake research on suitability of composted waste for
fertilization, land reclamation or regeneration of soils in environmentally degraded
areas. An important component of the research on possible environmental utilisation
of composts produced from sewage sludge or solid waste is monitoring, consisting
of quality checks based on analyses of the content of these compost constituents, and
xenobiotics in particular, which can migrate to ground and surface water.
The present monograph contains the results of studies on utilisation
or environmental use of various types of waste. Both positive and negative aspects
of waste influence on the environment have been raised and discussed.
Wiera Sdej
5
6
CHAPTER I
Krzysztof Jówiakowski1, Teresa Korniłłowicz-Kowalska2, Helena Iglik2
ESTIMATION OF SANITARY STATUS OF SEWAGE
TREATED IN CONSTRUCTED WETLAND SYSTEMS*1
Introduction
Domestic sewage is one of the factors causing bacteriological contamination
of surface and ground waters. Untreated sewage brings to the waters enormous
amounts of microorganisms – bacteria, viruses, fungi, and protozoa, called the
allochtonic i.e. introduced flora. Most of the allochtonic flora is typical microflora
living in the gastrointestinal tract of humans and higher animals, constituting socalled physiological flora of the organism. It is mainly composed of rods
of Escherichia coli, enterococci Enterococcus faecalis and sporifying clostridia
Clostridium perfringens, that are excreted together with the faeces (ZAREMBA,
BOROWSKI 1997, LIBUDISZ, KOWAL 2000, SMYŁŁA 2005). Untreated domestic
sewage may also contain pathogens and potential pathogens, e.g. those causing
typhoid fever, paratyphoid fever, bacterial dysentery, campylobacteriosis, tularemia,
tuberculosis and cholera (KLUCZEK 1999). According to KLUCZEK (1999), the most
frequent pathogenic bacteria occurring in sewage include rods of Salmonella and
Mycobacterium tuberculosis. Other pathogenic bacteria isolated from sewage
include Clostridium, Yersinia, Brucella, Campylobacter, as well as Bacillus
anthracis, Vibrio cholerae, Listeria monocytogenes and enteropathogenic strains
of Escherichia coli (VENGLOVSKY et al. 1997, OSEK 1999).
Intestinal bacteria are excreted with the faeces in enormous amounts, e.g.
1 gram of human faeces contains on average ca. 1.3 x 107 cells of E.coli and
3.0 x 106 cells of E. faecalis (SMYŁŁA et al. 2003). Such huge amounts of bacteria in
the faeces contribute to the bacterial contamination of waste waters. The species and
quantitative composition of microorganisms occurring in sewage are closely related
with the health status of inhabitants who produce the sewage (SIMMONS 1997,
OLACZUK-NEYMAN 2003). The survival time of pathogenic bacteria outside
of the organism of a sick person is usually long enough for the bacteria to constitute
a hazard of proliferation of contagious diseases through the water (SIMMONS 1997).
The numbers of bacteria in waste waters are subject to notable variation, but
generally depend on the population inhabiting a catchment (GEORGE et al. 2002,
OLACZUK-NEYMAN 2003). Therefore, the numbers of bacteria in waste waters are
*
The research has been financed from the science funds for the years 2007-2010 as a research
project of the Ministry of Science and Higher Education No. N N523 3495 33
7
greatly varied. Usually the numbers of bacteria from the coli group, of faecal origin
(thermo-tolerant), in raw sewage vary from 106 to 108 in 100 cm3 (GEORGE 2002).
The monitoring and assessment of the level of contamination of waters and
waste waters more and more often involve the application of comprehensive
analysis of sensory, physicochemical and microbiological indices. Those last ones
permit reasonable approximation of the sanitary status of the environment, the time
of occurrence and duration of microbiological contamination, the type and source
of contamination, and the potential health hazard caused by the presence of
pathogens (ŁOMOTOWSKI, SZPINDOR 1999). Current obligatory microbiological tests
performed within the scope of assessment of sanitary status of sewage are based
primarily on isolating faecal contamination indicating bacteria that constitute
permanent natural intestinal microflora of humans and higher animals (KOSAREWICZ
et al. 1999).
In microbiological studies on sewage performed to date much less attention has
been paid to microscopic fungi. They have mainly been focused on the qualitative
composition of those microbial groups, and on the occurrence of pathogenic species
in particular. In Poland, studies on the occurrence of pathogenic and potentially
pathogenic fungi in sewage and sewage sludge have been conducted by ULFIG
(1981, 1986) and by GRABISKA-ŁONIEWSKA et al. (1993). GRABISKAŁONIEWSKA et al. (1993) studied the occurrence of yeast and yeast-like organisms in
municipal sewage. Those authors demonstrated that typical sludge species included
Geotrichum candidum and Trichosporon cutaneum – fungi potentially pathogenic
for humans. Frequently observed in sewage and sewage sludge pathogenic and
potentially pathogenic fungi include also dermatophytes causing dermatomycosis,
also so-called geophilous dermatophytes (ULFIG 1986). Those fungi participate in
processes of sewage treatment, e.g. in the removal of keratin matter, but also are
indicator microorganisms in the assessment of degree of contamination with
pathogenic microorganisms (ULFIG 1983, KORNIŁŁOWICZ 1993a). Information on
other microscopic fungi in sewage and sewage sludge are highly fragmentary
(BECKER et al. 1954, COOKE 1970, WOLLETT, HENDRICK 1970, ULFIG et al. 1996).
In particular, there is a lack of data on the population sizes of those microbial groups
in sewage, and on changes in their numbers in relation to the degree of purification
of liquid wastes.
In recent years there has been an increase in the level of ecological awareness of
inhabitants of towns and villages in Poland, and therefore, for purposes of protection
of the water environment, more and more sewage treatment installations are being
constructed, collective ones as well as small household systems. Small household
systems are installed mainly is areas that have no connection larger sewage disposal
systems. A solution that has been gaining increasing popularity is the constructed
wetland system.
In the world constructed wetlands have been in use for more than 30 years, for
the treatment of household, industrial, rainfall, and agricultural sewage (SEIDEL
1967, KICKUTH 1977). In Poland, the oldest constructed wetlands have been
in operation for over a dozen years (KOWALIK, OBARSKA-PEMPKOWIAK 1998]. In
most cases those are single-stage installations, with horizontal (HF-CW “horizontal
flow constructed wetland”) or vertical (VF-CW “vertical flow constructed wetland”)
flow of waste waters treated, in which reed or willow are employed (HABERL et al.
8
1995, KOWALIK OBARSKA-PEMPKOWIAK 1998). Recently, however, multi-stage
constructed wetlands are built, so-called hybrid systems, composed of two or three
HF-CW and VF-CW beds that ensure better conditions for biological purification of
waste waters (KOWALIK, OBARSKA-PEMPKOWIAK 1998, LUEDERITZ et al. 2001,
OBARSKA-PEMPKOWIAK, GAJEWSKA 2003, ARIAS et al. 2004, GAJEWSKA et al.
2004, TUSZYSKA et al. 2004, OBARSKA-PEMPKOWIAK 2005; OBARSKAPEMPKOWIAK, GAJEWSKA 2005; VYMAZAL 2005).
Constructed wetlands are the object of research in Poland as well as in the world,
yet so far there is a shortage of results concerning the microbiological and sanitary
status of sewage treated in constructed wetland systems. The objective of the study
presented herein is estimation of the sanitary condition of sewage treated in
4 constructed wetlands.
Objects and methods of the study
The study on the sanitary status of treated sewage was conducted at 4
constructed wetlands located within the Lublin Province. A characterisation of the
objects under analysis is given below.
Object No. 1. A soil-plant (single stage) vertical flow constructed wetland with
common reed Phragmites australis Cav. Trin. Ex Steud., with maximum throughput
of 60 m3⋅d-1, located in Sobieszyn. At present the mean diurnal amount of sewage
supplied to the wetland is 24.7 m3⋅d-1, and the hydraulic loading rate of sewage on
the surface of the bed is on average 0.020 m3·m-2·d-1. The object has been in
operation since 1995 and is located at the Agriculture Schools Complex in
Sobieszyn near Kock.
The constructed wetland is made up of a two-chamber preliminary settler (with
active volume of 75 m3), a sewage pumping unit, a distribution well, four parallel
beds with reed, with a combined surface area of 1227 m2, and a collector well. The
beds are constructed of several layers of soil, one layer of fabric, and drains. The
surface layer, with a depth of 0.2 m, is a humus cover.
Underneath is a layer of loose sand of the same depth, directly overlying
a filtering fabric 1.2 mm thick. Beneath the filter fabric there is a 0.3 m layer
of dolomite gravel with grain sizes of 16–32 mm. Underneath that there is a layer of
drains collecting the effluent, each with a diameter of 100 mm.
The next down and final layer, with a thickness of 0.1 m, is sand, directly
overlying a PEHD geomembrane, 1 mm thick, the function of which is to provide
total isolation of the bed from the natural soil. The effluent flowing out of the
system is a forest-edge ditch that directs the treated sewage to the soil (ŁOSZAK,
PODLASZEWSKI 2000).
Object No. 2. A soil-plant (single stage) horizontal flow constructed wetland
with willow Salix viminalis L., with maximum throughput of 2 m3·d-1, located in
Jastków. At present the mean amount of sewage supplied to the wetland per day is
1.2 m3⋅d-1, and the hydraulic loading rate of sewage on the surface of the bed is an
average of 0.006 m3·m-2·d-1. The installation has been in operation since 1994 and its
sole function is purification of domestic sewage from an 11-person household.
The constructed wetland has a two-chamber preliminary settler (with active capacity
of 13.7 m3) and a soil-plant bed with average depth of 1.1 m and a surface area of
9
186 m2. The bed is filled with lose medium-grained sand. The surface layer is
a humus cover planted with willow. The bed is isolated from the natural soil with
PEHD foil 1 mm thick. The receptacle for treated sewage flowing out of the bed is
a pond with surface area of 1190 m2 (DRUPKA et al.1992).
Object No. 3. A multi-stage soil-plant constructed wetland with both vertical and
horizontal flow, with willow Salix viminalis L. and common reed Phragmites
australis Cav. Trin. Ex Steud., with throughput of 0.6 m3·d-1, located in Dabrowica.
The installation purifies domestic sewage from a 6-person household, and has been
in operation since September2006. The first element of the system is a threechamber settler with active capacity of 4.6 m3. The second element is an
arrangement of two parallel systems of soil-plant beds: system I – first bed with
horizontal flow and willow (A) and second bed with vertical flow and common reed
(B), system II – first bed with vertical flow and common reed (C) and second bed
with horizontal flow and willow (D). All beds (A, B, C, D) have the same surface
area of 24 m2. Average hydraulic loading rate of each bed system is 0.006 m3m-2d-1.
The beds with willow (A, D) have a depth of 1.0 m, while the beds with reed (B, C)
– 0.8 m. Inclination of the bed bottoms is 3% in the direction of sewage outflow.
The beds are filled with crushed stone and medium-grained sand. The beds are
isolated from the natural soil by means of PEHD foil 1 mm thick. The receptacle for
treated sewage is a mid-field ditch (JÓ
WIAKOWSKI et al. 2006).
Object No. 4. A multi-stage soil-plant constructed wetland with vertical and
horizontal flow, with common reed Phragmites australis Cav. Trin. Ex Steud. and
willow Salix viminalis L., with maximum throughput of 0.45 m3·d-1, located in
Janów near Garbów. The constructed wetland purifies domestic sewage from a 3person household. The object was built at the turn of 2007 and 2008. The first
element is a two-chamber settler with active volume of ca. 8.4 m3. The second
element is a system of two beds: 1 – with vertical flow, with common reed
Phragmites australis Cav. Trin. Ex Steud. (surface area of 18 m2 and depth
of 0.8 m), 2 – with horizontal flow, with willow Salix viminalis L (surface area of 30
m2 and depth of 1.2 m). The beds are filled with crushed stone and loose mediumgrained sand. They are isolated from the natural soil by means of PEHD hydroinsulating geomembrane with thickness of 1 mm. Treated sewage is deposited to the
ground by means of filtering drainage planted with Miscanthus giganteus
[JÓ
WIAKOWSKI, GORAL 2007].
Samples of sewage for microbiological analyses were taken from the above
objects from the particular stages of purification, as follows: from the preliminary
settler – raw sewage, from the tank after the settler – sewage after mechanical
purification, from the tank after biological treatment – biologically treated sewage.
The samples for analyses were taken in conformance with the relevant standards
(PN-74/C-04620/00, PN-EN 25667-2: 1999), in February, May, August and
November of 2008. In the samples the numbers of coli group bacteria were
determined with the fermentation method, and the numbers of faecal bacteria from
the coli group in compliance with the current standards PN-75-C-04615/05 and PN77-C-04615/07.
Bacteria from the coli group are Gram-negative rods that do not form spores,
grow under relatively anaerobic conditions, and ferment lactose, producing acid and
gas, within 24-48 hours at temperature of 35-37ºC. They are classified in the genera
10
Escherichia, Citrobacter and Enterobacter within the family Bacteriaceae. Faecaltype coli bacteria occurring in sewage include Escherichia coli that settle in human
faeces and have the capability of fermenting lactose, producing acid and gas, within
24-48 hours at temperature of 440C. Determination of bacteria from the coli group
with the fermentation method was done by inoculation of decimal dilutions
of samples (dilution in Ringer fluid – PN-ISO 9308-1) in a binary system into
Ejkman liquid medium (lactose, bromocresol purple) in test tubes with Dürham
tubes, followed by incubation at 37ºC and 44ºC. The results were read after 24 and
48 days of culturing. Results were accepted as positive when the medium changed
colour completely (from purple to yellow) and gas was produced. Doubtful results
(small amount of gas at no or weak acidification) were verified by inoculation into
Endo medium, and subjected to complementary testing by inoculation for repeated
fermentation, making coloured preparation with the Gram method, and performing
the cytochrome oxidase test. Final results were read from Tables included in the
standards and given in the form of the most probable number (MPN) of bacteria
from the coli group in 100 cm3 of sample, and as the coli form count, i.e. the
smallest volume of tested sample in which coli group bacteria can still be observed.
The numbers of fungi were determined with the plate dilution method, using the
Martin medium (saprotrophic fungi) and the Sabouraud medium (fungi potentially
pathogenic for humans and animals). Saprotrophic fungi were cultured at 25ºC,
while potentially pathogenic fungi - at 30ºC. Colonies grown were counted and the
results were given in cfu1cm-3 of sample. In all cases 3 parallel replications were
made.
The efficiency of elimination of bacteria and fungi was estimated on the basis
of their mean population values in the input and output sewage from the particular
elements of the constructed wetlands analysed in 2008. The obtained results
of populations of coli group bacteria were compared with the values given in the
REGULATION OF THE MINISTER FOR THE ENVIRONMENT [2004] which introduces
five classes of water purity in Poland. In the microbiological aspect, that division is
based on the numbers of coli group bacteria in 100 cm3 of water and faecal type coli
group bacteria in 100 cm3 of water (Tab.1).
Table 1
Limit values of microbiological indices of purity of surface waters according to the
REGULATION OF THE MINISTER FOR THE ENVIRONMENT (2004)
Microbiological indices
Number of faecal type coli group bacteria
in 100 ml
Number of coli group bacteria in 100 ml
I
Limit values in water purity classes I-V
II
III
IV
V
20
200
2 000
20 000
>20 000
50
500
5 000
50 000
>50 000
11
Effects of removal of coli group and faecal type coli group bacteria
The results of determinations concerning the populations of coli group and
faecal type coli group bacteria in the sewage from the constructed wetlands under
analysis, at the particular stages of purification, are presented in Tables 2 and 3.
Raw sewage contained very large numbers of bacteria of the type of Escherichia
coli – the mean MPN value varied from 8.3106 bacteria in 100 cm3 of the sample
from the constructed wetland in Sobieszyn to 4.2107 bacteria in 100 cm3 of the
sewage sample from the constructed wetland in Jastków. The values for faecal type
E. coli were generally several-fold lower – mean MPN value varied from 2.1106
bacteria in 100 cm3 in the sample of sewage from the constructed wetland
in Sobieszyn to 8.2106 bacteria in 100 cm3of sewage sample from the system
in Dbrowica.
Table 2
Numbers of coli group bacteria (MPN) in 100 ml of sewage
from the constructed wetlands in 2008
Kind of sewage
Raw sewage
Treated sewage
Object No. 1 – Sobieszyn
II
V
3
70010
24000103
3
2410
24103
VIII
2400103
62103
XI
6200103
2.4103
Kind of sewage
Raw sewage
Sewage after settler
Sewage after willow bed
Object No. 2 – Jastków
II
V
3
7000010
70000103
3
700010
1300103
3
710
70103
VIII
2400103
2400103
240103
XI
24000103
6200103
6.2103
VIII
24000103
24000103
6200103
23103
XI
6200103
24000103
2400103
1.3103
24000103
24000103
23103
2.30103
6200103
24000103
130103
0.62103
VIII
24000103
62103
0.24103
XI
2300103
62103
23103
Kind of sewage
Raw sewage
Sewage after bed A
Sewage after bed B
Raw sewage
Sewage after bed C
Sewage afer bed D
Kind of sewage
Raw sewage
Sewage after reed bed
n.sew. – no sewage
Object No. 3 – Dbrowica
System I
II
V
7000103
24000103
7000103
700103
7000103
130103
3
0.710
70103
System II
7000103
2400103
3
700010
700103
3
2410
240103
2.40103
2.40103
Object No. 4 – Janów
II
V
n.sew.
24103
n.sew.
2.4103
n.sew.
0.62103
12
For comparison, in raw sewage at the sewage treatment plant in Czstochowa
106 faecal coli bacteria were found in 100 cm3 of sewage sample (SMYŁŁA et al.
2003), while in sewage at the treatment plant in Gdynia as much as 1.81020100 cm-3
faecal type coli bacteria were noted, and in Gdask – 9.31018100 cm-3 (SZUMILAS
et al. 2001). Whereas, the numbers of bacteria of Escherichia coli type in sewage
supplied to household sewage treatment installations with filtration drainage located
in the communes of Lubraniec and Nakło varied from 2.51 107 to 7.39 107
cfucm-3 (BUDZISKA et al. 2007), which – converted to values per 100 cm3 - gives
from 2.51·109 to 7.39·109 bacteria. Comparatively, those were populations from 100
to 1000-fold greater than those in the raw sewage supplied to the constructed
wetlands in Jastków and Sobieszyn, respectively.
Table 3
Numbers of faecal type coli group bacteria (MPN) in 100 ml of sewage
from the constructed wetlands in 2008
Kind of sewage
Raw sewage
Treated sewage
Kind of sewage
Raw sewage
Kind of sewage
Raw sewage
Sewage after bed A
Sewage after bed B
Raw sewage
Sewage after bed C
Sewage after bed D
Kind of sewage
Raw sewage
Object No. 1 – Sobieszyn
II
V
3
24010
6200103
7.0103
2.4103
Object No. 2 – Jastków
II
V
3
2400010
2400103
70000103
620103
3
0.2110
24103
Object No. 3 – Dbrowica
System I
II
V
2400103
240103
620103
130103
24103
130103
3
0.2410
21103
System II
3
240010
240103
3
62010
130103
3
6.210
240103
1.3103
0.24103
Object No. 4 – Janów
II
V
n.sew.
62103
n.sew.
0.24103
n.sew.
0.062103
VIII
1300103
23103
XI
620103
0.62103
VIII
620103
1300103
62103
XI
2400103
2400103
1.3103
VIII
24000103
6200103
1620103
23103
XI
6200103
6200103
2400103
0.24103
24000103
6200103
6.2103
0.62103
6200103
6200103
50103
0.13103
VIII
6200103
2.4103
0.24103
XI
620103
6.2103
0.62103
During the mechanical purification of sewage in the settlers of the constructed
wetlands under analysis a low efficiency of removal of bacteria of E. coli. and E.
coli of faecal type was observed. It was only at the biological stage that clear
13
effluents were obtained, with mean values of coli form count at MPN of 1.93103 –
80.8103 cells in 100 cm3 of analysed sample. The mean values of faecal coli form
count varied within the range of MPN 0.31103 – 21.9103 cells in 100 cm3 of
analysed sample. The results obtained are lower by 1-2 orders of magnitude than
data given in the literature.
In a study at the sewage treatment plant in Gdask, OLACZUK-NEYMAN (2003)
found, at the outlet, populations of faecal type coli bacteria at the level
of 104 – 105100 cm-3, and at Dbogóra, 2.4104 – 2.5105 in 100 cm3. In a study at
the sewage treatment plant in Czstochowa, populations of faecal coli bacteria at the
outlet were of the order of 103 – 104100 cm-3 (SMYŁŁA et al. 2003). Whereas, the
numbers of Escherichia coli bacteria in sewage on the outlet of the household
sewage treatment systems with filtration drainage in the communes of Lubraniec
and Nakło was from 8.03 101 to 9.07 101 cfucm-3 (BUDZISKA et al. 2007), which
corresponds to 8.03103 – 9.07 103 bacteria in 100 cm3. These results are similar to
those obtained for the constructed wetland systems under analysis.
In the opinion of KOSAREWICZ et al. (1999), after mechanical and biological
purification of sewage the number of coli bacillus rods usually varies from 1000 to
100 000 in 1 dm3. Application of additional purification processes permits further
reduction of the content of those microorganisms.
Populations of coli group and faecal type coli group bacteria obtained in the
multi-stage constructed wetland systems (objects No. 3 and 4) most often
corresponded to water purity classes II, III or IV. Treated sewage with those purity
classes can be used for agricultural needs, e.g. for the watering of gardens or lawns.
In the single-stage constructed wetlands (objects No. 1 and 2) the mean numbers of
bacteria from the coli group and faecal type coli qualify the treated sewage under
analysis in water purity classes IV or V. The highest numbers of bacteria of the type
of E. coli and faecal E. coli were noted at object No. 2 (constructed wetland in
Jastków), in operation since 1994.
Based on the bacteriological analyses performed it was found that the small
constructed wetlands under study are characterised by very good efficiency of
removal of bacteria of faecal type coli group and those of the coli group (99.60 –
99.99%). The best efficiency of removal of coli group and faecal type coli group
bacteria (above 99.91%) was obtained in the multi-stage systems in Janów - object
No. 4, and in Dbrowica – object No. 3 (system 2), and the worst in the single-stage
reed system in Sobieszyn – object No. 1 (under 99.66%). SZUMILAS et al. (2001)
report that modern sewage treatment systems are capable of eliminating more than
99.999% of coil group bacteria through biological purification. According to
TALARKO (2003), the efficiency of elimination of coli form bacteria in soil-plant
filters amounts to approximately 99%, while BERGIER et al. (2002) maintain that
constructed wetlands with horizontal flow are characterised by faecal bacteria
removal rates at the level of 98.8%.
Effects of removal of saprotrophic and potentially pathogenic fungi
Wastewaters and surface waters are the habitat of numerous fungi. In surface
waters there occur typically aquatic fungi, primarily zoosporic, as well as yeasts
(KORNIŁŁOWICZ 1991, KORNIŁŁOWICZ, SZEMBER 1991, DYNOWSKA 1995,
14
CZECZUGA et al. 2002, KIZIEWICZ 2004a,b). Next to those, depending on the degree
of pollution with allochtonic organic matter, in surface waters there occur,
frequently in large numbers, so-called non-aquatic fungi, most often of soil or
sewage origin (PARK 1972, KORNIŁŁOWICZ 1993a,b, 1994a,b, DYNOWSKA 1995).
Sewage fungi are characterised by notable diversity of taxonomic and
physiological groups related with plant and animal organisms and with soil,
including phytopathogenic species as well as those pathogenic for humans and
animals (BECKER et al. 1954, COOKE 1970, WOLLETT, HENDRICK 1970, ULFIG and
KORCZ 1983, GRABISKA-ŁONIEWSKA 1993, ULFIG et al. 1996). Therefore,
determinations of fungi at sewage treatment installations are significant not only
from the general biological but also from the sanitary point of view (ULFIG 1986).
TOMLINSON and WILLIAMS (1975), as well as KORNIŁŁOWICZ (1993) and ULFIG
(1993), are of the opinion that certain fungi play an important role in processes of
sewage purification, and are also used as bioindicators of pollution of surface
waters.
The results of determinations concerning the populations of saprotrophic and
potentially pathogenic fungi in the sewage from the constructed wetlands under
analysis are presented in Tables 4 and 5.
Raw sewage from the constructed wetlands under analysis contained fairly large
amounts of saprotrophic fungi – their average populations varied from 553 cfu1cm-3
of analysed sample of sewage from the system in Sobieszyn (object No. 1) to
2292 cfu1cm-3 of sewage sample from the constructed wetland in Dbrowica (object
No. 3).
In turn, the numbers of potentially pathogenic fungi were usually slightly higher
– their mean populations in the raw sewage varied from 705 cfu1cm-3 (object No. 1)
to 2808 cfu1cm-3 (object No. 3).
The analysed constructed wetland systems were fairly efficient in the reduction
of populations of saprophytic and potentially pathogenic fungi, so their numbers
in purified sewage were low. The lowest numbers of saprophytic fungi were noted
in sewage on the outlet of the constructed wetland in Dbrowica (object No. 3) and
in Janów (object No. 4) – at 7.0 and 6.0 cfu1cm-3 of analysed sample, respectively.
The fungal populations observed were similar to the number of those microbial
groups determined by GRABISKA-ŁONIEWSKA et al. (2007) in mains water after the
process of purification. The authors quoted, in samples of river water after the
process of purification, found more than 2 cfu after conversion per 1 cm3 of water
(1506 cfudm-3). Our own study shows that also the numbers of potentially
pathogenic fungi were the lowest in objects No. 3 and 4, at 12.0 and 8.0 cfu1cm-3 of
analysed sample, respectively. The highest numbers of saprotrophic fungi – 81
cfucm-3 of analysed sample – were recorded in treated sewage on the outlet of the
constructed wetland in Sobieszyn (object No. 1), and largest populations of
potentially pathogenic fungi – 881 cfu1cm-3 of analysed sample – in sewage
flowing out from the constructed wetland installation in Jastków (object No. 2).
The research results obtained indicate that the highest efficiency of removal
of saprotrophic and potentially pathogenic fungi (above 99.28%) was recorded in the
multi-stage constructed wetlands in Janów (object No. 4) and in Dbrowica (object
No. 3 - system I). The lower efficiency of removing fungal groups observed in
objects No. 1 and 2 is due to the long period of operation of those systems.
15
Numbers of saprotrophic fungi (cfu.cm-3) in sewage from
the constructed wetlands in 2008
Table 4
Object No. 1 - Sobieszyn
Kind of sewage
Raw sewage
Treated sewage
Kind of sewage
Raw sewage
Kind of sewage
Raw sewage
Sewage after bed A
Sewage after bed B
Raw sewage
Sewage after bed C
Sewage after bed D
Kind of sewage
Raw sewage
II
183.3
V
1466.7
VIII
336.7
XI
226.7
(±57.7)
(±75.1)
(±35.1)
(±41.6)
4.7
4.7
253.3
60.7
(±2.0)
(±0.5)
(±50.3)
(±2.1)
VIII
500.0
XI
250.0
(±50.0)
Object No. 2 - Jastków
II
V
2266.7
180.0
(±26.5)
(±305.5)
(±100.0)
106.7
1766.7
290.0
133.3
(±11.5)
(±51.6)
(±78.1)
(±37.8)
6.0
29.0
5.0
203.3
(±2.6)
(±2.6)
(±1.7)
(±25.1)
Object No. 3 - Dbrowica
System I
II
V
3200.0
2533.3
VIII
2666.7
XI
766.7
(±351.2)
(±500.0)
(±321.4)
(±57.7)
1200.0
933.3
676.7
126.7
(±173.2)
(±115.4)
(±35.1)
(±20.8)
146.7
11.3
50.3
386.7
(±20.0)
(±3.0)
(±2.8)
(±31.1)
3.7
15.0
6.0
4.0
(±0.5)
(±2.6)
(±2.0)
(±1.0)
2533.3
System II
3200.0
2666.7
766.7
(±351.2)
(±500.0)
(±321.4)
(±57.7)
1200.0
933.3
676.7
126.7
(±173.2)
(±115.4)
(±35.1)
(±20.8)
9.7
63.0
6.0
4.0
(±3.6)
(±10.5)
(±2.0)
(±1.0)
8.7
122.0
8.3
72.0
(±5.7)
(±10.5)
(±1.0)
(±4.0)
VIII
1633.3
XI
720.0
(±17.3)
(±115.4)
(±30.0)
3.3
11.7
79.3
(±1.1)
(±2.1)
(±4.5)
Object No. 4 - Janów
II
V
140.0
n.sew.
n.sew.
n.sew.
16
1.3
8.0
9.3
(±0.5)
(±2.0)
(±1.1)
Table 5
Numbers of potentially pathogenic fungi (cfucm-3) in sewage from
the constructed wetlands in 2008
Kind of sewage
Raw sewage
Treated sewage
Kind of sewage
Raw sewage
Kind of sewage
Raw sewage
Sewage after bed A
Sewage after bed B
Raw sewage
Sewage after bed C
Sewage after bed D
Kind of sewage
Raw sewage
Object No. 1 - Sobieszyn
II
V
230.0
1966.7
VIII
406.7
XI
216.7
(±60.8)
(±251.6)
(±66.5)
(±28.8)
1.7
6.7
27.0
116.7
(±0.5)
(±1.5)
(±4.3)
(±15.2)
VIII
800.0
XI
323.3
Object No. 2 - Jastków
II
V
3233.3
130.0
(±10.0)
(±251.6)
(±173.2)
(±3.8)
83.3
2666.7
340.0
280.0
(±15.2)
(±155.7)
(±36.1)
(±6.4)
7.0
3333.3
11.7
173.3
(±1.0)
(±161.7)
(±5.6)
(±11.5)
Object No. 3 - Dbrowica
System I
II
V
3600.0
3833.3
VIII
2933.3
XI
866.7
(±208.1)
(±316.5)
(±450.9)
(±52.7)
2333.3
933.3
976.7
160.0
(±305.5)
(±321.4)
(±35.1)
(±20.0)
206.7
12.0
190.0
413.3
(±66.5)
(±3.0)
(±17.3)
(±37.8)
12.0
21.0
4.0
9.0
(±5.2)
(±6.2)
(±1.0)
(±1.0)
3833.3
System II
3600.0
2933.3
866.7
(±208.1)
(±316.5)
(±450.9)
(±52.7)
2333.3
933.3
976.7
160.0
(±305.5)
(±321.4)
(±35.1)
(±20.0)
18.3
123.0
9.0
77.0
(±3.7)
(±11.3)
(±1.7)
(±3.6)
13.0
148.3
15.0
6.0
(±2.6)
(±6.6)
(±0.6)
(±1.0)
VIII
2233.3
XI
1766.7
(±5.7)
(±387.6)
(±251.6)
4.0
19.7
70.0
(±1.7)
(±1.5)
(±4.0)
Object No. 4 - Janów
II
V
203.3
n.sew.
n.sew.
n.sew.
17
3.3
9.0
11.0
(±5.7)
(±2.6)
(±1.7)
Summary
Even though at present most sewage ends up in sewage treatment plants, it does
not solve the problem of bacterial contamination of surface waters. Classical sewage
treatment installations that do not perform specific disinfection reduce the numbers
of faecal bacteria by 1–3 orders of magnitude (GEORGE et al. 2002). As the level of
contamination of raw sewage supplied to treatment plants is very high, faecal
bacteria are also drained off, in enormous amounts, with treated sewage to the
environment (GEORGE et al. 2002). Even highly efficient purification of sewage,
with removal of biogenic substances, nitrogen and phosphorus, does not ensure
simultaneous effective elimination of microorganisms (OLACZUK-NEYMAN 2003),
as the efficiency of reduction of bacterial populations in the course of sewage
purification depends to a large extent on the numbers of bacteria in raw sewage.
SZUMILAS et al. (2001) found 99.999% reduction of the numbers of faecal type coli
group bacteria after sewage treatment, yet in spite of such an efficient operation of
the sewage treatment plant in question, with the initial pollution at the level of
1018- 1020, the numbers of those bacteria in treated sewage directed to the
environment were still huge.
The research results presented here indicate that multi-stage constructed wetland
systems ensure highly efficient – above 99% - elimination of bacteria and fungi,
while constructed wetland systems with a single soil bed (operated for more than a
dozen years) eliminate bacterial and mycological contaminations to a lesser degree.
To protect the aquatic environment from degradation it is necessary to employ
more and more efficient technologies of sewage treatment and to conduct
microbiological monitoring of the solutions applied. In recent years, certain
European countries have been introducing at least a partial disinfection of sewage
flowing out of purification plants. In Germany disinfection is applied to sewage
going to recreational areas, France disinfects sewage dumped in protected areas,
such as bathing zones and areas of mollusc growing, while in Spain disinfection is
applied with relation to sewage used for irrigation of arable fields, orchards, sports
fields, and gardens (SMYŁŁA et al. 2003).
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1
Dr Krzysztof JóĨwiakowski
Water and Sewages Analytics Laboratory
Department of Melioration and Agricultural Construction
University of Life Sciences in Lublin
ul. Leszczyskiego 7, 20-069 Lublin, POLAND
tel. +48 81 52 48 123, e-mail: [email protected]
2
Teresa Korniłłowicz-Kowalska, 2Helena Iglik
Mycological Laboratory
Department of Agricultural Microbiology
University of Life Sciences in Lublin
ul. Leszczyskiego 7, 20-069 Lublin, POLAND
21
22
CHAPTER II
Marek Kalenik, Maciej Cieluk
SEWAGE TREATMENT IN GRAVEL WITH ASSISTING
DOLOMITE LAYER
Introduction
The sewage economy in numerous villages in Poland and small cities is not well
organized. The most common method of removing sewage from apartment and farm
buildings is collecting sewage in the septic tank, then transporting it in a sewage
truck to a sewage treatment plant, sometimes on a field or to a ditch. Such a sewage
system is expensive in the exploitation, the septic tanks are often leaky and
improperly exploited. Sewages and sludge, carried away on field without
disinfection, create a big sanitary risk because of the presence of pathogenic bacteria
and eggs of parasites.
Expansion of the country water supply system and increase of the sanitary
facilities standard in flats evoked the increase of the sewage amount in households.
The construction of cumulative systems to collect and neutralize sewage is
impossible in many cases because of buildings dispersion, disadvantages of the
terrain topography and big investment costs. In these conditions a small sewage
treatment plant can be an alternative.
Small sewage treatment plants on country areas are recommended to apply on
terrains where the buildings are very dispersed, so the construction of sewage
systems is economically ungrounded. Sewage can be carried away to ground if
comes from detached houses, located outside of the underground water intake
protection zones and when the quantity of sewage does not exceed 5,0 m3⋅d-1
(ROZPORZDZENIE MINISTRA RODOWISKA [ORDER OF THE MINISTRY OF
ENVIRONMENT] 2006). There is also assumed an optimal unit quantity of sewage on
one inhabitant: in small settlement units (village) q = 120 dm3⋅d-1 in big settlement
units (city) q = 200 dm3⋅d-1 (PN-EN 752-4, 2001).
The purpose of the elaboration is to assess the effectiveness of sewage treatment
in the ground bed (gravel) with assisting layer (dolomite) under subsurface sewage
disposal field.
The small sewage treatment plants and technologies apply in them
On account of the applied technology of sewage treatment, small sewage
treatment plants can be divided on (KALENIK 2007):
• soil treatment plants – where the sewage is initially treated in mechanical way in
23
septic tanks and, as next, the sewage are cleaned thoroughly, directly in ground
bed (subsurface sewage disposal field) or in filter layers made of ground material
(sandy filter) (PN-EN 12566-3, 2007),
• soil-plant treatment plants - where the sewage is initially treated in mechanical
way in septic tanks and, as next, the sewages are being cleaned thoroughly in
filter layers made of ground with reed, willow or grass growing on the surface of
them,
• container treatment plants - small containers gathered in blocks, basing on the
technology of active sludge or bio-filter.
The small sewage treatment plants with subsurface sewage disposal field are
being built in well permeable grounds (gravels, sands) where the maximum level of
the ground water is at least 1.5 m below the sewage seepage level.
ENVIRONMENT] 2006). The purpose of this is to clean thoroughly the sewage in the
aeration zone, to stop the bacteria and viruses as well as to prevent contamination of
the natural environment. In fact, the small sewage treatment plants with subsurface
sewage disposal field should be built on the areas where the distance between the
sewage seepage level and maximum ground water level is at least 2.5 m. The recent
investigations point that during the exploitation of subsurface sewage disposal
fields, the humidity conditions of ground change within them depending on the
sewage seepage method being used (KALENIK, BŁAEJEWSKI 1999, KALENIK 2002,
KALENIK, KOZŁOWSKI 2007). Depending on the sort of ground, its hydraulic load
by sewages and impervious layer floor inclination, the ground water level is raising
up, reducing the real aeration zone (SROKA, KALENIK 1999, KALENIK 2000) in
which occur the oxygen processes of thorough cleaning. Whereas the small sewage
treatment plants with the sandy filter are being built in slightly permeable ground
(clay, silt) or if the ground water level is shallow under the ground surface. Small
container sewage treatment plants in the technology of active sludge or bio-filter can
be applied independently of hydro-geological conditions or landform features of
area.
The easiest and cheapest system is a septic tank cooperating with subsurface
sewage disposal field. This system is easy in construction and exploitation, it does
not require qualified service or technical and laboratory supervision. It can be
operated by a household owner, appropriately trained.
A subsurface sewage disposal field is a device serving to introduce the sewage,
initially treated in a septic tank, to ground. During filtering by natural layers of
ground, the sewage are being cleaned in biological processes under the influence of
oxygen bacteria and other microorganisms which take the oxygen from the ground
air. Small solid and colloidal suspensions are being stopped on the surface of sand
grains. Some part of sewage is being taken by plant roots, some raises toward the
ground surface thanks to the ground capillarity and evaporates, the remaining
infiltrates into ground waters. The arrangement of devices for the individual sewage
disposal with subsurface sewage disposal field is showed in Figure 1. The sewage
flows through a gravitational house sever (1) from the apartment building to the
septic tank (2), where should be kept for ca. 2 - 3 days, but no less than 1 day. From
the septic tank, the sewage flows to a distribution box (3) which directs them to a
perforated distribution pipe (5), finished with ventilation pipes (4). Then, the sewage
24
spills out in the sewage seepage bed (7) through the holes in perforated distribution
pipes (5) and further infiltrates into the ground.
3
2
a)
10
4
5
1
b)
1
2
3
5
7
4
6
d)
c)
6
7
8
5
e)
6
8
5
7
9
6
7
8
5
9
Fig. 1. Scheme of the subsurface sewage disposal field (KALENIK 2009):
a) horizontal projection, b) longitudinal cross-section, c) cross-section of drainage in averagepermeable ground, d) cross-section of drainage in slightly-permeable ground, e) cross-section of
drainage in easily-permeable ground, 1- supplying sewage pipeline, 2-septic tank, 3-distribution
box, 4-ventilation pipe, 5-perforated distribution pipe, 6-subsoil, 7-seepage bed, 8-barrier
material, 9-assist layer , 10- sewage infiltration surface.
In the septic tank an initial mechanical sewage treatment occurs, which must
reduce the value of BOD5 at least of 20% and the content of solid suspension at least
about 50% (ROZPORZDZENIE MINISTRA RODOWISKA [ORDER OF THE MINISTRY
OF ENVIRONMENT] 2006). It is so, because too big content of solid suspension in a
filtering sewage accelerates a silting-up of the ground under subsurface sewage
disposal field, what – as a result - diminishes the period of correct operation of the
device. In the septic tank, there occur sedimentation and flotation processes stopping
solid pollutants, as well as biological processes of anaerobic decomposition of
sludge gathered at the bottom of the container. The general capacity of the septic
tank cannot be smaller than 2.0 m3 (PN-EN 12566-1:2004/A1, 2006). If the capacity
is smaller than 4.0 m3, two-chamber septic tanks should be used, whereas if greater three-chamber, if septic tanks are not equipped with filter baskets. As a filling of the
filter basket, a ceramist or pozzuolana is being used. In order to limit the suspension
outflow from the septic tank, the bottom edge of the three-way pipe, the filter basket
or the shield should be plunged in the sewage at 0.4 m. Once or twice a year the
sludge gathered at the bottom of the container should be removed from the septic
tank.
25
Perforated distribution pipes are made of stiff PVC pipes of the minimum
internal diameter of 100 mm, in which round holes of the diameter 8.0 - 10.0 mm
spaced in 20-cm intervals are bored. The slope of the distribution pipe is 5.0 - 10 ‰.
The spacing of the pipes is assumed from 1.5 m to 2.0 m, and the arrangement depth
of the pipes - 0.8 - 1.2 m. Length of the pipes should not exceed 20.0 m. To provide
the ventilation of the seepage bed, ventilation pipes with the holes arose minimum
0.5 m above the level of the land are being installed on the ends of the distribution
pipes (KALENIK 2009).
The seepage bed consists of breakstone or rinsed gravel of the diameter of 15.0 40.0 mm. The bed’s thickness is equal 30.0 - 35.0 cm, its width - 50.0 - 120.0 cm.
The separating layer (8, Fig. 1), protecting the seepage bed before silting, can be
made of filtrating needled cloth or 5.0-cm layer of straw. (KALENIK 2009).
The construction of subsurface sewage disposal field in a low permeable ground
or in ground containing considerable amounts of decay, hummus or peat should be
avoided. After filtering in such bed, there remain some organic compounds in the
sewage and moreover the drowning of the seepage bed occurs. This phenomenon
should be avoided by decreasing of permissible unit hydraulic load of the subsurface
sewage disposal field and by applying under the seepage bed an assisting layer
which will extend the time of presence of the sewage in the layer deprived of
organic compounds.
The sewage, initially cleaned in the septic tank, is not safe on account of
bacteriological protection. Bacteria and eggs of parasites are removed in 99%
together with sedimentation sludge (HARTMANN 1999). In the accessible scientific
and technologic literature, there is little of a publication concerning the effectiveness
of sewage treatment in ground bed under subsurface sewage disposal field (REED I
IN. 1989, RETTINGER 1993, WILHELM I IN. 1994, SCHWAGER, BOLLER 1997,
SIEMIENIEC, KRZANOWSKI 2001, VAN CUYK I IN. 2001). Currently carried
investigations on the effectiveness of sewage treatment in the ground bed of coarse
sand (KALENIK, GRZYB 2001), dust sand (KALENIK, GRZYB 2003), gravel
(KALENIK, AMBROZIAK 2005), as well as with assisting layers of coarse sand
(KALENIK 2008) and of mineral ash (KALENIK, WILKOWSKA 2008) point out that
the kind of the ground bed affects the effectiveness of sewage treatment.
Conditioning of research
To measure the effectiveness of sewage treatment in the ground bed under
subsurface sewage disposal field, there was built a measurement stand in the form of
a tight container of the size: length 1.20 m, height 1.70 m, width 0.20 m (Fig. 2).
The container was made of plastic plates, fastened in metal frames. The sewage was
being pumped by a pump from the container through a delivery pipeline to a
perforated distribution pipe of the diameter of 100.0 mm laid on a ground bed layer
made of stones of the diameter of 20.0 - 40.0 mm. The pump was turned on and off
by a programmer. The size of the seepage bed layer is: length 0.50 m, width 0.20 m,
height 0.20 m. The sewage is filtered to the seepage bed layer through a hole of the
diameter of 8 mm placed in the bottom of the perforated distribution pipe. After
filtering through the seepage bed layer, the sewage is filtered through an assisting
layer into the ground bed. The assisting layer was made of dolomite and the ground
26
bed - of gravel. The researches were carried out for two assisting layers of the
thickness of 0.10 m and 0.20 m. The gravel layer thickness amounted to 1.30 m. In a
bottom of the measurement stand three holes were made, which enabled an outflow
of the filtered sewage through the assisting layer (dolomite) and ground bed (gravel)
to collecting vessels. The container was being filled by layers of the thickness of 5.0
cm, thickened by compacting.
5 6 7 8 9
2
3 4
1,70 m
11
10
12
1,20 m
Fig. 2. Scheme of the measuring stand:
1-tank, 2-pump, 3-programmer, 4-delivery pipe, 5-perforation distribution pipe,
6-seepage bed, 7-assist layer (dolomite), 8-ground bed (gravel), 9-transparent
plastic plate, 10-sewage outflow, 11-metal frame, 12-collecting vessel.
To researches, synthetic sewage was used, prepared according to PN-C04616/10 (1987). The synthetic (raw) sewage was being dosed three times a day and
its daily dose had been defined depending on the kind of the ground bed and
minimal acceptable hydraulic load of ground by sewage, according to Polish
recommendations. (CUGW 1971, TABERNACKI I IN. 1990). Before introducing the
raw sewage at the ground bed (gravel) with assisting layer (dolomite), as well as
after filtering them through the same layers, the following indicators of sewage
contamination were determined (ROZPORZDZENIE MINISTRA RODOWISKA
[ORDER OF THE MINISTRY OF ENVIRONMENT] 2006): solid suspensions, BOD5 and
COD, and additionally total nitrogen, total phosphorus, ammonia nitrogen, nitrate
nitrogen, nitrite nitrogen. The contamination indicators in the sewage were
determined once a week, taking into consideration the time of the sewage filtration
through the ground bed (gravel) with the assisting layer (dolomite).
The content of individual fractions of the ground graining was determined by
sieve analysis. The investigation was made on three samples and the obtained results
27
showed that it was gravel (KALENIK, AMBROZIAK 2005). The filtration coefficient of
gravel was determined in Wiłun apparatus ITB-ZW-K2. The measurement was
made for six samples. For the examined gravel, the filtration coefficient (k) amounts
to 0.005 m⋅s-1.
If the kind of ground (gravel) and its filtration coefficient (0.005 m⋅s-1) are
known, the daily dose of sewages can be determined - 3.0 dm3 - related to the length
of the perforated distribution pipe, according to the Polish recommendations
(CUGW 1971, TABERNACKI I IN. 1990). The hydraulic load of the distribution pipe,
according to Polish recommendations, relates to 1 m of its length and for gravel
amounts to 15,0 dm3⋅m-1⋅d-1. The daily dose of sewages was divided into three doses,
1.0 dm3 each, and they were being applied to the seepage bed at 7, 13 and 19
o’clock.
The synthetic (raw) sewage was being prepared every sixth day and the
indicators of contamination were being determined at the beginning, in the middle
and at the end of the dosing period, then they were being averaged (Table 1). The
solid suspension was determined by the gravimetric method. The BOD5 was
determined by the electrochemical Sensomat method of Lovibond. The COD was
determined by the titration with potassium dichromate. Total nitrogen, nitrite
nitrogen and total phosphorus was determined in the Hach spectrophotometer.
Ammonia nitrogen and nitrate nitrogen was determined by the colorimetric method.
Physical and chemical indicators of sewage
The carried-out research of the effectiveness of sewage treatment in ground bed
made only of gravel (KALENIK, AMBROZIAK 2005) shows (Table 1) that the solid
suspensions are not being removed in satisfactory degree and do not fulfill the
obligatory recommendations (ROZPORZDZENIE MINISTRA RODOWISKA [ORDER
OF THE MINISTRY OF ENVIRONMENT] 2006). However, the BOD5 and COD
indicators fulfill them.
The solid suspensions removing effectiveness fluctuated between 50% and 56%
and amounted to 53% on average. However, the effectiveness of decreasing of the
BOD5 indicator fluctuated between 98.4% and 99.5%, of the COD indicator between 84.7% and 88.1% and on average amounted to 98.7% and 86%
respectively. The total phosphorus in the cleaned sewage appeared in trace amounts
and its medium effectiveness of decreasing amounted to 98.3%.
Table 1
Physical and chemical indicators of raw sewage and treated sewage in ground bed of gravel
(mean values) (KALENIK, AMBROZIAK 2005)
Indicators
Solid suspensions
BOD5
COD
Total phosphorus
Unit
-3
[mg⋅dm ]
[mg O2⋅dm-3]
[mg O2⋅dm-3]
[mg P⋅dm-3]
Raw
sewage
172.00
109.70
308.00
1.70
9 week
86.00
0.60
47.00
0.00
28
Treated sewage
10 week 11 week
75.00
80.00
1.80
2.10
45.10
45.30
0.06
0.02
12 week
80.00
1.20
36.50
0.04
The analysis of the results presented in the Table 2 allows to state that after
filtering the raw sewage through the gravel with the dolomite assisting layer, the
content of the solid suspensions, BOD5, COD and total phosphorus in the cleaned
sewage decreased.
The ground bed with dolomite assisting layer of the thickness of 0.10 m started
to work properly after six weeks, however the ground bed with the assisting layer of
the thickness of 0.20 m - after four weeks. Under the seepage bed, a bacterial jelly of
the thickness of 2.5 cm formed, which is a site of bacteria and microorganisms. The
temperature in the room for all the research period was stable and amounted to
14°C.
Table 2
Physical and chemical indicators in raw and cleaned sewage in ground bed of gravel with
dolomite assisting layer (mean values)
Cleaned sewage
Indicators
Solid
suspensions
BOD5
COD
Total
phosphorus
Unit
Raw
sewage
Assisting layer thickness
0,10 m
0,20 m
7 week 8 week 9 week 5 week 6 week 7 week
[mg⋅dm-3]
117.0
0.0
0.0
0.0
0.0
0.0
0.0
[mg O2⋅dm-3]
[mg O2⋅dm-3]
213.0
261.0
1.5
34.8
1.8
34.5
1.5
30.1
1.3
29.8
1.4
26.5
1.3
27.1
[mg P⋅dm-3]
2.5
0.88
0.65
1.20
1.48
1.58
1.03
For the layer of the thickness of 0.10 and 0.20 m, the solid suspension removing
effectiveness amounted to 100%. A big amount of solid suspension introduced into
the ground bed causes its quick silting-up (ŁOMOTOWSKI 1999). As a result, the bed
permeability coefficient decreases and then – a life of sewage treatment plant with
subsurface sewage disposal field.
The BOD5 removing effectiveness for the layers of the thickness of 0.10 m and
0.20 m amounted to 99% on average. Next the BOD5 decreasing effectiveness for
the layer of the thickness of 0.10 m fluctuated between 86% and 88% and amounted
to 87% on average, however for layer of the thickness of 0.20 m fluctuated between
88% and 90% and amounted to 87% on average.
The total phosphorus in the all period of researches was being removed, because
in the raw sewages it occurred in small amounts. The total phosphorus removing
effectiveness for the layer of the thickness of 0.10 m fluctuated between 52% and
74% and amounted to 63% on average, however for the layer of the thickness of
0.20 m fluctuated between 37% and 59% and amounted to 48% on average. The
lower effectiveness of the total phosphorus removing for assisting layer of the
thickness of 0.20 m is caused by the bed saturation with phosphorus.
The analysed indicators: solid suspensions, BOD5 and COD fulfilled the Polish
recommendations concerning the introducing of cleaned sewage into ground
ENVIRONMENT] 2006).
29
Forms of nitrogen in sewage
The carried-out research on the effectiveness of sewage treatment in the ground
bed made only of gravel (KALENIK, AMBROZIAK 2005) shows (Table 3) that the
total nitrogen and ammonia nitrogen was decreasing in satisfactory degree. The
nitrate nitrogen quantity grew several dozen times, however the nitrite nitrogen in
the cleaned sewage occurred in trace amounts. The removing effectiveness of the
total nitrogen amounted to 26% on average and this of ammonia nitrogen - 99%.
Table 3
Forms of nitrogen in raw sewage and clened sewage in ground bed of gravel (mean values)
(KALENIK, AMBROZIAK 2005)
Indicators
Total nitrogen
Ammonia
nitrogen
Nitrate nitrogen
Nitrite nitrogen
Raw
sewage
Unit
-3
[mg N⋅dm ]
[mg N-NH4⋅dm-3]
-3
[mg N-NO3⋅dm ]
[mg N-NO2⋅dm-3]
Cleaned sewage
31.90
9 week
23.30
10 week 11 week 12 week
23.30
23.30
24.3
10.70
0.00
0.06
0.00
0.005
0.03
0.0019
5.20
0.0035
4.80
0.0027
5.10
0.0081
4.50
0.0022
Analysis of results presented in the Table 4 allowed to state that after filtering
the raw sewage through gravel with dolomite assisting layer, the content of the total
nitrogen, ammonia nitrogen and nitrite nitrogen decreased in the cleaned sewage.
However, increase of the nitrate nitrogen occurred.
Table 4
Forms of nitrogen in raw sewage and cleaned sewage in ground bed of gravel with dolomite
assisting layer (mean values)
Indicators
Total nitrogen
Ammonia
nitrogen
Nitrate nitrogen
Nitrite nitrogen
Unit
[mg N⋅dm-3]
[mg NNH4⋅dm-3]
[mg NNO3⋅dm-3]
[mg NNO2⋅dm-3]
31.8
Cleaned sewage
Assisting layer thickness
0,10 m
0,20 m
7
8
9
5
6
7
week week week week week week
20.4
20.5
20.9
18.0
17.6
17.4
4.2
0.150
0.150
0.120
0.050
0.110
0.095
0.3
64.38
61.52
40.54
44.38
41.85
41.78
0.0082
0.002
0.002
0.001
0.0
0.0
0.0
Raw
sewage
The total nitrogen removing effectiveness for the layer of the thickness of 0.10
m fluctuated between 34% and 36% and amounted to 35% on average, however for
the layer of the thickness of 0.20 m fluctuated between 43% and 45% and amounted
30
to 44% on average. Next, the ammonia nitrogen removing effectiveness for the layer
of the thickness of 0.10 m fluctuated between 96% to 97% and amounted to 96,5%
on average, however for the layer of the thickness of 0.20 m fluctuated between 97%
to 99% and amounted to 98% on average. The nitrite nitrogen in the cleaned sewage
for both layers occurred in trace amounts, however the nitrate nitrogen amount grew
several dozen times. The big quantity of the nitrate nitrogen in the cleaned sewage
proves that in the ground bed a nitrification occurs.
Summary
In the ground bed made only of gravel, the indicators BOD5 and COD are being
removed in appropriate amount according to the current recommendations
ENVIRONMENT] 2006). However, the solid suspensions are not being removed in
satisfactory degree and do not fulfill the current recommendations. Only after using
the dolomite assisting layer in the gravel ground bed the solid suspensions have been
totally removed from the cleaned sewage.
In the carried-out experiment it is possible to state that the variability of the
dolomite assisting layer thickness between 0.10 and 0.20 m has a small influence on
the sewage cleaning effectiveness. However, one should emphasize that a better
cleaning effectiveness was obtained for the dolomite layer of the thickness of 0.20
m. Filtering the sewage through the ground bed with assisting layer of the thickness
of 0.20 m, comparing to the layer of the thickness of 0.10 m, resulted for the cleaned
sewage in increasing of the BOD5 reduction effectiveness on 1% on average, COD on 2 %, total nitrogen - on 9 %, ammonia nitrogen - on 1.5 %. However, the total
phosphorus reduction effectiveness decreased on 15%. As the obtained differences
are lower than the accuracy of determination of the individual indicators, it is
possible to acknowledge them as little significant. Hence, the application of the
assisting layer of the thickness of 0.10 m is sufficient. The very good effectiveness
of removing solid suspensions from the raw sewage in gravel with dolomite layer
can be the reason of fast silting-up of the subsurface sewage disposal field. Thus, the
septic tank should be designed in such way that there could be removed as big
amount of suspensions from sewage as possible.
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naleĪy spełniü przy wprowadzaniu Ğcieków do wód lub do ziemi oraz w sprawie substancji
szczególnie szkodliwych dla Ğrodowiska wodnego. Dz. U. Nr 137, poz. 984.
SCHWAGER A., BOLLER M. 1997: Transport phenomena in intermitted filters. Water Science
and Technology, vol. 35, no. 6: 13-20.
SIEMIENIEC A., KRZANOWSKI S. 2001: Ocena skutecznoĞci oczyszczania Ğcieków przez filtry
gruntowe w warunkach terenowych. VII Ogólnopolskie Sympozjum Szkoleniowe.
„Projektowanie i eksploatacja przydomowych oczyszczalni cieków”, 28 II-1 III. PoznaKiekrz, Eko-Tech., 77-89.
SROKA Z., KALENIK M. 1999: Prognozowanie zmian poziomu zwierciadła wody gruntowej
pod systemami podziemnego rozsączania Ğcieków. Konferencja Naukowa. Przyrodnicze i
Techniczne Problemy Ochrony i Kształtowania rodowiska Rolniczego i Lenego.
Roczniki Akademii Rolniczej w Poznaniu. Melioracje i Inynieria rodowiska, z. 20,
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cz 2: 359-371.
TABERNACKI J., HEIDRICH Z., SIKORSKI M., KUCZEWSKI K., ŁOMOTOWSKI J., JASISKI P.,
LIPOWSKI K. 1990. Album wzorcowych rozwiązaĔ odprowadzania i unieszkodliwiania
Ğcieków bytowo-gospodarczych z wiejskich gospodarstw zagrodowych. IMUZ Falenty: pp
68.
VAN CUYK S., SIEGRIST R., LOGAN A., MASSON S., FISHER E., FIGUEROA L. 2001. Hydraulic
and purification behaviors and their interaction during wastewater treatment in soil
infiltration systems. Water Research, 35 (4): 953-964.
WILHELM S. R., SCHIFF S. L., ROBERTSON W. D. 1994: Chemical fate and transport in
domestic septic system unsaturated and saturated zone geochemistry. Environmental
Toxicology and Chemistry, vol. 13, no. 2: 193-203.
Marek Kalenik, Maciej CieĞluk
Division of Water Supply and Sewage Systems
Department of Civil Engineering and Geodesy
Warsaw Agricultural University
ul. Nowoursynowska 159, 02-776 Warsaw, POLAND
e-mail: [email protected]
33
34
CHAPTER III
Józef Koc1, Paweł Skonieczek2, Marcin Duda1
POTENTIAL FOR SEWAGE WATER PURIFICATION IN
AN AQUEOUS ENVIRONMENT BY A CONSTRUCTED
WETLAND
Introduction
Modern development and improvement of the quality of life has been
accompanied by increasing populations outside big cities. Formerly rural areas are
frequently assuming non-agricultural functions such as settlement, services,
logistics, industry and recreation. This has resulted in an increase in the amount
of waste matter and acceleration of its circulation; as sewage and waste it also
becomes a threat to the environment. Many years without sewerage systems and
sewage treatment plants or systems of waste collection and disposal, with waste
management systems based on makeshift septic tanks and rubbish dumps, has
resulted in an accumulation of waste in the environment, mainly in the soil and
subsurface water. Contaminants have been spreading and contributing to increased
fertility of underground waters and small streams. Such streams are also fed with
discharges from sewage treatment plants, which has negative consequence because
systems of lower treatment effectiveness and higher biogenic element concentrations
in post-treatment waters are acceptable due to a low building density and the small
amounts of sewage for treatment (Ministry Regulation of 29.11.2002). As a result,
waters are polluted by the following types of contaminations (KOC 1994):
− from natural sources
• products of rock erosion;
• products of organic matter transformations in soil;
• secretions and excrements of living organisms;
• substances contained in atmospheric precipitation;
− from agricultural sources
• components or mineral and organic fertilizers which have not been taken up
by plants;
• remains of pesticides which have not been decomposed;
• remains of fuels and products of their combustion;
• animal excrements as sewage from animal breeding farms;
• runoff from fodder silos and fodder storage houses;
• excrements of animals which are permanently grazing, washed down by
rain;
35
− from non-agricultural sources
• rainwater runoffs from roads and hardened areas;
• contaminations of transport-related origin from transport routes;
• runoff from legal and illegal rubbish dumps;
• products of fuel combustion;
• sewage treatment plant discharges;
• inflow of contaminated rainwater from urban areas;
In this case, water carries elements of natural matter circulation combined with
contaminants from the environment and those produced by day-to-day activities
of rural area inhabitants. Sewage discharges from treatment plants often cause the
quality of water in small streams to deteriorate due to higher acceptable limits
of contaminations and biogenic element concentrations as well as a higher
contribution of sewage discharge in flowing water than is the case with large urban
agglomerations, where discharges from treatment plants flow off to bigger rivers
(TUCHOLSKI et al. 2007). Therefore, rural area sanitation schemes are being carried
out to prevent increases in environment pollution, but their effects are below
expectations.
The unsatisfactory rate of improvement of water quality puts Poland at odds
with the commitments it has made under international agreements (GUS, 2009). By
ratifying the Helsinki Convention, Poland committed itself to take steps to arrest the
growth fertility of the Baltic Sea by improving the quality of inflowing rivers (which
are largely dependent on the quality of their tributaries), including the elementary
watersheds in the upper parts of the drainage basins.
Consequently, a crucial task is to improve the water quality in small streams. It
is a difficult problem and closely dependent on the natural factors which govern the
matter and energy circulation in the environment which has been considerably
transformed throughout the country.
Natural factors affecting water quality in rural areas
The transformation of primeval landscape into agriculturally-utilized areas has
affected hydrographic conditions by shortening and simplifying water circulation
cycles. The gradual increase in field and pasture areas at the expense of forests has
brought about a reduction of soil retention and deterioration of infiltration
conditions, while at the same time facilitating surface runoff of rainwater. These
processes have resulted in negative transformations of natural biogeochemical
cycles (STACHOWICZ 1997). The situation has been aggravated by drainage run-off,
which accelerates substance outflow from the soil (BOROWIEC, ZABŁOCKI 1990,
KOC, SZYMCZYK 2001).
Agricultural contaminates are primarily area-related. They include mineral and
organic fertilizers applied in field cultivation as well as erosion-related surface runoff from rural areas (PAWLIK-DOBROWOLSKI 1990; WODNO-BILANSOWE... 1998).
Agricultural contaminates also originate in dispersed point sources, such as
homesteads (residential buildings and septic tanks) and objects related to animal
production (livestock buildings, manure heaps, tanks of animal slurry and liquid
manure, silos and silage heaps, etc.). The intensity of contaminate inflow to water
36
from the sources scattered around the area depends on various factors, including:
land use, size of the area and its inclination, soil type, vegetation, population density.
Substance migration from soil with water is largely affected by the type of
agricultural activity, including: type of cultivated plants, coverage of soil with
vegetation, level and manner of organic and mineral fertilization, mass of postharvest remains decomposed in autumn and in winter and livestock density
(PAWLIK-DOBROWOLSKI 1990; GIERCUSZKIEWICZ-BAJTLIK 1990; WODNOBILANSOWE... 1998). An important role in the process is also played by hydrological
and soil-related factors, i.e. the amount and distribution of rainfall and spring thaw
water, occurrence of torrential rain, intensity of surface run-off, type of soil, organic
matter content in soil and air-water relations (WODNO-BILANSOWE... 1998).
It is of fundamental importance for the quality of water in elementary drainage
basins how many nutrients are discharged by small sewage treatment plants and how
many nutrients are supplied to rural areas in order to boost agricultural production.
This can be controlled by calculating a balance of biogenic elements on different
scales: for the country, the region, the drainage basin and by monitoring water
condition.
According to DYMACZEWSKI et al. (1997) and HARTMANN (1999), specific
saprobic zones, with different types of organisms dominating, form in the stream
flow below the sewage discharge inflow, depending on the stream load with
impurities. The following saprobic zones are identified: a deoxidation and
degradation zone, situated immediately below the discharge site of improperly
treated or untreated sewage (polysorbic zone); intermediate zones, where aerobic
conditions are being restored, (- and -mesosaprobic zones); a water restoration
zone or unpolluted zone (oligosaprobic).
Impurities flowing into the streams and water bodies are reduced, but this is the
case only until the threshold of water ability to self-purify is exceeded. The process
results in the formation of mineral forms of biogenic elements. Those which flow in
from various sources form the base of primary production, whose size exceeds the
consumptive capabilities of higher links in the food chain. Excess primary
production is decomposed by destruents, often under a shortage of oxygen, with
toxic substances excreted and non-mineralized organic substance accumulated. In
effect, biogenic inflow in excess of absorptive capabilities in hydrobiocenoses
results in secondary water contamination, especially in water bodies with low water
exchange indices in a lake-river system.
Degradation of aqueous ecosystems results in upsetting the balance between the
amount of contaminations of human origin and the capability of their neutralization
in the environment, i.e. of self-purification. This is the reason for increasing
environment pollution, excessive eutrophication and water productivity.
The decrease in pollution growth rate and reduction of pollution in small streams
is insufficient given the necessity to achieve a significant improvement in water
quality and lowering of the concentration of contaminations and biogenic elements
carried with water to the Baltic Sea. More effective sewage management should
include highly effective biological processes which stimulate biological
transformations and result in water purification (WÓJCIK 1993, STRUTYSKI 1997).
Each undertaking aimed at reducing the load of impurities in surface waters is
worth considering, especially if it requires small investment outlays. It is of great
37
importance in reducing the migration of nutrients to waters to properly manage the
banks of streams and coasts of water bodies and to create buffer zones (SÖDERGREN
1993; HAYCOCK et al. 1996; KOC et al. 2001).
Seeking new methods of effective utilization of existing objects in environment
protection involves studies of constructed wetlands. Characteristic features of such
objects include shallow depth, high heating/cooling rate and intense mixing –factors
which play an important role in matter transformations. The intensity of biogenic
elements absorption in a water body depends mainly on the season of the year, the
size of inflowing loads of impurities and the retention time. Studies have been
carried out in order to determine the reduction of concentrations and loads
of impurities in a stream carrying water from a forest and agricultural area which
was additionally contaminated with sewage from a sewage treatment plant, after
flowing through a constructed wetland. Rural area sanitation schemes provide the
possibility of channelling off sewage to larger, more effective treatment plants to
reduce impurity loads in this type of object. It is also important to determine how
a constructed wetland and the sludge accumulated in it will behave both during and
after a period of sewage discharge to the stream.
Fundamentals of constructed wetlands operating as biogeochemical
barriers in the environment
When seeking methods of surface water protection against eutrophication,
a possibility was suggested of using small constructed wetlands as impurity
eliminators (CZAMARA 2002). The amount and temporal variability of impurities
inflowing to water bodies depends on:
• use of watersheds above a constructed wetland;
• seasons of the year and the related periods of vegetation and agricultural
procedures;
• torrential rains and erosion-causing run-offs during intensive thaws;
• intensity of run-off from the drainage systems which supply water bodies;
• sewage inflow from settlements and loads of impurities carried by it.
Constructed wetlands should be situated below the existing inflows carrying
increased loads of impurities.
A newly constructed wetland in a stream valley results in the following:
• changes in matter and energy circulation within the system;
• creation of a new eco-system with a clear tendency to self-organize which is
intended to achieve stable equilibrium;
• change of intensity of factors which regulate the matter and energy
circulation rate.
The presence of a constructed wetland along the stream flow route reduces the
water flow rate, which results in sedimentation of impurities carried with water
(PARZONKA 1991, MADEYSKI, TARNAWSKI 2004). A decrease in the water flow rate
in the initial part of the wetland results in separation of sedimenting particles by size
and specific weight. The largest mineral particles sediment first, followed by smaller
and smaller ones in consecutive sections of the stream. Organic-mineral and organic
38
(colloidal) particles do not sediment until they reach the middle part of the wetland
and before the weir, where the flow rate is the lowest. The substances dissolved in
water may form insoluble deposits of calcium sulphates and phosphates and
aluminum and iron phosphates.
Biogenic substances dissolved in water are taken up by plants and contribute to
the production of organic mass (primary production) in the wetland. After being
consumed, plants are returned to water as excrements of herbivores and predators.
A special role is attributed to macrophytes, which act as filters of suspensions in the
littoral zone. CHUDYBA and KALWASISKI (1998) claim that retaining inflowing
matter in the littoral zone is closely connected with the presence of higher plants.
Their morphological structure, combined with appropriate growth density, turns
them into mechanical filters. This effect applies mainly to dispersed impurities with
high levels of suspension. Settlement of suspensions on plants is closely related to
decreased water flow in the littoral zone. The largest amount of suspensions is
retained by spiked water-milfoil (Myriophyllum spicatum L.), and the smallest is
retained by aquatic moss (Fontinalis antypyretica). Higher plants also retain oil and
petrochemical impurities. Discussion of the role of littoral as a filter must not be
confined to mechanical action of macrophytes. Higher vegetation can take up
organic and mineral substances and use them in metabolism, or accumulate them in
cells. Many authors claim that macrophytes are a fundamental factor which affects
water quality in water bodies, and the more of them there are, the more capable the
water bodies are of self-purifying (SZYPEREK 2003). As the eutrophication and
primary production increases, the food chain breaks down due to biomass excess,
reduced amount of light reaching the depth of water as well as a shortage of oxygen
in the water during periods when no photosynthesis takes place (night, winter) and
its consumption for plant respiration. The living conditions of herbivores and
predators deteriorate. Unconsumed plants settle and become part of bottom deposits.
Sedimenting substances carried with water which are produced in the constructed
wetland form a deposit layer, which is initially resuspended and decomposed. As
time passes, the deeper layers of more intensely mineralized sediments stabilize and
fresh layers are superimposed on them. Water movement in the constructed wetland
causes deposits to accumulate in its deepest places where deposits containing more
organic matter settle due to slower sedimentation in the flowing water and its easy
transport with water currents. An excessively high content of biogenic substances in
water contributes to development of vegetation in the littoral and sub-littoral zone
and in shallow waters. Due to intensive eutrophication, competition is won by plants
with high nutritional needs and those which produce high levels of biomass.
Impurities are absorbed and substances sediment in the overgrown zone of the
constructed wetland. Water and macrophyte movement, caused mainly by wind,
disturbs the sediments and brings about their mineralization and transfers them to
deeper layers of the wetland (SIWEK et al. 2009). However, intense growth of
phytoplankton in over-fertile wetlands reduces light access to the water depths and
brings about extinction of submerged and partly submerged plants. Oxygen shortage
in water may occur as it is consumed in respiration.
Impurities carried with water are removed by sedimentation, precipitation
of insoluble compounds, coagulation of dispersed colloids and biosorption. This
produces sediments which are transferred, resuspended and decomposed – resulting
39
in the release of accumulated substances. The initial period of the constructed
wetland existence is dominated by the processes of water purification and its
renewal. This is followed by a period of equilibrium of the processes of impurity
trapping and release from sediments. The wetland then stops playing the role of
a biogeochemical barrier. This is preceded by a transition period when the wetland
plays its role only during the vegetation period. The biological, physicochemical and
chemical processes mentioned above, and especially the equilibrium between
impurity absorption and release, are considerably affected by weather conditions,
including temperature, wind and light, as well as the amount of inflowing impurity
loads. Therefore, after the initial period of the wetland existence, when water is
purified all year round, it is highly effective in water purification during the
vegetation period, but its effectiveness is reduced outside it, especially in winter.
When the phase begins with water being purified in summer and contaminated in
winter due to the domination of impurity release from sediments, it is a sign that the
water body has to be rehabilitated by sediment removal or stabilization.
The equilibrium between impurity absorption and release depends on its
concentration in flowing water. If impurity inflow to a constructed wetland (which
purifies strongly contaminated waters) ceases, e.g. because sewage treatment plant
effluent is channeled off outside the drainage basin, such a wetland may become
a source of impurities due to a shift in the equilibrium towards impurity release from
sediments. The wetland is supplied with biogenic substances and impurities from
within (BORÓWKA 2007). Hence, after a period when the wetland is used as
a biogeochemical barrier for the impurities carried with water, its control (and
frequent rehabilitation) is necessary.
The high effectiveness of water purification by macrophytes has encouraged the
use of artificial systems mimicking marsh ecosystems in the purification processes
of sewage and sewage-contaminated waters. Natural treatment facilities make use
of the ability of the soil and vegetation of marshy ecosystems to retain and
decompose impurities present in waters and sewage. Constructed wetlands which
mimic marshy ecosystems are more effective in purification than natural water
bodies (COVENEY 2002).
Using constructed wetlands to reduce impurity loads inflow leads to their
degradation. Water bodies situated in urban drainage basins and those in the vicinity
of farms are particularly susceptible to degradation. Surface run-off from built-up
areas carries impurities at high concentrations, including nitrogen and phosphorus.
It is important to determine the threshold values and impurity loads absorbable by
these systems. They could provide a basis for calculating the maximum acceptable
anthropogenic loads in systems of small streams and constructed wetlands on their
way. Such systems usually receive effluent from rural sewage treatment plants.
Failure to find an accurate solution to the problem results in progressive degradation
of the rural areas where small streams are particularly valuable, performing many
functions that no other systems can perform.
Owing to their features, small water bodies are specific ecosystems which are
significantly different from each other and different from other environments in our
climatic zone. They can perform several valuable and complex functions. These
include: biocenotic, hydrologic, sozologic, landscape-related, educational,
economic, leisure-related (KOC et al. 2001). Consequently, their use in order to
40
improve water quality has to be subordinated to its broader biocenotic role in the
environment. It seems that the complicated relations that govern them have not been
sufficiently explored and require new research into their operation, natural
importance and transformation.
Studies of the operation of a constructed wetland situated in a stream
flow
The issue of the effectiveness of a constructed wetland as a filter of impurities
carried with the stream water is illustrated by a study conducted in the Olsztyn Lake
District. The operation of a constructed wetland situated on a stream carrying water
contaminated with effluent from a sewage treatment plant was examined in a small
closed-circulation object, typical of rural areas – the drainage basin of the stream of
Szbruk, whose waters flow into Lake Wulpiskie. The stream is 5.1 km long and
its drainage basin area is 13.2 km²; it is an agricultural area with human settlements,
forest accounts for 30% of its area. There are 630 people living there.
The study determined the relationship between the environment components and
the factors affecting the operation of the stream-constructed wetland system,
including especially: hydrological relationships, variability of water and sewage
composition caused by various factors (intensified anthropo-pressure, season of the
year, topographic and weather conditions), as well as the effect of various ecosystem
parts (water, bottoms, vegetation) on self-purification of flowing water.
In the backwater area, the water of the stream of Szbruk may flow into the pond
or through the surrounding ditch which is a continuation of the stream around the
pond, or divided into the outflow into the pond and the surrounding ditch.
Agricultural area runoff flows into the main stream, to the surrounding ditch and
directly to the pond. The outflow from the pond and from the surrounding ditch
below the pond join and flow on as the Szbruk stream to Lake Wulpiskie.
A constructed wetland, where fish used to be bred, is now a water body which
intercepts point and area impurities flowing in with the stream, which collects water
from the drainage basin. This creates a peculiar aspect of the wetland as a specific
biogeochemical barrier for biogenic substances inflowing directly to Lake
Wulpiskie. The wetland acts as a biofilter – it intercepts and accumulates biogenic
substances, thereby protecting the lake waters, and evolves as its eutrophication
level, biocenosis and bottom deposits change.
Evaluation of the water quality was based on the physicochemical analyses in 4
cross-sections of the water in the Szbruk stream on two tributaries flowing in from
the drainage basin and on the outflow from the constructed wetland (Fig. 1). During
the period covered by the study, samples of water were taken from the stream and
from the constructed wetland every month.
The studies were conducted during a period when two sewage treatment plants –
no. 1 in Unieszewo and no. 2 in Szbruk – were in operation, when their outflows
were channeled off to the Szbruk stream (2002-2003). This was also when a
sewerage system was operating in the area and the sewage from the settlements was
channeled off to a treatment plant situated outside the drainage basin (2006-2007).
41
Fig. 1 Positions of measurement sites in the Szbruk stream drainage basin
Water flow was measured with an electromagnetic flowmeter manufactured by
Valeport, model 801 (UK).
Unit water outflow at different measurement sites varied throughout the study
period. The unit inflows determined in the study were affected by the size of the
drainage area above the flow measurement site.
The water flow up to the place where it was divided into the surrounding ditch
and the constructed wetland increased with the size of the drainage basin, which
is consistent with the fundamental relationship between the increase of waterflow in
a stream with the increasing area of its drainage basin. Such streams are referred to
42
as draining streams (BYCZKOWSKI 1996). Different situation was in the case of the
flow in the surrounding ditch supplied from the agricultural drainage basin and
in the outflow from the constructed wetland. This was caused by evaporation from
the water surface and by transpiration of emergent vegetation. The highest (extreme)
fluctuations of water inflow – from no inflow to the highest observed in the system –
were recorded in the inflow to the drained areas. The constructed wetland may play
the role of outflow regulator because its structure reduces rapid outflow fluctuations;
it can also receive large amounts of water and release it after a period of delay
(Fig. 2).
50
dm-3.s-1
40
30
20
10
0
1
2
3
4
5
6
7
14,2
26,8
5,3
3,6
15,4
3,1
18,5
min
7,2
12,2
0,0
0,2
0,0
0,0
2,1
max
33,1
92,1
40,2
18,6
94,8
24,1
103,9
average
Fig. 2 Water flow at specific flow measurement sites [dm-3. s-1], description of flow
measurement sites as per Table 1
The effect of the constructed wetland on impurity concentrations and
loads in flowing waters
It is very difficult to monitor the processes of impurity loads flowing through
water bodies in natural conditions due to their complexity and temporal variability.
The processes’ dynamics are associated with the variability of factors which affect
the supply of external impurity loads and changes that take place inside the water
bodies.
In general, the lowest values of the parameters were determined in the upper
parts of the Szbruk stream (measurement site no. 1) where it carries waters from
the agricultural and forest drainage basin (Table1). The water is regarded as good
quality water [REGULATION …]. The concentration of impurities grew with the
increase in the drainage basin area and was especially high during the period when
the treatment plants were in operation.
Discharge from the Unieszewo treatment plant, which consisted of three
filtration plots and two tanks, supplied 7.8 m3 of pre-treated sewage a day, which is
equivalent to 0.09 dm-3.s-1. These amounts did not significantly increase the flow.
The Szbruk treatment plant, which worked periodically with two stoppages for
clearing and sewage discharge to the stream, daily discharged 50 m3 of pre-treated
sewage. The sewage was discharged twice a day for 30 minutes. First, it reached the
intermediate ditch, which considerably prolonged the time of inflow of pre-treated
sewage to the Szbruk stream (measurement site no. 2). This prevented any
43
significant flow increase in the stream and made it dilute, which favours selfpurification. The total daily inflow of sewage from both treatment plants to the
Szbruk stream was 58 m3. Sewage contributed 2.5% to the temporary flow and
33% at the time of discharge (Fig. 2).
Effluent from the first treatment plant lowered the water quality, which
manifested itself in an increase in the analyzed parameter values. However, the
increase was small compared to the impact of effluent from treatment plant no.2;
this allowed the stream to retain its ability to self-purify. It was only after sewage
was discharged from treatment plant no. 2 (in Szbruk) that water quality
dramatically deteriorated, with a resulting oxygen deficit (KOC ET AL. 2004). The
tested parameters values increased, and those of ChODCr, Ntot, Ptot, K, Na, Cl even
multiplied during the sewage discharge. The water quality in the stream deteriorated
at that time and was classed as category III (medium quality water).
Water contaminated with sewage, self-purified while flowing through the
constructed wetland situated below (Table 1). Ash content was reduced by 28%,
ChODCr by 40%, conductivity by 45%, total nitrogen by 88%, total phosphorus by
84%, potassium by 68%, magnesium by 7%, sodium by 76%, chlorine by 59% and
sulphates by 61%, calcium concentration increased by 33%. However, the quality
improvement in the part of water that flowed through the surrounding ditch was
much smaller. The dry residue content decreased by 13%, ash content by 8%, ChOD
by 51%, conductivity by 30%, total nitrogen by 70%, total phosphorus by 90%,
potassium by 56%, sodium by 74%, chlorine by 52, sulphates by 58%, calcium
concentration increased by 55% and that of magnesium - by 1%. The better water
quality improvement in the constructed wetland as compared to the surrounding
ditch must be attributed to slower water flow, which ensured more favourable
conditions for biological and physicochemical processes. A decrease in the
concentration of water impurities during the water flow through their wetland is not
fully equivalent to the actual effect of the process due to a significant reduction
in water volume as a result of evaporation (Fig. 2).
Only after flowing through the constructed wetland, where its oxygenation
improved and suspension sedimentation and mineral substances phytosorption took
place, did the tested parameters return to the values from before the sewage
discharge from the treatment plant. The water quality can be regarded as restored.
During the sewage discharge from the treatment plant in Szbruk, the values
of all the parameters at the outflow to the lake, except for Ca and Mg, were lower
than at the site before the inflow to the constructed wetland, despite an increase in
the drainage basin area and inflow of impurities. The constructed wetland not only
reduced the load of impurities discharged from the treatment plant, but also that
inflowing from the drainage basin. The nutrients present in the contaminated water
(and subsequently flowing through the constructed wetland of considerable area –
24.8 ha) were used for primary and secondary production in those ecosystems and
accumulated in their bottom deposits (ALLAN 1998).
44
Table 1
Concentration of substances dissolved in the water of the stream of Szbruk, carrying water from agricultural areas and from sewage treatment plants (mgdm-3)
Measurement site
Dry residue
Ash
ChOD
Conducti
vity
Ntot
Ptot
Potassiu
m
Calcium
Magnes
ium
Sodium
Chlorin
e
Sulphates
Forest runoff (1)
216*
192-272**
137
102-164
14.6
8.4-26.0
241
155-281
1.73
1.02-2.81
0.19
0.11-0.31
1.0
0.6-1.7
43.4
40.8-46.4
5.5
3.7-6.7
4.3
3.6-4.8
7
6-8
52.8
23.9-117.8
±23***
±20
±5.8
±43
±0.57
±0.069
±0.3
±1.7
±0.9
±0.3
±1
±30.71
350
220-948
215
140-532
23.9
8.4-56.8
360
291-555
4.51
1.05-19.88
0.565
0.22-2.84
3.2
1.0-8.4
53.2
40.8-69.6
8.2
4.5-17.6
8.0
4.4-23.5
9
7-12
57.4
33.4-134.1
±207
±112
±12.7
±74
±0.76
±2.4
±8.4
±3.6
±5.2
±2
±31.6
420
312-784
261
128-496
67.6
21.6-136
675
419-1230
3.55
1.02-7.82
10.8
1.3-34.2
49.4
35.6-67.4
9.4
5.1-19.2
32.6
4.8-97.8
27
9-42
121.16
25.0-280.9
Before the wetland (2)
±138
±100
±43.2
±224
±5.63
26.25
12.6041.58
±10.65
±2.48
±8.5
±9.5
±4.0
±25.9
±12
±84.0
Before the wetland –
24 h average (2b)
365
231-923
±203
222
145-516
±110
25.4
13.3-56.7
±12.1
373
315-556
±74
5.42
2.51-20.02
±5.27
0.69
0.32-2.88
±0.74
3.5
1.3-8.5
±2.3
53.0
40.9-69.5
±8.4
8.4
4.7-17.7
±3.7
9.0
5.2-23.7
±5.1
10
8-12
±1
60.06
37.71-131.27
±29.48
Inflow through the
drain pipe to the
wetland (3)
529
212-936
±270
372
236-604
447
112-692
±191
189
144-284
43.8
23.2-59.2
±11.1
40.9
30.0-69.2
639
205-888
±267
372
264-527
7.82
2.30-23.30
±6.81
3.16
1.66-7.30
0.15
0.04-0.28
±0.095
0.36
0.17-0.67
4.0
2.2-6.9
±1.6
3.5
0.9-6.2
101.2
31.4-188.0
±56.1
65.8
36.2-103.0
14.1
4.4-26.2
±7.1
8.8
5.7-12.0
7.6
1.8-10.5
±2.7
7.9
4.8-12.0
15
5-22
±5
11
9-16
138.5
34.7-529.4
±149.5
46.8
18.1-74.2
±130
±41
±11.5
±77
±1.86
±0.15
±1.9
±21.8
±2.1
±1.9
±2
±17.7
630
344-784
±138
365
236-464
371
160-572
±131
240
136-292
42.6
25.0-56
±10.8
33.1
24.8-47.6
775
361-1042
±190
471
220-629
8.50
4.99-13.04
±2.78
3.85
1.79-6.51
0.33
0.08-1.36
±0.43
0.36
0.10-0.92
10.1
8.2-13.4
±2.0
4.8
1.3-8.4
104.2
53.1-152.0
±33.1
76.9
31.3-101.0
14.7
6.7-23.6
±5.7
9.5
4.2-14.6
11.6
4.0-15.8
±3.5
8.5
2.9-10.5
28
19-38
±6
13
9-18
122.7
64.8-414.2
±110.4
51.1
24.0-82.8
±81
±52
±6.5
±115
±1.71
±0.21
±2.2
±20.3
±3.6
±2.2
±4
±18.7
332
248-408
236
132-316
31.7
24.0-39.6
442
291-540
3.53
1.14-6.66
0.35
0.19-0.48
4.4
1.0-7.6
74.9
43.4-111.0
9.5
5.2-11.7
8.0
2.5-10.5
13
9-18
47.0
23.8-92.5
±58
±53
±4.0
±70
±1.91
±0.11
±2.1
±17.6
±1.9
±2.3
±3
±19.6
Before the wetland at
the time of sewage
discharge (2a)
45
Outflow from the
wetland (4)
Inflow through the
drain pipe to the
surrounding ditch (5)
Outflow from the
surrounding ditch (6)
Outflow to the lake (7)
* average, ** min-max, *** ±SD
45
An analysis of the impurity load carried with water showed that it was
considerably reduced after the water had flowed through the constructed wetland
(Table 2). The dry residue load decreased by 87%, ashes – by 90%, ChODCr - by
80%, total nitrogen - by 93%, total phosphorus - by 92%, potassium - by 87%,
calcium - by 85%, magnesium - by 87%, sodium - by 88%, chlorine - by 86% and
sulphates - by 91%. The reduction level of the load of impurities were found to be
lower in the surrounding ditch: 55% for dry residue, 54% for ash, 54% for ChODCr,
53% for total nitrogen, 25% for total phosphorus, 67% for potassium, 58% for
calcium, 56% for magnesium, 48% for sodium, 67% for chlorine and 59% for
sulphates. The system consisting of the surrounding ditch and the constructed wetland
brought about reduction of the dry residue by 70%, ash - by 73%, ChODCr by 67%,
total nitrogen - by 75%, total phosphorus - by 86%, potassium - by 60%, calcium - by
67%, magnesium - by 70%, sodium - by 74%, chlorine - by 61% and sulphates - by
71%.
Table 2
Load of dissolved substances in the water of the stream of Szbruk carrying impurities from
sewage treatment plants and agricultural areas (kgyear-1)
Measurement
site
Forest runoff
(1)
Before the
wetland, 24hour
average(2)
Stream
inflow to the
wetland (2)
Stream
inflow to the
surrounding
ditch (2)
Drain pipe
inflow to the
wetland (3)
Outflow
from the
wetland (4)
Drain pipe
inflow to the
surrounding
ditch (5)
Outflow
from the
surrounding
ditch (6)
Outflow to
the lake (7)
Dry
residue
Ash
ChOD
Ntot
Ptot
K
Ca
Mg
Na
Cl
Sulphates
96932
61487
6566
777
86
450
19505
2463
1937
3005
23764
309230
188061
21525
4588
584
2987
44882
7116
7655
8498
50853
222893
135554
15515
3307
421
2153
32351
5129
5518
6125
36655
86337
52507
6010
1281
163
834
12531
1987
2137
2373
14198
59377
50128
4913
878
16
450
11360
1582
853
1628
15548
36367
18486
3997
309
35
347
6434
859
767
1089
4574
106292
62594
7185
1433
55
1708
17582
2482
1961
4777
20714
176908
116355
16027
1861
172
2306
37208
4616
4138
6447
24726
192647
137153
18378
2049
201
2548
43435
5507
4647
7464
27257
46
Load of biogenic substances in bottom deposits
Bottom deposits in rivers and water bodies are a useful geomedium used in
control of the quality of surface waters in terms of the level of eutrophication and
contamination with heavy metals and harmful chemical compounds. Since the
concentrations of harmful substances in deposits are several times higher than in
water, an analysis of deposits enables detection and monitoring of changes of their
content even when the level of contamination is relatively low (URBAN et al. 1997;
BAUDO, BELTRAMI 2001). Therefore, analyses of deposits near various sites of
contamination, e.g. near the sites of sewage discharge, are important in monitoring
the level of environment pollution.
Impurities and detritus settle on a continuous basis and form bottom deposits,
which are a kind of “archive”. Deposit particles bind both biogenic substances and
heavy metals, which are among the most persistent toxic substances entering
aqueous ecosystems (BAUDO, BELTRAMI 2001; SOBCZYSKI, SIEPAK 2001). The
process of settlement of particles, formation of a deposit layer on the water bodies
bottom and consolidation of such a layer depend on a number of factors (Parzonka
1991), which can be divided into geomorphologic, hydraulic, hydrodynamic and
exploitational factors. The processes are greatly affected by the particle features,
their accumulation, organic matter content and the presence of soluble salts in water.
The bottom deposit sampling sites are shown in Fig. 3. The bottom deposits
varied in terms of thickness and physicochemical properties (Table 3, 4). The
highest thickness was recorded for the cores taken in the middle of the constructed
wetland – up to 20 cm at a water depth of 1.51 m, whereas the lowest thickness was
recorded in the stream flow. In fact, there was no typical layer of bottom deposits as
is usually the case with water bodies. The stream bottom is covered with a coarse
material, as finer particles are washed away, especially with increased flow, and are
transported along the stream. They settle when the flow rate decreases or when they
meet with the vegetation resistance or when flowing through the constructed
wetland (MADEYSKI, TARNAWSKI 2004).
Due to the small thickness and low variability in terms of the colour and
structure of the bottom deposits from the stream and the constructed wetland
(probably due to their age), the cores of those deposits were not divided into layers.
The pH value of the sediments in the research site ranged from 4.98 in KCL
(5.34 in H2O) to 7.66 (7.45 in H2O). The lowest values were found at the first line in
the wetland where contaminated water flows in (Table 3). The pH value was
recorded in the 10-20 cm layer of the water body (Table 4). The lowest
concentration of carbonates was recorded in the initial sections of the wetland in the
10-20 cm layer (0.13%) whereas the highest (67.89%) was in the final section of the
surrounding ditch.
The bottom deposits varied in terms of the content of biogenic compounds
(Table 5, 6).
Of the elements whose content in the bottom deposits were analyzed, calcium
dominated, which is a consequence of it flowing in from the drainage basin. In
general, the largest concentrations of Ca were recorded in the sample taken in the
final section of the surrounding ditch (measurement site no. 8), where the parameter
value was 100100 mg Ca⋅kg-1 d.m. The lowest concentrations of the element were
47
recorded below the forest runoff inflow – 858 mg Ca⋅kg-1 d.m. The chemical
compositions of the sediments can modify the substrate, which was the case with
Ca. Calcium content was the most variable, both in the vertical and horizontal
profile. An over 4-fold difference in the element concentration was recorded in the
wetland between the neighbouring measurement sites.
Fig. 3. Sites of bottom deposits sampling in the stream and in the constructed wetland
Nitrogen is another element whose content differs from that of the others;
however, it dominates mainly in the wetland, both in the 0-10 cm and in the 10-20
cm layers. This shows that the element accumulates in the pond bottom, which
considerably contributes to the protection of the next system component – the lake.
The highest concentrations of nitrogen were recorded in the sediments of the central
part of the pond, close to the depths. The values were the highest both in the surface
layer 0-10cm (15.6 g⋅kg-1 d.m.) and in the layer 10-20cm (16.0 g⋅kg-1 d.m.); this
48
shows that sediments flow and settle in the deepest places of the wetlands. It is in
the central part of the pond (measurement site 11b) that bottom deposits of up to 20
cm thick were recorded, whereas 3-cm layers occurred in the extreme lines of
sediment sampling - 10 a, c and 11 a, c. SKWIERAWSKI (2003) showed the
concentration and distribution of nitrogen to be strongly correlated with the amount
of organic matter, which indicates that the nitrogen in the sediments is mainly found
in organic compounds. The dominant position of nitrogen in bottom deposits has
been reported by SZYPEREK (2004), who recorded the highest contents of nitrogen in
tributaries, MÜLLER et al. (1998) in lakes of Central Europe (63-Switzerland, 2France, 3-Italy), and GAWROSKA (1989) in sediments of Lake Bskie.
Phosphorus content ranged from 218 mg⋅kg-1 d.m. below forest runoff inflow to
1788 mg⋅kg-1 d.m. in the initial section of the pond and at the outflow to the lake,
which – as was the case with nitrogen – is indicative of the effect of the drainage
basin on biogenic compounds depositing in sediments. That the pond was supplied
by runoff from the drainage basins is indicated by a decrease in the phosphorus
concentration with increasing distance from the inflow site. The concentration of the
element in bottom deposits may result from its intense exchange in the sedimentwater interface. The amount of phosphorus released from sediments to water may be
particularly high, especially during the vegetation period, when primary production
demand for the element is high, but also when conditions favour re-suspension and
with oxygen deficit (KAJAK 2001). Accumulated in sediments and being the main
factor in water eutrophication, phosphorus plays a double role in bottom deposits.
On the one hand, the considerable amounts of phosphorus retained in the deposits
shows that deposits are effective as a trap for phosphorus migrating in the
environment but, on the other hand, high concentrations of phosphorus may trigger
processes which take place in a water body, accelerating its turning into land
(SKWIERAWSKI 2003). Under anaerobic conditions or under such that favour
resuspension (water rippling at small depths), the amount of phosphorus released to
water may be very high, especially during the vegetation period (KAJAK 2001).
Potassium is not regarded as an element which affects the process of
eutrophication, but it is used in agriculture and may be an indicator of the intensity
of agricultural soil use in the drainage basin. Its content ranged from 996 mg⋅kg-1
d.m. in the bottom deposits below the forest drainage basin to 6308 mg⋅kg-1 d.m. in
the deposits of the central part of the pond. No significant differences were recorded
between the surface and subsurface layers.
The highest concentrations of magnesium were found in samples taken in the
central part of the pond – 5126 mg⋅kg-1 d.m. in the layer 10-20 cm. The lowest
concentration was found below the forest drainage basin 362 mg⋅kg-1 d.m.
As compared to other macroelements, the concentration of Na and S in the
analyzed samples was less varied.
Analysis of bottom deposit cores provides grounds for examination of the
abundance of selected macroelements in the deposits. They show the variability of
component concentration through the pond bottom in the cross section and the
longitudinal section, from the inflow to the outflow. It is a general tendency that
larger amounts of macroelements accumulate along the pond axis and their
accumulation depends on the deposits thickness. The concentration of elements, i.e.
Mg, Na, decreased as the distance from the supply site increased.
49
Table 3
Some physical properties and selected elements (mg⋅kg-1 d.m.) in the surface layer of the bottom deposits – 0-10 cm
Sampling site
pH in
KCl
pH in
H2O
% CaCO3
% d.m.
%
Org.
matter
N
P
K
Ca
Mg
Na
S
Cl
1
After the flows join below the forest
7.14
7.25
0.42
80.3
0.51
240
218
996
858
362
134
200
4
2
Below the Unieszewo treatment plant
7.18
7.13
0.25
74.5
2.59
-
-
-
-
-
-
-
-
3
Below the Szbruk treatment plant
7.38
7.31
12.26
75.4
3.50
1290
610
3652
59345
4824
475
900
16
4
Beginning of the surrounding ditch
6.93
6.86
3.70
66.5
3.65
-
-
-
-
-
-
-
-
5
Surrounding ditch
7.54
7.31
2.22
75.1
2.00
-
-
-
-
-
-
-
-
Nr
6
7
50
8
9
10
11
12
Surrounding ditch
7.39
7.30
0.72
70.5
3.65
1390
610
2324
24668
2050
230
1100
13
Surrounding ditch below the
agricultural drainage basin inflow
7.27
7.22
4.65
54.2
7.10
-
-
-
-
-
-
-
-
End of the surrounding ditch
After joining the wetland outflow
7.23
7.02
7.25
7.15
67.89
11.24
46.8
53.8
6.00
6.30
2060
3170
1395
1788
2656
3984
100100
52910
3377
3377
653
490
1700
1400
22
23
a
6.72
6.87
2.25
26.1
23.63
9630
1788
3320
10296
4040
475
1600
22
b
5.77
6.03
0.25
34.9
13.75
4940
1788
3320
2574
1387
341
1700
18
c
4.98
5.34
0.42
20.3
33.13
13210
1177
4980
5148
1447
326
2300
30
a
6.79
7.07
16.07
19.0
26.85
-
-
-
-
-
-
-
-
b
6.10
6.41
0.32
18.8
35.53
15550
1003
4980
2860
4703
549
2900
52
c
6.49
7.12
7.64
31.4
16.75
-
-
-
-
-
-
-
-
a
7.03
7.15
10.4
42.5
9.43
4370
610
2988
35750
2472
475
1400
22
b
6.73
6.91
6.26
20.6
25.80
11420
610
4980
25025
4824
564
2900
44
c
7.09
7.30
2.21
56.4
6.39
3070
785
2324
10010
2653
326
1000
15
Constructed wetland,
beginning
center
end
- not determined
50
Table 4
Some physical properties and selected elements (mg⋅kg-1 d.m.) in the layer of the bottom deposits from 10 to 20 cm
Nr
10
Constructed
wetland,
beginning
51
11
12
pH in KCL
pH in
H2O
% CaCO3
% d.m.
%
Org. matter
N
P
K
Ca
Mg
Na
S
Cl
a
7.08
7.37
3.82
52.7
11.67
5160
785
2988
12870
4040
371
900
14
b
6.77
6.99
0.85
56.4
7.49
2900
1177
3320
4433
1387
223
1000
9
c
5.24
5.78
0.13
37.6
23.06
10080
1003
4980
1430
4100
482
1900
22
a
6.83
7.22
38.6
33.0
16.20
-
-
-
-
-
-
-
-
b
5.43
5.93
0.32
22.9
30.77
16030
785
6308
1430
5126
549
2300
33
c
6.89
7.28
4.12
46.5
13.55
-
-
-
-
-
-
-
-
a
7.05
7.27
20.37
45.6
9.66
4240
610
3320
42900
3075
519
1500
25
b
6.79
7.27
6.79
42.4
14.18
7030
610
3320
92950
3980
564
1500
25
c
7.00
7.38
3.73
60.3
7.29
3670
610
2988
10010
3678
341
900
17
Sampling site
Constructed
wetland,
center
Constructed
wetland, end
- not determined
51
Of all the determined elements, it was calcium that had accumulated in the
bottom deposits of the pond in the largest amounts – 37525 kg (Table 5). The weight
of the deposits reached 2515 tones, containing several tones of each of the elements.
Table 5
Accumulation of selected elements in the pond bottom deposits [kg]
Layer
N
P
K
Ca
Mg
Na
S
Cl
0-5 cm
13110
1636
5669
19323
4538
644
2909
43
5-10 cm
5398
674
2334
7956
1869
265
1198
18
10-15 cm
2399
273
1330
8111
1240
149
489
7
15-20 cm
631
72
350
2135
326
39
129
2
Total
21538
2655
9683
37525
7973
1097
4725
70
The thickness of the pond bottom deposits is associated with its depth. The
largest thickness (reaching 20 cm) was found near the depths. The thickness of the
bottom deposits decreased as the pond became more and more shallow so, in effect,
the layer of 0-5 cm occupied the largest area of 11.9 ha (Fig. 3). Consequently, the
highest concentrations of the analyzed elements were determined in that layer.
One hectare of the pond bottom was covered by 101 tones of deposits,
containing: 868.5 kg N, 107.0 kg P, 390.5 kg K, 1513.1 kg Ca, 321.5 kg Mg, 44.3 kg
Na, 190.5 kg S, 2.8 kg Cl.
The effect of a constructed wetland on the quality of water after
sewage discharge had ceased
After contaminations stopped inflowing with the purified sewage, their
concentrations in the stream water significantly decreased. The total nitrogen
content decreased in the constructed wetland by 45%, that of total phosphorus - by
44 %, calcium - by 21%, magnesium content increased by 3% (Table 6). The
dynamic equilibrium between water and deposits shifted towards releasing
impurities from sediments deposited in the pond, especially in the surrounding ditch,
which accelerated mineralization when the water oxygenation was sufficient. During
the period when the impurity load in water was reduced, the impurities
concentration and their loads were found to be reduced to a lesser extent (Table 7).
Soluble substance concentrations were even found to have increased, which is
indicated by the higher electrical conductivity in outflowing water as compared to
the water before the pond. The dry matter inflowing the Szbruk stream from its
drainage basin was found to have decreased by 18%, ashes - by 6%, total nitrogen by 17%, total phosphorus - by 25%, potassium - by 25%, calcium - by 19%,
magnesium - by 16%, sodium - by 25%, chlorine - by 21% and sulphates - by 24%;
52
Table 6
Concentration of dissolved substances in the water of the stream of Szbruk carrying impurities from agricultural area, after the treatment plants
had been shut down (mgdm-3)
Measurement
site
Forest runoff
(1)
Before the
wetland (2)
Drain pipe
inflow to the
wetland (3)
53
Outflow from
the wetland
(4)
Drain pipe
inflow to the
surrounding
ditch (5)
Outflow from
the
surrounding
ditch (6)
Outflow to the
lake (7)
Dry residue
Ash
ChOD
Conductivity
Ntot
Ptot
Potassium
Calcium
Magnesium
Sodium
Chlorine
Sulphates
220*
168-272**
176
124-200
14.8
6.4-18.8
283
247-318
1.45
0.3-2.2
0.15
0.01-0.44
1.0
0.7-1.3
41.2
24.2-49.2
6.6
5.1-7.8
4.0
3.6-4.4
5.1
5.0-6.0
54.1
25.5-95.1
±45***
±28
±4.8
±26
±0.8
±0.1
±0.2
±9.3
±1.0
±0.4
±0.4
±29.4
281
208-352
197
144-244
23.4
16.4-28.0
393
299-554
3.61
0.8-7.5
0.25
0.01-0.45
1.9
0.7-2.8
58.6
43.4-72.4
8.1
5.9-9.5
6.6
4.8-8.8
9.8
8.0-12.0
69.9
31.2-113.9
±50
±32
±4.1
±88
±2.5
±0.2
±0.8
±11.3
±1.4
±0.6
±1.5
±30.4
640
502-764
589
448-596
42.9
33.7-50.8
914
854-957
8.93
0.1-19.7
0.38
0.1-1.01
7.3
5.4-9.7
104.9
81-137
14.7
11.9-16.5
9.7
7.2-13.2
21.2
13.0-38.0
53.2
29.6-76.8
±204
241
220-276
±64
174
144-240
±11.9
27.8
18.7-37.2
±143
359
302-493
±8.9
1.99
0.1-2.6
±0.3
0.14
0.01-0.45
±0.5
2.2
0.7-3.7
±23.4
46.5
34.8-57.8
±2.1
8.3
4.0-10.5
±1.7
5.5
4.8-6.3
±9.5
9.5
8.0-11.0
±36.3
43.2
12.3-107.6
±19
±36
±7.1
±51
±0.9
±0.1
±0.6
±10.7
±2.3
±0.7
±1.4
±34.0
546
523-740
518
416-596
41.0
28.7-58.2
852
600-941
6.21
0.44-14.3
0.53
0.1-1.97
3.7
3.0-4.9
108.1
86.9-131
16.4
14.3-20.2
7.7
6.0-10.5
14.3
12.0-18.0
52.3
26.5-77.8
±111
±53
±7.2
±40
±5.8
±0.5
±0.7
±19.3
±1.7
±2.5
±1.5
±33.3
397
344-556
356
292-468
38.7
27.5-54.8
692
620-758
6.02
0.8-12.5
0.41
0.01-1.17
5.5
1.0-7.4
81.0
65.2-91.4
10.4
8.8-13.0
7.2
6.3-9.2
12.0
11.0-15.0
84.5
16.6-131.2
±79
±77
±9.8
±51
±5.1
±0.3
±0.7
±10.2
±1.7
±1.2
±1.5
±45.0
292
224-400
258
204-296
31.6
25.0-42.4
509
365-720
3.90
1.1-9.7
0.37
0.01-0.43
3.5
0.9-7.1
72.5
65.2-81.0
9.4
7.9-10.8
7.0
5.6-9.2
11.5
10.0-14.0
52.7
25.9-96.12
±58
±34
±7.8
±120
±3.9
±0.2
±0.7
±5.9
±1.2
±1.3
±1.4
±25.2
* average, ** min-max, *** ±SD
53
however, ChODCr increased by 5%. This was the result of sediment re-suspension
and mineralization, as well as primary production in the pond. The study showed
that the initially achieved good effects of purification of highly contaminated waters
in the pond gradually worsen.
Passing water of lower contamination level through a constructed wetland
previously used for water purification results in reduced process effectiveness. The
decrease in the water contamination level in a ditch system shows that if clean water
is introduced in the next stage, its quality may deteriorate as a result of the object
washing-out.
Table 7
Load of dissolved substances in the water of the stream of Szbruk, carrying impurities from
an agricultural area (kgyear-1)
Measurement
site
Dry
residue
Ash
ChOD
Ntot
Ptot
K
Ca
Mg
Na
Cl
Sulphates
Forest runoff
(1)
75900
60720
5117
500
51
357
14225
2271
1380
1783
18653
Before the
wetland(2)
165143
115835
13746
2119
147
1859
34388
4735
3854
5772
41007
66057
46334
5498
848
67
744
13755
1894
1542
2309
16402
99086
69501
8247
1271
80
1115
20633
2841
2313
3463
24604
12864
12422
985
214
9
91
2595
394
184
344
1251
56472
40872
6501
466
40
449
10873
1958
1291
2223
10119
30736
25848
2046
298
25
350
5035
706
464
1016
2553
113810
104120
11043
1715
96
1278
23076
2955
2071
3420
24088
170282
144992
17545
2181
136
1726
33949
4912
3362
5643
34208
Stream
inflow to the
wetland (2)
Stream
inflow to the
surrounding
ditch (2)
Drain pipe
inflow to the
wetland (3)
Outflow
from the
wetland (4)
Drain pipe
inflow to the
surrounding
ditch (5)
Outflow
from the
surrounding
ditch (6)
Outflow to
the lake (7)
54
Table 8
Load of substances carried in the water of Szbruk stream (kgyear-1)
Load
Dry
residue
Ash
ChOD
Ntot
Ptot
K
Ca
Mg
Na
Cl
Sulphates
During the period of sewage discharge
Inflowing to
the wetland
282270
185682
20428
4185
437
2603
43711
6711
6371
7753
52203
Outflowing
from the
wetland
36367
18486
3997
309
35
347
6434
859
767
1089
4574
Difference
245903
167196
16431
3876
402
2256
37277
5852
5604
6664
47629
192629
115101
13195
2714
218
2542
30113
4469
4098
7150
34912
176908
116355
16027
1861
172
2306
37208
4616
4138
6447
24726
15721
-1254
-2832
853
46
236
-7095
-147
-40
703
10186
Inflowing to
the ditch
Outflowing
from the
surrounding
ditch
Difference
During the period without sewage discharge
Inflowing to
the wetland
78921
58756
6483
1062
76
835
16350
2288
1726
2653
17653
Outflowing
from the
wetland
56472
40872
6501
466
40
449
10873
1958
1291
2223
10119
Difference
22449
17884
-18
596
36
386
5477
330
435
430
7534
Inflowing to
the ditch
129822
95349
10293
1569
105
1465
25668
3547
2777
4479
27157
Outflowing
from the
surrounding
ditch
113810
104120
11043
1715
96
1278
23076
2955
2071
3420
24088
16012
-8771
-750
-146
9
187
2592
592
706
1059
3069
Difference
Summary
A constructed wetland situated on a small stream may also play the role
of a biogeochemical barrier for point and non-point contamination inflow in an
agricultural and forest drainage basin in a low-population area. The constructed
wetland improved the water quality up to the quality level of water outflowing from
the forest area throughout the period of its existence, both when sewage was
discharged to the stream and after the sewage discharge ceased.
Water flow through the pond resulted in a decrease in contamination and
eutrophication indexes. The dry residue content decreases by 12%, ashes - by 28%,
ChODCr – by 40%, conductivity – by 45%, total nitrogen - by 88%, total phosphorus
- by 84%, potassium - by 68%, magnesium - by 7%, sodium - by 76%, chlorine - by
55
59% and sulphates - by 61%. Except for the dry matter content, the reduction
of these parameters was considerably greater in the water flowing through the pond
than in the surrounding ditch.
Water flow through the pond resulted in reduction of dry matter content by 87%,
ashes - by 90%, ChODCr - by 80%, total nitrogen - by 93%, total phosphorus by 92%, potassium - by 87%, calcium - by 85%, magnesium - by 87%, sodium by 88%, chlorine - by 86% and sulphates - by 91%. Reduction of impurities load is
much greater than that of the concentrations, as water loss by evaporation and plant
transpiration is greater.
The effectiveness of water purification is higher in the vegetation period than
in winter, which is a confirmation of the thesis that the main role in water
purification is played by biological processes, which depend on temperature and
solar exposure.
The pond was degraded over the many years of its existence, which is indicated
by its being overgrown by nitrophilous rush vegetation and sediment accumulation
on the bottom, which are more abundant in nitrogen and phosphorus than are typical
sediments in lakes which are not loaded with sewage.
Water movement in the pond and the biological processes result in accumulation
of mainly mineral deposits at the water inflow to the pond, with the contribution of
organic matter and biogenic substances increasing in the further parts of the pond.
The highest accumulation of deposits was recorded in the depths and before the
weir.
Reduction of the pond load with impurities by diverting the sewage discharge
outside the drainage basin resulted in a change of relations between the processes
of accumulation and release of impurities due to their lower concentration in the
inflowing water; the effectiveness of impurities and biogenic compounds reduction
in the water flowing through the pond decreased.
Accumulation of deposits in the lake cove where the water from the stream
flows in is a sign that the water quality improvement is insufficient. Further quality
improvement can be achieved by passing all the stream water through the pond (but,
in that case, it would have to be larger) or by a series of ponds (“bead-on-a-string”
system) on the stream.
Constructed wetlands which act as natural sewage treatment plants should be
rehabilitated as the effectiveness of water purification decreases over time, with the
intensity dependent upon the impurity load-to-wetland volume ratio. A nonrehabilitated wetland may periodically reduce the water quality as a result
of lowering the accumulation/release ratio, which may happen during unfavourable
periods, such as a rapid flow increase in winter or water waving.
The study was financed by funds for financing scientific research in the years
2007-2010
56
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IMUZ, 27: 25-30.
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CHUDYBA H., KALWASISKI K. 1998. Samooczyszczanie wody. Nowocz. Roln. 05.06. pp. 48.
COVENEY M. F., STITES D. L., LOWE E. F., BATTOE L. E., CONROW R. 2002. Nutrient removal
from eutrophic lake water by wetland filtration. Ecological Engineering 19: 141-159.
CZAMARA W., WIATKOWSKI M. 2002. Zastosowanie zbiornika wstĊpnego w MĞciwojowie do
ochrony retencjonowanej wody. Roczniki Akademii Rolniczej w Poznaniu, CCCXLII: 4352.
GAWROSKA H., 1989. Skład chemiczny osadów dennych Jeziora BĊskiego. Acad. Agricult.
Tech. Olst. ProtectioAquarum et Piscatoria, No 17: 35-43.
GIERCUSZKIEWICZ-BAJTLIK M. 1990. Charakterystyka obszarowych Ĩródeł zanieczyszczeĔ w
Polsce. W: Zanieczyszczenia obszarowe w zlewniach rolniczych. Mater. semin. IMUZ
Falenty 26: 143-161.
GUS 2009. Ochrona rodowiska. GUS Warszawa
HAYCOCK N.E., PINAY G., BURT T.P., GOULDING K.W.T. 1996. Buffer zones: Current
concerns and future directions. W: Buffer zones: their processes and potential in water
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KAJAK Z., 2001. Hydrobiologia, limnologia. Ekosystemy wód Ğródlądowych. Wyd. Nauk.
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KOC J. 1994. ZagroĪenia Ğrodowiska rolniczego. Rodzaje Ĩródła rozmiary i skutki. Orodek
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KOC J., CYMES I., SKWIERAWSKI A., SZYPEREK U. 2001. Znaczenie ochrony małych
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KOC J., SZYMCZYK S. 2001. Wpływ uĪytkowania obszarów rolniczych na eutrofizacjĊ wód.
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KOC J., TUCHOLSKI S., SKONIECZEK P. 2004. Znaczenie zbiornika wstĊpnego w ochronie
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1
Józef Koc, 1Marcin Duda
Department of Land Improvement and Environmental Management
pl. Łódzki 2, 10-718 Olsztyn, POLAND
2
Paweł Skonieczek
Department of Environmental Protection Engineering
ul. Prawocheskiego 1, 10-957 Olsztyn, POLAND
58
CHAPTER IV
Justyna Koc-Jurczyk
TREATMENT TECHNOLOGIES OF MUNICIPAL WASTE
LANDFILL LEACHATES
Introduction
For many years now, one of the most important problems in the area
of environment protection consists in the waste management defined as the whole
of activities aimed at reduction of quantity, development of effective methods of
utilization, and neutralization of waste.
Waste utilization by means of landfilling creates a number of hazards to the
environment. Leachates can be counted among them.
The main source of leachates consists in penetration of surface and underground
waters as well as external water. Water from a landfill is emitted in the form of
vapor and as a component of the landfill gas. Water not absorbed by the waste
generates leachates that accumulate at landfill bottom or fill empty spaces at
different landfill levels in the form of the so-called suspended water (OBRZUT 1997).
The leachates are created when the damp content in the landfill bed exceeds its
retention capacity defined as the maximum water quantity that can be retained in
porous material of the landfill (EL-FADEL et al. 2002).
Leachate composition reflects microbiological activity of a landfill. Leachates
represent a complex and variable mixture of organic, inorganic and microbiological
substances and suspensions of solid substances in water. Leachate composition
depends also on waste type, e.g. plaster or gypsum can be transformed by anaerobic
bacteria into sulfides. However, the most important factor affecting leachate
composition is the landfill age (OBRZUT 1997, EL-FADEL ET AL. 2002, OZKAYA
2005).
Municipal waste and some industrial waste, neutralized by means of landfilling,
contain mixtures of hazardous substances of various types. Their sources include
unwanted products disposed to landfill sites after being used in households or
industrial enterprises. As a result of long stay in a landfill bed, they are subject to
partial degradation. Consequently, a number of potentially hazardous substances
occur in leachates. Many of them demonstrate xenobiotic nature.According to
SLACK et al. (2005), more than 200 organic substances were identified up to date in
leachates from municipal waste landfill sites, including 35 substances considered
potentially hazardous. On the other hand, in ground waters existing in areas
surrounding waste landfills, more than 1000 substances of different types were
59
detected. This means that transformation and/or partial degradation of hazardous
substances of various types lead to release of intermediate products and can result in
contamination of ground waters.
Chemical transformations in leachates in waste dumping period
KURNIAWAN et al. (2006), and earlier KANG et al. (2002), have classified
landfill sites, depending on period of their operation, as: young; those in the phase
of maturation and stabilization (medium-aged); and stabilized (old).
It follows from numerous published reports that in a landfill’s initial operation
phase, leachates contain organic substances that are relatively easily subject to
biochemical transformations in biological wastewater treatment plants. With
increasing landfill age, content of simple organic compounds decreases and highmolecular-weight compounds with little susceptibility to biodegradation first come
out, and finally dominate in leachates. According to SURMACZ-GÓRSKA et al.
(2000), in case of organic substance content expressed as COD being low and not
exceeding 2,000 mg⋅dm-3, one deals with compounds that are hard to decompose
biologically. Simultaneously with biochemical transformations, processes
of adsorption, dissolution, dilution, ion exchange and precipitation occur as a result
of which concentration of organic and inorganic substances varies in time
(TREBOUET ET AL. 2001, KANG ET AL. 2002, TATSI ET AL. 2003, RIVAS ET AL. 2004,
OZKAYA 2005).
With increasing landfill age, strong decrease of both BOD5 and COD values can
be observed as well as reduction of BOD5:COD ratio in leachates, according to data
available in the literature (Table 1).
It follows from studies carried out by KACZOREK, LEDAKOWICZ (2002) and
SURMACZ-GÓRSKA et al. (2000) that concentration of ammonia nitrogen in
leachates from domestic waste landfill sites reaches the value of 3,000 mg⋅dm–3.
OBRZUT (1997) reports that ammonia nitrogen concentration in leachate samples
taken from ten landfill sites in Poland varied from 1.7 to 1,520 mg⋅dm–3, with the
average value of 398 mg⋅dm–3. Studies carried out by KULIKOWSKA (2002) revealed
that concentration of ammonia nitrogen increased together with increasing landfill
age within the first five years of operation from about 100 mg⋅dm–3 up to
600 mg⋅dm–3.
An attempt to determine the nitrogen content in leaches from mature and
stabilized landfills was carried out by other authors. EL-FADEL et al. (2002) report
that together with increasing landfill age, ammonia nitrogen content in leachates
decreases from 1,500 mg⋅dm–3 in the methane fermentation phase to less than
50 mg⋅dm–3 in stabilization phase. Similar trend was observed by KANG et al.
(2002). In the period of first ten years of landfill operation, ammonia nitrogen
concentration remained at virtually constant level of 1,826–1,896 mg⋅dm–3, and then
decreased down to the value of 892 mg⋅dm–3.
Sources of heavy metals in leachates include industrial waste such as ashes,
batteries, dyes etc. (ERSES, ONAY 2003) as well as municipal waste containing
60
electronic parts, fluorescent lamps, thermometers, batteries, pesticides and other
(WARD et al. 2005).
Table 1
Concentration of organic compounds in leaches versus landfill age measured by means of
BOD and COD indicators
Landfil
l age
[years]
<5
5-10
> 10
> 20
Value
Parameter
COD
BOD
[mg.dm-3]
[mg.dm-3]
2640
600
1727
1058
1183
331
41507
32790
15000-40000
10000-25000
480 - 1801
76 - 721
3000-15000
300-15000
1660 - 1700
100 - 160
2150
215
10000-20000
1000-4000
5348
2684
< 3000
< 300
550
16.5
1000 - 5000
50 -1000
1367
145
7400 - 8800
475
2422 - 3945
106 - 195
< 1000
< 50
BOD/COD
0.22
0.61
0.28
0.78
0.6
0.1-1
0.06-0.09
0.1
0.1-0.2
0.5
< 0.1
0.03
0.05 - 0.2
0.1
< 0.1
0.03 - 0.05
< 0.05
References
LO i in. (1996)
SURMACZ-GÓRSKA i in. (2000)
SURMACZ-GÓRSKA i in.. (2000)
KANG i in. (2002)
EL-FADEL i in. (2002)
KULIKOWSKA (2002)
KURNIAWAN i in. (2006)
LO i in. (1996)
TREBOUET i in. (2001)
KANG i in. (2002)
KURNIAWA i in. (2006)
TREBOUET i in. (200)]
KANG i in. (2002)
RIVAS i in. (2004)
BILA i in. (2005)
Solubility and mobility are closely related to transformations occurring in the
landfill and depend on reaction value, redox potential, and, moreover, on presence
of organic and inorganic substances able to form complexes (BOZKURT et al. 2000).
The highest concentrations of metals were detected in leachates from young
landfills still in the acid fermentation phase, when the reaction value was low. In
both maturation and stabilization phases, the reaction value becomes neutral and
solubility of metals decreases (ERSES, ONAY 2003).
With increasing landfill age, change in type of organic matter degradation
products occurs — from low-molecular-number volatile organic acids to fulvic and
humus acids. Solubility of high-molecular-weight acids is different, but most
of them demonstrate metal sorption ability. Metals adsorbed on surfaces of humus
substances may therefore appear in both colloidal and suspension fractions.
In the landfill’s maturation phase, insoluble metal sulfides are created easily.
Metals may be also precipitated in the form of carbonates, hydroxides, and even
phosphates (ERSES, ONAY 2003).
SLACK et al. (2005) report that only 0.02% of heavy metals dumped on a landfill
site penetrate to leachates within the period of 30 years. Similarly, BOZKURT et al.
61
(2000) claim that more than 99.9% of metals are retained in the landfill bed as
a result of sorption on both organic and inorganic molecules (e.g. iron hydroxy
oxides) and precipitation. Similar observations were made by other authors
(AL-YAQOUT, HAMODA 2003, ERSES, ONAY 2003, WARD et al. 2005). In some
cases, concentration of metals can be quite high. SLACK et al. (2005) report that
concentration of zinc can reach the level of up to 1000 mg⋅dm–3, that of nickel —
13 mg⋅dm–3, copper — 10 mg⋅dm–3, and lead — 5 mg⋅dm–3. ROBINSON et al. (2005)
demonstrated that upper limit of chromium concentration can reach the value
of 13.1 mg⋅dm–3. Concentration of mercury in leachates can be even as high as
2 mg⋅dm–3 (BILA et al. 2005).
Presence of organic and inorganic hazardous substances and, in many cases, also
unidentified products of degradation of organic substances, is the cause of leachate
toxicity. In the opinion of CLÉMENT et al. (1997), ammonia nitrogen, heavy metals
(Ag, Hg, Pb, Cd, Mn, Zn and Cu) and organic compounds such as tannins, lignin
and phenols, can result in toxicity of leachates.
Studies on toxicity of leachates are rather rare. To date, the following groups
of organisms were used for this purpose: destroyers — Vibrio fisheri; producers —
Scenedesmus subscapitatus, Lemna minor, Selenastrum capricornutum; and
consumers — Brachionus calciflorus, Daphnia magna, Thamnocephalus platyurus
(CLÉMENT et al. 1997, SILVA et al. 2004, MERIÇ et al. 2005).
Determinants of the research
In recent years, intensive research work is carried out on treatment of landfill
leachates. For the studies, leachates from municipal waste landfill site in Wysieka
near Bartoszyce (Warmisko-Mazurskie province) were used. The work was carried
out in the period when stabilization of biochemical processes in the landfill bed
occurs typically. Total nitrogen concentration in leachates was then at the level of
749 mg⋅dm–3, and that of ammonia nitrogen — 636 mg⋅dm–3.
Technological research work was carried out simultaneously on four research
stands in reactors denoted as SBR 1, 2, 3 and 4. The leachate retention time was
fixed and amounted to 3 d, with the cycle period of 24 h.
Reactors SBR 1 and 3 contained active sludge, while reactors SBR 2 and 4 were
filled with active sludge and stationary packing suspended below leachate surface.
The packing consisted of 42 strips of PVC sponge with dimensions of 2 × 11 cm.
The operational cycle of SBR 1 and 2 reactors included filling, aeration,
sedimentation and aeration phases (reactor’s aerobic operation conditions), while
operation of SBR 3 and 4 reactors consisted of filling, stirring, aeration,
sedimentation and decantation phases (anaerobic-aerobic conditions).
In order to determine effectiveness of the process, the following contamination
indicators were controlled in the SBR reactors’ inlets and outlets: ammonia nitrogen,
nitrate nitrogen, and nitrite nitrogen (HERMANOWICZ et al. 1999).
For the purpose of further removal of organic substances from biologically
treated leachates, they were subject to further chemical treatment with the use
of Fenton’s reagent. The research work on leachate treatment by means
62
of the advanced oxidation method was carried out in static conditions of laboratory
reactors with capacity of 1 dm3 equipped with magnetic stirrer. Chemical reagents
were dosed on one-time basis in the beginning of each cycle, directly to the reactor.
Reaction proceeded at pH = 3. In the experiment, the hydrogen peroxide dose was
applied amounting to 3 g⋅dm–3 at decreasing share of Fe2+. Three Fe2+ : H2O2 rates
were examined: 1 : 10, 1 : 5 and 1 : 3. Effectiveness of oxidation of organic
compounds versus time was controlled. Measurements were carried out: at the
moment of reagent application; then after 1 min; 5 min; 30 min; 1 h; 1.5 h;
and finally after 2 h.
In order to determine effectiveness of leachate treatment process from advanced
oxidation chambers, concentration of organic substances was analyzed expressed as
COD (determined by means of bichromate method) (HERMANOWICZ et al. 1999) the
reaction value (HI 8818 pH-meter).
Effectiveness of treatment of leachates from municipal waste landfills
Biological methods
It was found that in aerobic conditions, presence of packing had no effect on
forms of nitrogen presents in reactor outflows. In reactors operating in anaerobicaerobic conditions, introduction of packing resulted in increase of ammonia nitrogen
concentration and decrease of nitrates. Individual nitrogen forms occurring in
leachates are presented in Figure 1.
800
concentration N [mg.dm-3]
700
600
0,12
0,55
0,79
500
400
2,95
3,78
0,7
300
179,2
200
100
685,2
680
650
4,71
8,61
0
SBR 1
nitrate
SBR 2
nitrite
SBR 3
SBR 4
ammonium
Fig. 1. Nitrogen forms in treated leachates
Introduction of packing resulted in losses of ammonia nitrogen, the fact being an
indication of the effect of simultaneous nitrification and denitrification in active
sludge.
Research work on leachates aimed at selection of optimum treatment technology
is carried out since introduction of waste landfilling to the engineering practice.
High concentration of biomass and long age of sludge permit for more effective
removal of organic substances with participation of slowly multiplying heterotrophic
bacteria and maintenance of sufficient concentration of nitrification bacteria
63
population. In order to increase biomass concentration and age of microorganisms,
attempts are made to utilize highly efficient reactors such as fluidized beds
or reactors with suspended biomass. Studies aimed at nitrification effectiveness
improvement were also carried out with the use of combined methods — active
sludge with biological membrane growing on a packing (movable or immovable
carriers).
It follows from data published in the literature that packing materials may
include: activated carbon, sand, plastics or PVC sponges. Organism grow on carrier
surfaces or inside their porous structures (GIESEKE et al. 2002). To date, attention
of researchers was focused only on increasing of the nitrification rate (VAN DE
GRAAF et al. 1995). Simultaneous use of biomass carriers allows to improve yield
of biomass and extend the period for which microorganisms of active sludge dwell
in the reactor.
GIESEKE et al. (2002) in the SBR reactor packed with Kaldnes mouldings
(SBBR) reached a decrease of ammonia nitrogen from 13 mg⋅dm–3 to 0.8 mg⋅dm–3,
with simultaneous increase of nitrates(V) up to 1.5 mg⋅dm–3 at oxygen concentration
in the reactor amounting to 5.4 mg⋅dm–3. After oxygen concentration being increased
up to 6.3 mg⋅dm–3, the ammonia nitrogen level decreased from 40 mg⋅dm–3 to
11 mg⋅dm–3, and content of nitrates(V) increased to 14 mg⋅dm–3.
It follows from studies of LOUKIDOU and ZOUDOULIS (2001) that it is possible
to reduce ammonia nitrogen from leachates by 60%. The research work was carried
out in a reactor packed with polyurethane cubes, and raw leachates were
characterized with ammonia concentration at the level of 1,800 mg⋅dm–3. Low degree
of reduction of ammonia nitrogen from leachates resulted from the fact that 18 hlong aeration period was insufficient to achieve full nitrification. Another reason can
consists in presence of large amount of organic compounds prohibiting sufficient
number of nitrification bacteria from multiplication.
ROSTRON et al. (2001) treated synthetic wastewater characterized with ammonia
nitrogen concentration of 500 mg⋅dm–3. The research work was carried out in CSTRs
under aerobic conditions. As packing, Linpor polyurethane cubes, Kaldnes
polyethylene mouldings and capsules made of PVA (polyvinyl alcohol) were used.
Full nitrification was achieved at the period of wastewater retention in reactor of 6;
3.4 and 2.2 d, respectively. Reduction of the retention time to 1.5 d resulted in
occurrence of nitrites in the outflow. That could be caused by washing out
Nitrobacter bacteria or too little amount of biomass with respect to the increasing
load. At the retention time of 1 d, full nitrification in the reactor packed with Linpor
and PVA blocks was achieved, while 90% nitrification was obtained with Kaldnes
mouldings. After reduction of the retention time down 0.5 d, phase-II nitrification
was stopped in all reactors. In case of packing made of PVA, 30% nitrification was
observed. At the same time, 30% of ammonia nitrogen remained not removed.
In aerobic conditions, oxidation of ammonia nitrogen to nitrates(III) is carried
out by Nitrosomonas sp. (phase I nitrification). Then, nitrite(III) nitrogen is oxidized
to nitrate(V) nitrogen with the aid of Nitrobacter sp. (phase II nitrification) (VAN
DER STAR et al. 2008).
64
Presently, a matter of significant importance for practical purposes related to
removal of ammonia nitrogen from wastewater consists in processes of partial
nitrification, oxidization of ammonia nitrogen in anaerobic (anoxic) conditions
or combination of both processes. Among processes with dominant mechanism
of ammonia nitrogen removal one can rate: partial nitrification and Sharon,
Anammox and Canon methods (KHIN, ANNACHHARTE 2004).
Partial nitrification can constitute the first stage in a system with conventional
denitrification or Anammox. In systems with partial nitrification, the denitrification
process consists in reduction of nitrate(III) nitrogen to molecular nitrogen and
occurs at higher rate compared to reduction of nitrate(V) nitrogen (TURK, MAVINIC
1989). In systems with partial nitrification and denitrification, reduction of demand
for oxygen (by 25%) and organic carbon (by 40%) occurs. An additional advantage
consists in lower biomass production and CO2 emission (SCHMIDT et al. 2003, KHIN,
ANNACHHARTE 2004).
Presently it is assumed that in reactors with full mixing, maintaining a short
retention period (e.g. one day) and high temperature (30–40ºC) is favorable for
partial nitrification as it leads to washing out Nitrobacter sp. from the reactor. From
the point of view of Nitrobacter sp. cultivation, low concentration of dissolved
oxygen (less than 0.4 mg⋅dm–3) and high concentration of ammonia nitrogen
(SCHMIDT et al. 2003) are unfavourable.
Modifications of Sharon process towards increase of its stability consist in
shortening of period for which the sludge is retained in the reactor through
abandonment of sludge recirculation. That way, instead of wastewater retention
period, the sludge age is controlled that should be long enough in order to ensure
sufficient multiplication of Nitrosomonas sp. in the active sludge chamber, but at the
same time long enough to achieve complete washout of Nitrobacter sp. population
from reactor (SCHMIDT et al. 2000). The final product of Sharon process is
nitrate(III) nitrogen, but also not oxidized ammonia nitrogen. That is why Anammox
method is recommended as the second stage for the purpose of complete ammonia
nitrogen removal.
The Anammox process was discovered by MULDER and SCHMIDT et al. (1995).
The authors examined anaerobic oxidization of ammonia nitrogen in presence of
nitrate(V) nitrogen in the laboratory scale using an anaerobic fluidized reactor. Later
research work carried out by VAN DE GRAAF, SCHMIDT et al. (1995) and BOCK,
SCHMIDT et al. (1995) proved that rather nitrate(III) nitrogen, and not nitrate(V)
nitrogen, is preferentially used as the electron acceptor. Nowadays it is a common
assumption that Anammox consist in denitrification of nitrates(V) or (III) with
ammonia nitrogen as the electron donor (KHIN, ANNACHHARTE 2004).
VAN DONGEN, SCHMIDT et al. (2001) examined removal of ammonia nitrogen in
a two-stage system using the Sharon method as the 1st stage and Anammox as the 2nd
stage treatment. At wastewater retention period amounting to 24 hours, they
achieved a 53%-reduction conversion of ammonia nitrogen into nitrate(III) nitrogen
in the 1st stage, while in the 2nd stage, at reactor load amounting to about 1.2 kg⋅m–
3. –1
d , the nitrogen removal efficiency exceeded 80%.
In the Canon process, elimination of nitrogen in a single reactor proceeds with
participation of bacteria of genus Nitrosomonas sp. and plankton bacteria. In
65
conditions of limited concentration of dissolved oxygen, Nitrosomonas sp. oxidize
ammonia nitrogen to nitrate(III) nitrogen, and plankton bacteria transform ammonia
and nitrate(III) nitrogen to molecular nitrogen and, to some small concentration, to
nitrates(V). In order to increase efficiency of the process and eliminate nitrate(V)
nitrogen it is recommended to introduce inorganic substances as an external source
of carbon (HAO, VAN LOOSDRECHT 2004).
Chemical methods
-3
concentration COD [mg.dm ]
It was found that degradation of organic substances in leachates, at H2O2 dose
amounting to 3 g⋅dm–3 and iron(II) doses equalling 1, 0.6 and 0.3 g⋅dm–3, efficiency
of removal of organic compounds was high and amounted to 48% at Fe2+ : H2O2
ratio of 1 : 5 and 1 : 3 and 45% at Fe2+ : H2O2 = 1 : 10. Variations of content
of organic substances in time are presented in Figure 2.
1000
800
600
400
200
0
0
0,016 0,08 0,5
1
01:10 01:05 01:03
1,5
2
time [h]
Fig. 2. Variations of content of organic substances in time at various
Fe2+ : H2O2 ratios
The best treatment results are achieved in two-stage systems representing
a combination of biological and physicochemical methods. Physicochemical
methods differ in efficiency, degree of complexity of technological solutions and
related apparatus as well as in costs of individual processes. In the engineering
practice, leachate treatment is commonly carried out by means of adsorption,
coagulation/flocculation and reverse osmosis.
Degradation of refraction substances can be achieved by means of advanced
oxidation. The method consists in generation of reactive hydroxyl radicals (OH•).
Hydroxyl radicals (with oxidization potential of 2.8 V) react stronger than chemical
oxidizers such as ozone or H2O2 and are not selective, thus making possible for them
to react with many chemical compounds.
It follows from review of the literature that leachate treatment can be quite
effectively carried out with the use of Fenton’s reagents (Fe2+ : H2O2). In two-stage
systems it can be applied both before and after biological treatment of leachates
(LOPEZ et al. 2004; ZHANG et al. 2005).
66
The Fe2+ : H2O2 is of special importance, as it allows to avoid undesirable free radical reactions that may occur in case of excess of any of the reagents. When the
proportion of Fe2+ with respect to H2O2 is optimal, OH radicals are used mainly for
oxidization of organic substances (LOPEZ et al. 2004).
For that reason, many authors dealt with the problem of proper selection
of proportion between reagents used for oxidation of organic substances. LOPEZ et
al. (2004) demonstrated that at fixed H2O2 dose amounting to 6.3 g⋅dm–3, COD in
treated leachates decreased with increasing concentration of iron(II). After increase
of the hydrogen peroxide dose up to 10 g⋅dm–3 and at the proportion Fe2+ : H2O2 =
1 : 8, the increase of BOD5 : COD from 0.2 to more than 0.5 was observed.
The authors concluded that final reaction products consist in short-chain organic acids
resistant to further oxidation.
The advanced oxidation reaction is most effective al low reaction value. KANG,
HWANG (2000) and LOPEZ et al. (2004) report that the highest efficiency of Fenton’s
reaction occurs at pH within the range 2.5–4. In the research work reported here,
reaction of advanced oxidation with Fenton’s reagent was carried out at pH = 3.
ZHANG et al. (2005) examined effectiveness of leachate treatment from
municipal waste landfill in Sandtown characterized with organic substance
concentration (COD) within the range 8,298–8,894 mg⋅dm–3. They demonstrated
that at pH = 2.5, the course of Fenton’s reaction was the most effective and rate
of generation of iron (III) ions was the highest. The authors shown that with
increasing temperature, efficiency of organic substance removal also increased. For
instance, at initial COD level of 1,000 mg⋅dm–3, the efficiency increased from 42.3%
(at temperature 13ºC) to 56.2% (at 37ºC), while at the initial level of 2,000 mg⋅dm–3,
COD reduction increased from 31.6% to 44.8%, when temperature raised from 15°C
to 35ºC.
Summary
Usefulness of combining biological methods with chemical ones is beyond any
doubt proved in case of removal of organic substances from leachates. In that
situation, it is also important to ensure equally high nitrogen removal efficiency.
In the opinion of ALBERS, KRÜCKEBERG (1992), KACZOREK et al. (2002) and
ZHANG et al. (2005), even in cases when physicochemical processes such as
ozonization or reverse osmosis are used, preliminary removal of nitrogen by means
of biological methods is both appropriate and economically justified.
Research work on treatment of leachates from municipal waste landfill sites
by means of the active sludge method in SBR reactors demonstrated that oxygen
concentration reduction through introduction of an anaerobic phase into the SBR
reactor operation and/or packing impairs the process of nitrification. Introduction of
packing leads to increase of nitrogen losses and/or ammonia nitrogen concentration
in effluents from reactors operating in anaerobic-aerobic conditions.
Further attempts to remove organic compounds from leachates with the use
of Fenton’s reagent proved that Fe2+ : H2O2 ratio has insignificant effect on the
67
reaction rate and process effectiveness. Increase of H2O2 in the mixture resulted
in slight decrease of effectiveness of the process.
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VAN
Justyna Koc-Jurczyk
The Chair of Natural Theories of Agriculture and Environmental Education
Faculty of Biology and Agriculture
University of Rzeszow
ul. Cwiklinskiej 2, 35-601 Rzeszów, POLAND
70
CHAPTER V
Wiera Sdej1, Zbigniew Luliski2, Janusz Posłuszny2
IMPACT OF MUNICIPAL LANDFILLS ON QUALITY
OF GROUND AND SURFACE WATERS
Introduction
Any human activity generates waste, which can be a serious threat to the natural
environment. Comprehensive waste management is one of these aspects
of environmental protection which, as the amount of waste is growing, have become
a problem awaiting an urgent solution. The current model of waste management
in Poland needs to be termed as a typically extensive one, with nearly 97%
of the generated waste ending up on dumping sites. The result is the progressing
degradation of all components of the natural environment, which have an important
influence on man’s comfort of life (DRZAŁ et al. 1995, SZYSZKOWSKI 1995,
BŁASZCZYK, GÓRSKI 1996, DOBRZYSKA et al. 1998, GRUSZKA, PLEWNIAK 1999,
BARTOSEWICZ 2002, ROSIK-DULEWSKA 2007).
Before selecting an adequate method of waste management, one should become
acquainted with the characteristics of all methods, many of which involve modern
technologies that incur high investment expenses. When waste is deposited on
a landfill, high outlays go to the selection of a right site as well as construction and
maintenance of this object. A waste dump that has been properly designed will
produce less negative impact on people, animals, plants and other elements of nature
such as water, soil and air (AVERESCH 1995, JURKIEWICZ et al. 1998, SZYMASKA –
PULIKOWSKA 2001b, YGADŁO 2001, KULIG 2002, AL-YAGOUT, HAMODA 2003,
SLACK et al. 2005).
Regarding the lithosphere, degradation of the environment caused by presence
of landfills includes penetration and accumulation of various substances in the
ground. With respect to the hydrosphere, the immediate risk is caused by emissions
of polluted leachate to lakes, rivers or groundwater (BELEVI, BACCINI 1989a,b,
ALISTAIR 2000, BOZKURT
et al. 2000, SZYMASKA-PULIKOWSKA 2001a,
AL-YAGOUT, HAMODA 2003; WÓJCIK et al. 2005).
Permeation of pollutants in leachate to the hydrosphere can occur both while
a landfill is used and after it has been closed down. Therefore, an important aspect
of waste management consists of proper reclamation of the surface of a waste
dumping site after its exploitation has been terminated.
This paper present an analysis of the impact of leachate from the currently
operating municipal landfill in Brodnica on some properties of ground and surface
71
water. A detailed analysis was performed to test values of physical, oxygen-related
and salinity indices.
Conditions of the research
The study has been conducted at the municipal waste landfill in Brodnica, which
is run by the Municipal Management Company, Ltd. The landfill, which was opened
in 1997, is situated about 3 km south-west of the town centre. It lies 350 m away
from the Drwca River and 900 m from the border of the zone of indirect protection
of the town’s water intake. To the north, the landfill is adjacent to the municipal
wastewater treatment plant, whereas to the west it borders with a municipal animal
asylum (Fig. 1).
Fig. 1. Location of Brodnica Landfill
The landfill basin was designed and constructed as an earthen tank, which
is surrounded by an earthen dyke raising about 2.5 – 4.0 m above the land.
The superstructure of the basin’s subsoil was constructed from the silts originating
from the trenches dug out during the macro-levelling of the south-western slope.
The landfill was prevented from producing negative impact on the environment with
a screen isolating the landfill basin from the groundwater and limiting migration
of pollutants into the ground. It was particularly important to provide the sealing
with maximum efficiency as the Drwca River, flowing near the landfill, supplies
potable water to the municipality of Torun.
Considering the required resistance to the expected load of waste and the
pressure generated by waste compactors as well as the required resistance to
72
aggressive effects produced by waste, a leachate drainage system was installed at the
bottom of the basin. At the ends of drainage pipes as well as at the sites where
drainage pipes were fitted into a collector, control wells made of Ø 1,000 mm
concrete rings were installed. The drainage system collector ends with a well made
of Ø 1,200 mm concrete rings, from which a PE Ø 250 mm pressure pipe
transports leachate, by the gravitational force, to a prefabricated Ø 1,400 mm
pumping station. From the pumping station, leachate is transported by a piston
pipeline to a concerte ring well located near the entrance gate to the landfill
(PRZEGLD EKOLOGICZNY, 2002).
Regarding the climate, Brodnica and its environs lie between the mild climate of the
Great Valleys region and a more severe lakeland climate of the Masurian Lake
District. Typical features of this area are quite chilly, snowy and long winters,
generally cool summers and low rainfall. The average annual air temperature is
7.5ºC, while the average annual precipitation is 556 mm. The average monthly
relative air humidity varies from 70% in May to 88% in November and December
(KULIG et al. 2003).
In order to establish the effect of leachate on quality of subsurface (ground and
deep groundwater) and surface water, measurements of the water table were taken
and physicochemical properties of leachate and water samples were tested. Water
samples for analyses were collected according to the Polish norms: PN-ISO 566711:2004; PN-76/C-04620 and PN-88/C-04632, once every three months, at six
different locations.
The following were determined in the samples of subsurface and surface water:
reaction (pH), proper electrolytic conductivity, dissolved substances, sulphates
(SO4+2), chlorides (Cl-) and content of calcium (Ca+2) and magnesium (Mg+) ions.
The laboratory tests were performed according to the analytical recommendations
contained in the Polish Norms: [PN-90/C-04540/01; PN-78/C-04541; PN-74/C04578/03; PN-ISO 9297/1994; PN-79/C-04566/10; PN-ISO 6058/1999 and PN-ISO
6059/1999].
Samples of leachate were collected from the pumping station situated behind the
landfill basin. Samples of subsurface water were taken from four piezometers
located around the landfill basin. Deep groundwater was sampled from piezometer
P1 (a model observation borehole for determination of the hydrochemical
background), drilled to test inflowing groundwater, and from piezometer P2, located
where water flows away from the landfill. Samples of groundwater were taken from
piezometer P3, situated on a bank of a narrow-gauge railway track, and from
piezometer P4, drilled around 8 meters from the leachate pumping station. Both
piezometers capture water flowing away from the landfill.
The localization of the leachate pumping station, the piezometers for
measurements of the quality of subsurface water and the water pond is illustrated in
Fig. 2.
73
Fig. 2. Location of the pumping station, piezometers for monitoring quality of subsurface
water and the water pond
The results of the tests were processed statistically using the software Statistica
ver. 6, by StatSoft, Inc. (2001). The least significant differences (LSD) were
determined at the level of significance p=0.05.
The quality of waters was assessed according to the criteria expressed in the
legal regulations contained in the Ordinance of the Minister of the Environment of
11 February 2004 on classification of surface and subsurface water, monitoring
methods, interpretation of the results and presentation of the quality of water
(Journal of Law, No 32, item 284) and Ordinance of the Minister of the
Environment of 14th July 2006 on execution of duties laid on industrial sewage
suppliers and conditions for disposal of sewage to sewage facilities (Journal of Law,
No 136, items 963 and 964).
Production and amounts of landfill leachate
With the variety of disposed waste, changeable atmospheric conditions and
microorganisms present in the bed, landfills are referred to as a certain type
of bioreactors, in which complex processes of degradation and biotransformation
occur. These processes, both aerobic and anaerobic ones, lead to the formation of
highly mineralized substances characterized by various toxicity to live organisms
(AVERESCH 1995, COSSU et al. 2000, LEDAKOWICZ, KACZOREK 2004).
74
The intensity of processes taking place in the deposited mass of waste is affected
by many factors, mainly water content and oxygen availability. Some precipitation
water falling on the surface of a landfill evaporates, some flows over the surface and
some, alongside the water supplied with the waste and originating from
decomposition of organic waste, migrates through the bed, where it is enriched with
soluble compounds. As a result, a by-product of landfills, called leachate, appears.
Most researchers (ROSIK- DULEWSKA, KARWACZYSKA 1998, GRUSZKA,
PLEWNIAK 1999, YGADŁO 1998, 2001, BARTOSEWICZ 2002) claim that leachate
appears primarily due to the penetration of precipitation water to the landfill
reservoir and, to a much lesser extent, via decomposition of the organic fraction
found in the waste mass. Increased production of leachate can be also caused by
surface and subsurface water reaching the landfill which lacks a proper system for
draining such water (BELLEVI, BACCINI 1989A,B, LO 1996, YGADŁO 1998,
OLESZKIEWICZ 1999, BOZKURT et al. 2000).
Because of its content of chemical substances and compounds that alter the
natural composition of water, leachate is considered to be wastewater. Moreover,
because it contains elevated amounts of halogen derivative compounds, it is
classified as dangerous wastewater (SURMACZ – GÓRSKA et al. 2000).
The composition and amount of leachate can be highly varied, depending on the
type of waste, its fragmentation and density, amount of water trickling through the
bed, age of the landfill, technology for storing waste, physicochemical
transformations occurring in the landfill body and the way the landfill is reclaimed
(OBRZUT 1997, RUBACHA, ROGOWSKA 1997, BERGIER, WÓJCIK 2001, KLOJZY –
KARCZMARCZYK et al. 2003, ROBINSON et al. 2005).
The amount of generated leachate depends primarily on the volume of
atmospheric precipitation as well as on the evaporation and insulation. When the
annual precipitation reaches 700 mm, the density of deposited waste is 600 kgm-3
and its water content is 30%, the amount of generated leachate per 1 ha of the
landfill surface area is ca 450 mmyear-1, i.e. 4.500 m3. Loss of water through
evaporation and surface runoff is 250 mm. However, this is a simplified calculation
as it does not take into consideration all possible factors. Nonetheless, it gives very
close approximation of the volume of generated leachate (SUCHY et al. 1998,
OLESZKIEWICZ 1999, BARTOSEWICZ 2002, GÓRSKI 2002). Moreover, amounts of
leachate change seasonally. Most leachate will appear between September and April,
while in the late spring and summer only minimal amounts of leachate are produced.
There are also possible daily peaks caused by rapidly melting snow or heavy rains.
When this happens, amounts of leachate can be up to ten-fold higher than observed
under natural conditions (KODA 2001).
With respect to the waste deposited on the landfill in Brodnica, the amount of
leachate discharged to the municipal wastewater treatment plant was steadily
increasing (Fig. 3).
The highest amount of leachate was recorded in December, and the smallest one
occurred in September 2004, which confirms the results of studies completed by
other authors. From October 2004 to March 2005, the quantity of leachate was
observed to have increased considerably relative to the summer season. This was
caused by a very high volume of precipitation which occurred in the autumn and
75
3
Quantities of leachate [m ]
winter of that year. Another reason was the fact that a draining system was installed
in Section II of the landfill.
1000
934
800
861
702
683
600
400
910
803
453
349
251
269
179
200
121
0
IV
V
VI
VII
VIII
IX
X
XI
XII
I
II
III
Months 2004/2005
Fig. 3. Quantities of leachate in 2004-2005
Chemical composition of leachate
The precipitation water trickling through the landfill as well as subsurface and
surface water cause leaching of water soluble substances. These three sources of
water largely affect the qualitative (chemical) composition of generated leachate. It
is assumed that the chemism of leachate depends mainly on the content of organic
substance in waste, the stage of waste transformation and technology of waste
deposition. Another important factor is the chemical composition of deposited waste
as well as decomposability and leachability of particular waste components
(CLÈMENT et al. 1997, ALLEN 1999, DBROWSKA et al. 1999, KANG et al. 2002,
KODA et al. 2006).
The main process which takes place on a landfill is the microbial decomposition
of organic matter, followed by the reaction of decomposition products with other
components. As a result, substances found in leachate are a mixture of compounds
originating from solid components dissolved in water and liquid components as well
as intermediate products occurring during the fermentation processes. The final
composition of leachate, as the above implies, is a resultant of processes occurring
in a mixture of old and fresh waste.
The load of pollutants in leachate produced during the early stage of waste
deposition is much higher than in leachate produced during later stages. This
dependence holds particularly true for organic pollutants, as can be demonstrated
with an aid of oxygen indices, i.e. BOD5, COD, as well as TOC (total organic
carbon). The value of leachate reaction reveals a reverse regularity. In the initial
years after opening a landfill (up to five years), landfill leachate is acidic (pH 3.7 to
6.4), which is directly caused by the processes occurring in the waste bed. During
this phase of the landfill exploitation, the decomposition of waste generates shortchain volatile acids, which make up 70-90% of organic components, as well as
hydrogen and carbon dioxide. These chemicals are directly responsible for
76
acidification of leachate. In the later years of exploitation, leachate becomes neutral
or slightly alkaline (7.0-7.6) and after about ten years it is alkaline in reaction (8.08.5) (SZUBSKA 1997, SUCHY in. 1998, VADILLO et al. 1999, PRZYWARSKA 2001,
SZYMASKI 2006, ROSIK-DULEWSKA 2007).
The reaction of the leachate produced at the landfill in Brodnica oscillated
around 7.5 to 7.8, which is characteristic for stabilized landfills. No significant
differences in the value of this parameter were observed between the leachate
sampling dates (Fig. 4).
The values of oxygen indices reported by different authors (SURMACZ-GÓRSKA
et al. 1997, SZUBSKA 1997, VADILLO et al. 1999, SZPADT 2006, SZYDŁOWSKI 2007)
oscillate within broad ranges and depend mainly on the age of a landfill. In leachate
from young landfills, the values of these indices are much higher than from older
ones. In leachate from landfills exploited for three years, the values of BOD5 vary
between 1,500 and 45,000 mg O2 m-3 and the values of COD – between 3.600 and
62,000 mg O2 m-3. In leachate from landfills used for over 3 years, BOD5 equals 250
to 16,000 mg O2 m-3 and COD reaches 2,800 to 19,000 mg O2 m-3.
pH value
7,8
7,7
7,6
7,5
7,4
7,3
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates, LSD0.05 = 0.60
Fig. 4. Value of reaction (pH) of leachate
Higher content of organic compounds in leachate from younger landfills is
caused by the fact that during the initial years of waste deposition on a landfill,
processes of organic substance decomposition are the most intensive. At that time,
during the acidogenic phase, the highest amount of easily soluble organic bonds are
created. At landfills older than 3 years, organic matter decomposition processes slow
down, which is clearly reflected by numerical values of oxygen indices. During that
time, due to the progressing stabilisation processes (waste methane fermentation
phase), small quantities of hardly decomposable organic compounds, mainly humic
and fulvic acids, appear in leachate (VADILLO et al. 1999, ALISTAIR 2000, COSSU
et al. 2000, LEDAKOWICZ-KACZOREK 2004).
Among the oxygen indices characterising properties of leachate from the waste
landfill in Brodnica, the following were determined: BOD5, CODCr and total organic
carbon. During our study, fluctuations in BOD5 in leachate were within 45.00 to
100.00 mg O2 dm-3 (on average 65.00 mg O2 dm-3) whereas CODCr varied from
170.00 to 682.00 mg O2 dm-3 (on average 373.90 mg O2 dm-3). Differences in the
values of these parameters between the sampling dates were in most cases highly
significant (Fig. 5, Tab. 1). The highest values of BOD5 were recorded in August
77
2004, and those of CODCr – in June 2004. The results of quantitative determinations
of oxygen (the values of COD and BOD5) for the analysed samples of leachate
suggest that small amounts of organic pollutants were present (Fig. 5). The
BOD5/COD ratio for all the samples reached between 0.03 and 0.38, which proves
that the landfill is stabilized.
The total carbon content in leachate ranges from 195.0 to 12,060.0 mg dm-3.
In the leachate from the municipal waste landfill in Brodnica this ratio was very low,
at 39.66 mg C dm-3. Although there was a large variation in the content of organic
carbon during our study (3.54 – 64.81), in none of the cases the determined values
were higher than the monthly values within the ranges quoted by various authors
(SZUBSKA 1997, KULIG 2002, KLOJZY-KARCZMARCZYK et al. 2003, ROSIKDULEWSKA 2007). The highest total carbon content was recorded in October and
the lowest – in August 2004. The differences between the values of TOC obtained in
June 2004 and March 2005 were not significant, while these between the other
months were highly significant (Fig. 6).
BOD5
CODCr
100
-3
mg O2 dm
60
.
mgO2.dm-3
80
40
20
0
VI 2004
VIII 2004
X 2004
III 2005
700
600
500
400
300
200
100
0
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates
Sampling dates
Fig. 5. Values of BOD5 and CODCr in leachate from the landfill
mg C•dm-3
80
60
40
20
0
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates
Fig. 6. Content of total organic carbon in leachate
Among the indices characterising salinity of leachate are proper electrolytic
conductivity and content of soluble substances such as sulphates, chlorides
and calcium and magnesium ions. Many authors report that values of these
78
parameters in leachate can be highly varied (BELLEVI, BACCINI 1989a,b,
DBROWSKA et al. 1999, KULIG 2002, KLOJZY-KARCZMARCZYK et al. 2003,
MOCZULSKA 2006A,B, KŁACZKO, ROSIK-DULEWSKA 2007, SZYMASKI et al.
2007). The highest concentrations of soluble substance appear in the first 2-3 years
of the exploitation of a new landfill.
In the leachate from Brodnica Landfill, the values of indices expressing the
salinity of effluents were also highly varied. The value of proper electrolytic
conductivity was within 1,115.1 – 7744.30 S cm (on average, 4,675.00 S cm).
High values of this parameters in leachate from municipal waste dumping sites have
been recorded at other locations (VADILLO et al. 1999, SZYMASKI et al. 2007). The
value of proper electrolytic conductivity at Brodnica Landfill demonstrably declined
during our study, which may have been a result of fitting a draining system
to Section II of the landfill. Consequently, the wastewater discharged to the WTP
was more strongly diluted. Similar tendencies appeared with respect to the values
characterising the content of soluble substances in leachate (Fig. 7).
-3
.
8000
7000
6000
5000
4000
3000
2000
1000
0
Disolved substances
mg dm
µS•cm
Proper elecrolytic conductivity
VI 2004
VIII 2004
X 2004
III 2005
4000
3500
3000
2500
2000
1500
1000
500
0
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates
Sampling dates
Fig. 7. Value of proper electrolytic conductivity and content of dissolved substances
in leachate
Values of salinity indices expressed as the content of sulphates and chlorides
were much lower compared to the composition of leachate from other landfills
similar in age (KLOJZY-KARCZMARCZYK et al. 2003, MOTYKA et al. 2005, KODA
et al. 2007, SZYMASKI et al. 2007) (Fig. 8). Noteworthy is the elevated value
of sulphate ions versus chloride ions, which indicates that the hydrochemical type
of leachate had been shaped as a result of leaching sulphate minerals, which are
a product of sulphate weathering. During our study, the concentration of sulphates in
leachate grew demonstrably – from 28.00 to 101.00 mg SO4-2 dm-3. Reverse
dependencies occurred in regard to the content of chlorides, whose ions fell from
897.75 mg Cl- dm-3 to 106.50 mg Cl- dm-3.
SURMACZ-GÓRSKA et al. (2000), who analysed composition of leachate form
three municipal waste landfills, different in age and exploitation technology,
demonstrated that high salinity in leachate, mainly the content of chlorine ions, is
caused largely by depositing street waste collected during winter season as well as
the release of chlorine during mineralization of organic substance in fermentation
processes that take place in masses of deposited waste. This, however, has found no
79
confirmation in the authors’ own study, as shown by much lower values of this ion
in leachate sampled in spring than in summer or autumn.
Chlorides
120
100
80
60
40
20
0
-3
1000
800
600
.
mg Cl dm
.
mg SO4 dm
-3
Sulphates
VI 2004
VIII 2004
X 2004
400
200
0
III 2005
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates
Sampling dates
Fig. 8. Content of sulphates and chlorides in leachate from the landfill
The average content of calcium ions was 72.64 mg Ca+2 dm-3, with values of this
parameter being significantly higher in leachate samples collected in autumn than in
spring (Fig. 9, Tab. 1). During our study, the level of magnesium fell demonstrably,
from 226.18 to 22.97 Mg+2 dm-3, which was obviously caused by the dilution of
leachate when a draining system was installed in Section II of the landfill. High
levels of magnesium ions in leachate from municipal waste landfills have been
noticed by other authors (VADILLO et al. 1999, MOTYKA et al. 2005).
Magnesium
250
80
200
-3
100
mg Mg dm
60
.
mg Ca.dm-3
Calcium
40
20
0
VI 2004
VIII 2004
X 2004
150
100
50
0
III 2005
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates
Sampling dates
Fig. 9. Content of calcium and magnesium cations in leachate from the landfill
Elements classified as heavy metals, whose salts are mostly toxic substances, are
particularly noxious pollutants in leachate. Leachate from landfills tend to contain
most of Fe ions, but other elements such as Cr, Ni, Cu, Cd and Pb appear as well,
albeit in lower concentrations. Heavy metals undergo more intense leaching during
the early years of operating a landfill than in later years. This is a consequence of
processes occurring in a young landfill which lead to acidification of leachate
(ROSIK-DULEWSKA, KARWACZYSKA 1998, SUCHY et al. 1998, BUDEK et al. 2000,
ROSIK-DULEWSKA 2003, KOZAKIEWICZ, MIKOŁAJCZYK 2003, SZYMASKI et al.
2007). The presence of this group of elements in leachate is mainly caused by
80
disposal of batteries, fluorescent bulbs, accumulators and empty paint, varnish or
solvent containers, etc. In countries where waste recycling is well-developed, the
content of heavy metals in landfill leachate is much lower (WARGAN 2002, WARD et
al. 2005, SZYMASKI 2006).
Table 1
Statistical calculations for values of physical, oxygen and salinity indices in landfill leachate
Index
LSD0.05
Standard deviation
Standard error
Reaction
BOD5
CODCr
Total organic carbon
Proper electrolytic
conductivity
Dissolved substances
Sulphates
Chlorides
Calcium
Magnesium
0.60
7.93
23.47
8.11
0.14
24.83
217.50
25.99
0.07
12.42
108.75
13.00
85.75
2903.90
1451.95
57.25
9.35
28.80
6.39
14.72
1294.60
34.51
327.47
16.15
85.60
647.30
17.25
163.74
8.07
42.80
In the early years of operating a landfill, the leachate also contains
bacteriological contaminants. OLESZKIEWICZ (1999), KA
MIERCZUK, KALISZ
(2001) and NIEMIEC and ZAMORSKA (2002) report that landfill leachate is severalfold more loaded with bacteria than municipal wastewater and sewage. In addition,
it also demonstrates a much larger variation of the bacterial fauna. Landfill leachate
contains numerous pathogenic microorganisms, including the ones responsible for
intestinal infections (typhois fever, dysentery, diarrhoea in children), tuberculosis,
tetanus, gas gangrene, anthrax, diphtheria and viruses of jaundice or Heine-Medina
as well as enteroviruses and adenoviruses. The most common bacteria are rods
of Salmonella typhi and Salmonella paratyphi. These bacteria are claimed to be
a potential source of pathogenic microorganisms, which can considerably affect the
level of pollution of ground and surface waters.
The effect of municipal waste landfill leachate on quality of ground and
surface waters
Landfills are typically situated on the surface or near the surface of the ground,
which means that they are within the natural circulation of water in the environment.
Atmospheric precipitation rinses trickles through a whole landfill, carrying leached
pollutants to subsurface, surface and even deep groundwater. Migration of pollutants
in leachate to the hydrosphere is certainly one of the gravest problems caused by the
presence and exploitation of landfills (AVERESCH 1995, BŁASZCZYK, GÓRSKI 1996,
BARTOSIEWICZ 2002, KLOJZY-KACZMARCZYK, MAZUREK 2003, KRYZA, CHUDY
2003, KŁACZKO-SZYMASKI 2007). The pollutants found in leachate, due to their
toxicity, lead to persistent contamination of surface waters, disrupting their natural
81
balance and inhibiting their self-purification. Pollutants can also enter groundwater,
mainly in the first aquifer, causing contamination (SZYSZKOWSKI 1995, KODA 2001,
MORYL, MORGA 2001a, b, KLOJZY – KARCZMARCZYK et al. 2003).
The extent of the influence of a landfill on the quality of water is measured as
a distance from the edge of the landfill cap to the line surrounding the landfill along
which the values of pollutants equal the values of the hydrogeochemical
background. As SZYMASKA – PULIKOWSKA (2001) reports, leachate infiltrating
from the landfill to the ground can be partly purified in the aeration zone and further
purification takes place in the zone of saturation of the aquifer. Under favourable
hydrogeochemical conditions, pollutants from leachate can migrate with
groundwater over large distances, exceeding 1,000 m.
According to BŁASZYK and GÓRSKI (1996) or MORYL and MORGI (2001a, b),
migration of pollutants from a landfill is mainly conditioned by the permeability of
rock formations in the substrate directly under the deposited waste. Apart from
migrating with the rainfall trickling through the landfill, pollutants can also reach
groundwater as a result of leaching the waste in the saturation zone if the water table
is high.
The total load of pollutants removed from the landfill depends on the type of
deposited waste and biological as well as physicochemical transformations which
occur in the landfill body. Reduction of the stream of precipitation trickling deep
through landfill has a significant influence on limiting the penetration of leachate to
the environment. Considering the variety of hydrogeological conditions, it is
extremely important to select a good location for a new landfill and to create
appropriate barriers reducing the outflowing infiltration water stream (TWARDY,
JAGU 2001A, SIKORSKA-MAYKOWSKA et al. 2002).
The volume of pollutants escaping the landfill leachate to ground and surface
waters can be evaluated through by monitoring the quality of water through
a network of piezometers or by analysing water quality in nearby homestead wells.
In our study, the level of deep groundwater measured in model piezometer P1,
situated in front of the landfill basin, where the groundwater was flowing to the
landfill, and in piezometer P2, drilled into a water stream flowing away from the
landfill, ranged within 1.75 and 3.50 m. In both piezometers, higher levels of water
were observed in summer and spring than in autumn (Tab. 2).
Table 2
Values of physical parameters of subsurface and surface water
Piezometers
Parameter
unit
Water level
m
Reaction
pH
1
2
Deep groundwater
3
4
Groundwater
*3.48
1.91
0.95
0.88
**3.44÷3.50
1.75÷2.10
0.65÷1.50
1.15÷0.25
7.46
7.00
7.18
7.36
7.30÷7.70
6.80÷7.35
6.95÷7.50
7.05÷7.85
*average
**fluctuations
82
Surface
water
(water
pond)
7.60
The average level of groundwater collected from piezometers P3 and P4, drilled
into the water flowing away from the landfill, was similar, oscillating between 0.88
and 0.95 m. In the former piezometer, the highest water level occurred in spring, and
the lowest one – in autumn; in the latter piezometer, higher values were determined
in summer and lower – in spring.
The average value of the reaction (pH) of groundwater was similar in all
piezometers, ranging within 7.00 and 7.46 (Tab. 2). Lower pH values were
determined in ground than in surface waters, in which the average reaction was 7.50.
According to this parameter, all the analysed water samples could be classified as
water purity class I.
Waste, next to wastewater, sewage and mineral fertilizers, is one of the major
factors responsible for degradation of water supplies, especially resources of
groundwater. Threats posed by a landfill to the surface of the earth or to air are just
as noxious, but they will appear only as long as a given landfill is operated. The
subterranean sphere, however, is threatened not only during the life of a landfill but
also when it has been closed, which makes landfills a danger to groundwater for tens
or even hundreds of years after their exploitation was terminated (GÓRSKI 2002,
KODA 2001, TWARDY, JAGU 2001 b, ZAŁATAJ 2001, KODA et al. 2006, 2007).
Deep groundwater collected from the area adjacent to the municipal waste
landfill in Brodnica, which is being exploited, were characterised by small variations
in the values of BOD5. The average value of this parameter in water sampled from
both piezometers P1 and P2 varied at a level of 1.50 mg O2 dm-3, which classifies
this water as belonging to water purity class I (Fig.10). Several-fold higher values of
BOD5 were determined in groundwater, where the average value of this parameter
was 4.33 mg O2 dm-3 in piezometer P3, water purity class III, and 17.00 O2 mg dm-3
in piezometer P4, water purity class V. In the former case, higher values were
recorded in water samples collected in summer and lower – in water samples taken
in autumn and spring. Regarding piezometer P4, situated behind the landfill cap,
near the wastewater pumping station, very high variations in BOD5 were noticed,
oscillating from 4.00 to 36.00 mg O2 dm-3. The highest values were observed in
autumn, and the lowest ones in summer.
CODCr
60
-3
50
mg O2.dm
mg O2.dm-3
BOD5
40
35
30
25
20
15
10
5
0
VI 2004
VIII 2004
X 2004
40
30
20
10
0
III 2005
Key:
– Piezometer 1;
– Piezometer 2;
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates
Sampling dates
–Piezometer 3;
– Piezometer 4;
– Water pond
Fig. 10. Values of BOD5 and CODCr in subsurface and surface water
83
Surface waters sampled from the water pond showed the value of BOD5 equal 5.87
mg O2 dm-3, which corresponds to water purity class I.
During the whole study, deep water sampled from model piezometer (P1) was
characterised by low and only slightly varied values of CODCr (Fig. 10). The value
of this index enabled us to classify the water as belonging to water purity class I. On
the other hand, the average value of CODCr, determined in deep water collected from
piezometer P2, located at the animal asylum, was 4-fold higher compared to the
average value of this parameter in deep groundwater collected from model
piezometer P1. Based on the values of this index, the water should be classified as
water purity class IV. Higher values of CODCr were recorded in August 2004, and
lower – in March 2005. High values of CODCr in deep groundwater collected from
piezometer P2 suggest that the contamination of that water was affected by both the
landfill and other sources of pollution (the animal asylum). This is confirmed by the
lower values of CODCr found in shallower groundwater collected from piezometer
P3, which is 10 m west of the landfill. In that case, higher values of this parameter
occurred in summer 2004 than in spring 2005. A noticeable decline in the values of
CODCr was within the range of 17.77 and 6.00 mg O2 dm-3, which classifies these
waters as water purity class II. The ground waters sampled from piezometer P4,
located behind the landfill basin, about 8 m from the pumping station, were
characterised by much higher values of CODCr than ground waters from piezometer
P3. The average value of this parameter was 46.92 mg O2 dm-3, corresponding to
water purity class IV.
The average value of CODCr in surface waters sampled from the water pond was
51.00 mg O2 dm-3, which means that they belonged to water purity class IV.
The total organic carbon content in deep and ground waters was highly varied.
The average content of this component in waters sampled from the model
piezometer P1 was 24.04 mg C Ca dm-3, which classifies them as belonging to water
purity class V (Fig. 11). Significant changes in the total organic carbon content were
observed between the sampling dates. In waters collected in summer and autumn the
organic carbon content was much lower than in the spring samples. A high content
of this component in deep waters collected from model piezometer P1 situated over
the inflowing waters, is not a measure of the effect of the landfill on the quality of
waters but indicates some pollution from other sources in the area where
the piezometer is located.
The deep waters collected from piezometer P2 situated over the water flowing
away from the landfill, regarding the total organic carbon content, were classified as
water purity class IV. However, it was impossible to state firmly that the poor
quality of these waters was caused exclusively by the proximity to the landfill.
A series of analyses seems to imply that other pollutants, from the area where the
piezometer is located, can be involved. Moreover, among the parameters most
highly exceeded there were the ones which did not reach high values in ground
waters sampled from piezometers P3 and P4, situated in close proximity to the
landfill. Pollution of deep waters sampled from piezometer P2 is therefore caused
jointly by the landfill and animal asylum.
The average content of organic carbon in ground waters collected from
piezometer P3 was 9.94 mg C dm-3, which corresponds to water purity class IV; in
84
ground waters collected from piezometer P4, located behind the landfill cap, it was
20.47 mg C dm-3, which means they belonged to water purity class V. In both cases,
the highest values were observable in spring and the lowest ones – in summer.
The total content of organic carbon in surface waters from the water pond equalled
36.44 mg C dm-3, which corresponds to water purity class V (Fig. 11).
The actual threat to ground waters depends not only on the amounts of waste
deposited but also on its physicochemical properties, such as water solubility,
toxicity, capability of water soluble substances, once they have entered ground
waters, to undergo self-purification processes. The extent of threat to ground waters
is also dependent on hydrogeological conditions near the landfill. How fast
pollutants will spread in ground waters depends on such factors as the volume and
quality of leachate, purifying properties of the aeration and saturation spheres, flow
properties (hydraulic slope and thickness of strata) which condition the speed and
intensity of flow, type of ground in the layer above the water table and in the aquifer
(BŁASZYK, BYCZYSKI 1986, VADILLO in. 1999, SICISKI, MYKÓW 2000,
SZYMASKA-PULIKOWSKA 2001a,b, KLOJZY-KACZMARCZYK et al. 2003).
70
mg..dm-3
60
50
40
30
20
10
0
VI 2004
VIII 2004
X 2004
III 2005
Sampling dates
Fig. 11. Content of total organic carbon in subsurface and surface water
Key:
– Piezometer 1;
– Piezometer 2;
–Piezometer 3;
– Piezometer 4;
– Water pond
In the analysed deep waters sampled near the landfill in Brodnica, the average
value of proper electrolytic conductivity was between 730.2 µS cm (P1) and 712.3
µS cm (P2), which corresponds to water purity class II (Tab. 3). In both cases,
significant differences were determined in values of this parameter between
particular water sampling dates. The highest value of proper electrolytic
conductivity appeared in the first three months and the lowest – in the last three
months of a year.
The average value of proper electrolytic conductivity in groundwater was 780.0
µS cm (P1) and 1,444.8 µS cm (P2), which classifies this water as water purity class
II and III, respectively. In both cases, the highest values were recorded in October
2004 and the lowest ones – in March 2005. The differences in the recorded values of
proper electrolytic conductivity between the sampling dates were significant (Tab.
3).
The water sampled from the water pond demonstrated proper electrolytic
conductivity around 681.00 µS cm, which means it belonged to water purity class II.
85
One possible measure of the influence the landfill has on the hydrosphere is the
increase in concentration of water soluble substances in ground and surface water.
Our analyses of the chemical composition of water samples collected from the four
piezometers near the landfill in Brodnica indicate certain pollution of ground water:
both deep and subsurface one. This pollution consisted of the increased
mineralization of the water and elevated values of such indices as the concentration
of chlorides and sulphates. Similarly to the parameters discussed earlier, it is not
possible to state firmly whether this situation was an effect produced solely by the
landfill. Other factors may have been involved, which is suggested by raised values
of many parameters found in deep groundwater collected from model piezometer
P1. A possible example is the content of soluble substances, which would enable us
to classify the water as water purity class III. Similar levels of soluble substances
were found in deep groundwater sampled from piezometer P2, located on the
premises of the animal asylum and in groundwater collected from piezometer P3.
The groundwater collected from piezometer P4, situated close to the leachate
pumping station, contained the highest concentration of soluble substances among
all the analysed water samples. Based on the average content of soluble substances,
the water from that piezometer belonged to water purity class IV.
Table 3
Parameter
Proper
electrolytic
conductivity
LSD0.05
Dissolved
substances
LSD0.05
Sulphates
LSD0.05
Chlorides
LSD0.05
Calcium
LSD0.05
Magnesium
LSD0.05
Values of salinity indices for subsurface and surface water
Piezometers
unit
1
2
3
4
Deep groundwater
Groundwater
µScm
mgdm-3
mg SO4-2dm-3
mg Cl-dm-3
mg Ca+2dm-3
mg K+dm-3
*730.20
** 677.4789.0
19.98
498.00
448.00 594.00
24.34
70.45
61.45 77.91
7.71
44.32
39.05 53.96
7.72
122.24
120.24 124.25
3.78
17.63
13.98 24.32
6.43
712.30
602.00 805.70
15.70
557.00
466.00680.00
16.06
58.36
38.00 94.00
7.88
39.01
26.63 53.96
5.37
82.16
68.14 96.19
5.97
15.50
12.16 17.02
2.41
*average
**fluctuations
86
780.00
755.80 817.00
15.23
588.70
511.00 690.00
25.58
89.79
73.55 119.00
13.44
14.55
12.42 17.04
4.07
137.61
132.26 148.30
8.12
29.39
22.50 34.05
6.59
1445.05
1019.60 1828.00
30.22
897.00
760.00 1032.00
19.51
40.12
21.79 - 58.00
Surface
water
(water
pond)
681.00
463.00
48.00
6.20
118.48
85.20 - 170.40
49.70
10.23
137.78
94.19-158.32 -
106.21
8.76
49.55
27.97 - 70.53
17.02
6.88
The average content of soluble substances in water samples obtained from the
water pond was 463.00 mg dm-1, which corresponds to water purity class I.
Migration of pollutants to groundwater becomes a health hazard when it is used
as a source of potable water. The rate and direction of the migration of pollutants to
groundwater are mainly conditioned by geological factors. In sedimentary rocks
(e.g. limestone, sandstone, dolomite), groundwater tend to flow along bedding
planes. The channels created by flowing water in such rocks enable water to travel
over relatively big distances without any changes in the concentration of pollutants.
Metamorphic rock, on the other hand, such as slate, enable polluted groundwater to
travel fast along fracture zones (BŁASZCZYK, GÓRSKI 1996, ALLEN 1999, VADILLO
et al. 1999, GOLIMOWSKI, KODA 2001, KRYZA-CHUDY 2003, MOCZULSKA 2006a,b,
KŁACZKO, SZYMASKI 2007).
The content of sulphides and chlorides in deep groundwater sampled from the
model piezometer and the piezometer at the animal asylum was within the range that
would classify it as belonging to water purity class I. The quality of groundwater
collected from piezometer 3 and of surface water from the water pond was similar
with this respect. The values of these parameters determined in the groundwater
sampled from the piezometer located near the leachate pumping station were higher,
suggesting that the water belonged to water purity class II. Variations in the content
of sulphides and chlorides in water samples from each of the piezometer were
relatively large and highly significant differences were noticed at most of the
sampling collections. Higher values of these ions were recorded in spring than in
summer and autumn.
The content of calcium and magnesium in groundwater samples collected from
the piezometers were typical of water representing water purity class I and II. The
content of calcium ions in deep groundwater oscillated around 68.14 and 124.25 mg
Ca+2 dm-3, whereas that of magnesium ions ranged between 12.16 and 24.32 mg
Mg+2 dm-3. The average content of calcium in this water was higher by 9.52 to 49.52
mg Ca+2 dm-3 than in leachate. Thus, it can be assumed that the concentration of
calcium ions in deep groundwater was conditioned not only by the influence
produced by the landfill but also by other sources of pollution nearby. In the
groundwater collected from piezometers P3 and P4, there were evident oscillations
in the concentrations of both ions, reaching an amplitude of tens of mg dm-3, with a
clear decreasing tendency for the content of calcium in water samples collected in
spring versus samples obtained in summer and winter.
In the surface water, the average content of calcium corresponded to the level
assigned to water purity class III, while that of magnesium would classify the water
to water purity class I.
Our study has demonstrated that the negative influence of well designed and
properly maintained landfills on the hydrosphere can be greatly reduced by using
adequate barriers (seals). However, groundwater and surface water need to be
constantly monitored, in order to prevent potential contamination, which landfills
can cause due to uncontrollable migration of leachate.
87
Summary
Municipal waste landfills are among the objects claimed to exert adverse
influence on the natural environment, mainly on the aquatic-terrestrial environment.
Landfills can cause very strong pollution of the hydrosphere with a variety of
components, more often than not toxic ones. Under favourable hydrogeological
conditions, leachate, produced by an operating landfill, can travel over large
distances and pose a threat to subsurface water, and consequently local sources of
potable water. Thus, constant monitoring of pollutants which can escape to the
hydrosphere both while a landfill is operated and after it has been closed down, is
necessary. It is also necessary to control the efficiency of the applied sealing systems
and potential faults of the used insulation. With such constant monitoring of the
effect produced by the landfill on the natural environment the risk of leachate
permeating to the aquifer is much lower. This is confirmed by the results of the
authors’ own studies, which demonstrated that although the examined landfill had
insulation barriers, it could be a certain risk to the aquatic and terrestrial
environment, as was made evident by the differences in the physical, oxygen and
salinity indices determined in leachate, subsurface and surface water. Based on the
results of our determinations, it was not possible to state firmly that the quality of
groundwater from a given piezometer (including the model one) corresponded to
one class of water purity. In some cases, the values of the physical, oxygen and
salinity indices exceeded the permissible values of water purity class IV or V.
However, it is not possible to state firmly if this situation was a result of the
migration of pollutants from the landfill. Identification of sources of components
degrading the quality of subsurface water is rather difficult due to the fact that there
are several potential points of pollution, e.g. the animal asylum, arable fields. The
effect of the landfill on the water physical, oxygen and salinity indices depended on
the sampling site. Samples collected from the cross-section above the landfill were
characterised by a lower content of chemical substances that the ones sampled at the
same time from the cross-section below the landfill. This tendency held true for
nearly all of the analysed components (pH, Ca, Mg, Cl-, SO42-). The reaction of the
analysed water samples was in most cases alkaline, with small variation.
Comparing the water samples taken from the piezometers drilled along the
direction of groundwater flow (behind the landfill basin and at the animal asylum), it
was found out that the further away from the landfill, the lower the values of the
parameters. There was only one exception to this rule, namely sulphates, whose
concentration was not correlated with the distance to the landfill.
The results of our study prove that rational management and exploitation of a
landfill, proper sealing to the substrate bed and regular monitoring, one can limit the
negative influence of a landfill on the environment.
88
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1
Wiera Sądej
Department of Environmental Chemistry,
Plac Łódzki 4, 10-727 Olsztyn, POLAND
2
Zbigniew LuliĔski, 2Janusz Posłuszny
Brodnica Municipal Management Company Ltd.
ul. Gajdy 13, 87-300 Brodnica, POLAND
[email protected], jposł[email protected]
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94
CHAPTER VI
THE EFFECT OF THE MUNICIPAL WASTE LANDFILL
ON THE HEAVY METALS CONTENT IN SOIL
Introduction
The factors responsible for the degradation of the soil environment include an
excessive cumulation of heavy metals which contain all metal elements with atomic
mass higher than that of calcium and density higher than 5 gcm-3 when exceeding
toxic concentrations (KOC 1994, SANECKI 1995, KABATA-PENDIAS, PENDIAS 1999,
ROSIK-DULEWSKA 2007). They occur as natural components in nature, but belong to
particularly dangerous elements, which create a potential hazard to the biological
environment and affect human health. The elements of a very high degree of risk
include, but are not limited to, cadmium, lead, copper and zinc, and those
of a medium degree of risk – arsenic (KOC 1994, KABATA-PENDIAS, PENDIAS 1999,
GAMBU, GORLACH 2001a). Harmfulness of heavy metals occurs sooner in animal
organisms than in plants, because the safe level for a plant is often toxic in the case
of its use as fodder or human food (KOC 1994, GAMBU, GORLACH 2001a,
ZGNILICKA 2002). In addition, in the environment they are often susceptible to
bioaccumulation, and in living organisms they are easily absorbable from the
alimentary canal, permeate the biological barrier which is the blood and brain, form
connections with sulphohydryl groups of proteins and damage the nucleic acids
chain.
One of the major and still topical issues is the estimation of the effect of various
factors and processes on the soil quality. Unfavourable changes in the physical,
chemical or biological soil properties may result not only in a decrease of its
fertility, but they can even totally exclude it from production. Their cumulation in
the soil surface layer is particularly noteworthy (BIERNACKA, MAŁUSZYSKI 2007;
NIED
WIECKI et al. 2007). The presence of metals such as lead, cadmium, and
mercury plays a great role, particularly in the case of their washing out to ground
waters in an amount that would threaten the quality of potable water and create a
hazard to human and animal health (TERELAK et al. 1995, GAMBU, GORLACH
2001a, KARWACZYSKA et al. 2005).
In terms of the contents of some compounds and heavy metals, human activity
has the highest effect on soil formation, natural conditions (mother-rock, climate,
landform features) being of secondary importance (DOMSKA et al. 2005, OLEKÓW
2007).
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The spreading of heavy metals such as copper, zinc, arsenic, cadmium and lead,
among others, is caused particularly by the chemical industry (Cu, Zn, As), artificial
fertilizers industry (Cu, Zn, Cd, Pb), cellulose and paper industry (Cu, Zn, Pb),
petroleum refineries as well as metallurgy and ferrous metallurgy (Cu, Zn, Cd, Pb),
glass-making, ceramic and cement industry (Cu, Pb) and power stations (elements
occurring in fuels). A great influence on the cumulation of heavy metals in soils is
also made by the location of industrial plants, motorization and herbicides, fertilizers
and waste used for soil fertilization (SANECKI 1995, SIUTA 2000, GAMBU,
GORLACH 2001b, ZGNILICKA 2002, KUSZA, CIESIELCZUK 2007, ROSIK-DULEWSKA
2005, NIEWIADOMSKI, TOŁOCZKO 2005, KARCZEWSKA, KRÓL 2007, MEDYSKA,
KABAŁA 2007). Soil contamination with heavy metals has also been reported in
urban areas of a high degree of urbanization, located close to industrial plants and
transport routes (LASKOWSKI, TOŁOCZKO 1995, KARCZEWSKA 2002, KUSZA,
CIESIELCZUK 2007, OLEKÓW 2007, PLAK 2007).
Although soil contamination with heavy metals is mostly caused by industrial
activity and coal or oil burning (SIUTA 2000, ZGNILICKA 2002, KUSZA,
CIESIELCZUK 2007), solid, liquid or gaseous contaminants can also get into the soil
from post-flotation and municipal waste landfill and mining activity (SANECKI 1995,
GAMBU, GORLACH 2001b, ROSIK - DULEWSKA 2007, NIEWIADOMSKI, TOŁOCZKO
2005, KARCZEWSKA, KRÓL 2007, MEDYSKA, KABAŁA 2007). However, some
authors (OWCZARZAK, MOCEK 2004, DOMSKA, RACZKOWSKI 2008) do not indicate
any unfavourable effect of brown coal opencast mines located on autogenic soils
on the availability to plants or on the content of nutrients in the soil.
As the civilisation develops and large population centres increase, the process
of continuous cumulation of waste in industrial and municipal landfills progresses.
This waste can also be a hazard to the environment as a result of a release
of numerous components from them, including heavy metals, through dusting, flow,
washing out or ignition and smoking (KOC 1994, KABATA-PENDIAS, PENDIAS 1999,
ROSIK-DULEWSKA et al. 2008). The qualitative composition of municipal waste
consists of flammable and non-flammable waste. The former comprises organic
waste, paper, fabrics, plastics, leather and rubber, while the latter – metals, glass and
ceramic goods. Municipal landfills located outside urbanized areas and illegal
rubbish dumps can be a source of contamination not only of adjoining farm land but
also of water poisoning in addition to the fact that they occupy another area, often at
the cost of agriculture or forestry (SZYMASKA, PULIKOWSKA 2003, KARCZEWSKA,
KRÓL 2007, ROSIK-DULEWSKA et al. 2008).
Waste is one of the major problems of environmental protection because
it creates hazard to all environmental spheres – lithosphere, hydrosphere,
atmosphere and biosphere (KARWACZYSKA 2001, SZYMASKA-PULIKOWSKA
2003, NIED
WIECKI et al. 2007, ROSIK-DULEWSKA 2007, ROSIK-DULEWSKA et al.
2008). A lot of waste management regulations have been introduced in Poland
recently, but its state is still unsatisfactory. It results, among others, from the
uncontrolled composition of dumping grounds, in which one can come across not
only debris, household appliances or electronic equipment but also hazardous waste
(remains of electrolytes, paint, lacquer, pigments, anti-corrosion agents, seed
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dressings, solvents, herbicides, batteries and overdue pharmacological agents, ash
from individual heating systems as well as organic fractions showing considerable
abilities to cumulate heavy metals (GAMBU, GORLACH 2001A; KARCZEWSKA 2002,
NIED
WIECKI et al. 2007).
The purpose of this research was to determine the effect of a municipal waste
landfill on the cumulation of copper, cadmium, lead and arsenic in the soil
of an adjoining area and to estimate risks connected with their contents and
distribution.
Research conditions
The investigation was carried out in 2008 in the area around the municipal waste
landfill of the town and commune of Wgorzewo near Czerwony Dwór. It occupies
a territory of the total area of 3.6 hectares and has been exploited since 1996.
The basic technical and exploitation parameters of the landfill amount to: upper area
- 4950 m2, bottom area - 1700 m2, total volume – 20000 m3, target dumping datum –
158,20 m above sea level. The quantity of the waste cumulated so far is estimated
at 99 thousand m3. The waste deposited at the landfill is not separated and mostly
consists of household and building waste (ORYCZAK 2008). The area adjoining the
landfill has been additionally secured by a 10-metre wide tree planting strip.
Soil was sampled from the surface soil layer on the southern side of the stockpile
at a distance of 5, 10, 20 and 30 m. In mean soil samples (formed by mixing
10 individual samples), the granulometric composition was determined with the
Bouyoucos aerometric method modified by Casagrande and Prószyski, pH –
electrometrically in 1 mol⋅dm-3 KCl, the organic carbon content – according to
Tiurin, the phosphorus and potassium content – with the Egner-Riehm method,
magnesium – with the Schachtschabel method, and the contents of copper, zinc,
cadmium, lead and arsenic – with the atomic absorption spectrometry technology
after a sample mineralization using nitric acid and hydrochloric acid.
The data from particular sampling sites did not show any significant variations,
therefore the findings are presented in tables as mean values.
The significance of variations has been calculated using the Tuckey’s test, at the
level of p=0,05.
Physical and chemical properties of the soils under study
It has been found out that the soil samples taken from the area adjoining the
landfill had similar physical and chemical properties typical of soils of good
agricultural usefulness. It was proved by their granulometric composition of light
and medium loam, acidity of pH values from 7.0 do 7.2, and the humus content from
0.4 to 0.5% (Tab.1).
The soil conditions prevailing in the neighbourhood of the landfill did not show
properties which would favour excess cumulation of heavy metals.
Decisive role not only in the contents of mineral components in the soil, but also
their mobility and availability to plants is played by not only the soil acidity, but also
97
the type and properties of the soil, including the granulometric cmposition and
humus content (KABATA-PENDIAS, PENDIAS 1999, KARCZEWSKA 2002).
Relatively high soil acidity (close to neutral) during the investigation, like
in previous investigations by DOMSKA et al. (1996) and DOMSKA and WOJTKOWIAK
(2000) was probably conditioned by its granulometric composition and applied
agricultural technology. This acidity was not favourable for a high mobility of heavy
metals, thus limiting the penetration of contaminants into the plants’ root system.
Setting in motion of the forms which are easily available to plants occurs with acid
reaction; under such conditions, there is generally a larger content of heavy metals
(GAMBU, GORLACH 2001a). Soil graining typical of sands and sandy clay indicates
a possibility of an occurrence of water permeability and easy migration
of contaminants into the soil profile. In soils with a very small fraction of clay
and a high content of organic carbon there are favourable conditions for potential
accumulation of contaminants only in organic and mineral complexes (KUSZA,
CIESIELCZUK 2007). Humus, in turn, shows high abilities of heavy metals
absorption, which makes it more difficult to wash them out of the soil (MEDYSKA,
KABAŁA 2007).
Table 1
Some physical and chemical soil properties
Place
of sampling
1
2
3
4
Acidity
(pH in 1n KCl)
7.2
7.2
7.2
7.0
Granulometric
composition
light loam
light loam
light loam
medium loam
Humus
%
0.5
0.5
0.4
0.4
Distance from landfill: 1 – 5m, 2 – 10 m, 3 – 20 m, 4 – 30 m.
The contents of available phosphorus (54.4-61.1 mg⋅kg-1), potassium (124.5132.8 mg⋅kg-1) and magnesium (57.7-61.5 mg⋅kg-1) around the landfill corresponded
to the average soil abundance in relation to these nutrients (Tab.2).
Table 2
Available phosphorus, potassium and magnesium content (mg⋅kg-1 d.m.)
Place of sampling
1
2
3
4
LSD p=0,05
∗see tab. 1
Phosphorus
54.5
56.7
56.7
61.1
2.5
Potassium
124.5
128.6
124.5
132.8
4.6
Magnesium
57.7
60.1
57.9
61.5
3.9
A slightly larger cumulation of the analysed forms (by 4.4-6.6, 4.2-8.3 and 1.43.8 mg⋅kg-1, respectively) occurred in the margin of the area under study
in comparison with the terrain located directly next to the landfill. It was probably
98
connected with a variation of the granulometric composition and absorbing capacity
of the soils studied (light and medium loam). In the available literature, phosphorus
is pointed to as a factor which may contribute to changes in zinc or cadmium
cumulation in soil, but only in the case of applying large doses of phosphorus
fertilizers (TERELAK et al. 1995).
Copper and zinc
Copper and zinc belong to components of high biological importance, necessary
for a proper functioning of an organism. We count them among heavy metals,
affecting unfavourably the growth and yield of plants, when toxic concentrations are
exceeded (KOC 1994, GAMBU, GORLACH 2001A, KABATA-PENDIAS, PENDIAS
1999). In Polish soils, copper occurs in amounts from 0.2 to 293.3 mg⋅kg-1 forming
low-mobile connections in the form of carbonate and sulphate and with organic
matter and clay minerals. In the case of zinc, its content ranges from 0.5 to 1725.0
mg⋅kg-1 and is closely connected with the reaction, because it forms compounds of
high solubility, which grows with acidification and decreasing soil absorption ability
(TRELAK et al. 1995). According to research conducted by the Institute of
Cultivation, Fertilization and Soil Science in Puławy together with Regional
Chemical and Agricultural Stations, the permissible total content in the surface soil
layer in relation to copper amounts to 25 to 74 mg⋅kg-1, and that of zinc – from 80
to180 mg⋅kg-1 (GAMBU, GORLACH 2001b). No unfavourable effect on particular
ecosystems is revealed below these numbers.
The copper content in the studied area's soil near the municipal waste landfill
was of little variation (showed no significant differences) and ranged within the
values corresponding to good soil abundance for this element from 4.93 to 6.15
mg⋅kg-1 of dry matter (Tab.3).
Table 3
Copper and zinc content in landfill site area (mg⋅kg-1 s.m.)
Cu
5.02
6.15
4.93
5.61
1.30
Place of sampling∗
1
2
3
4
LSD p=0,05
∗ see tab. 1
Zn
45.71
44.25
43.28
38.24
2.56
More varied results, but not exceeding (just like in the case of copper) a good
soil abundance, were obtained when analyzing the cumulation of zinc in the studied
area. More zinc, i.e. from 43.28 to 45.71 mg⋅kg-1 of dry matter occurred in the
vicinity of the landfill at the distance of 5, 10 and 20 m, while at a farther distance
(30 m) there was only 38.24 mg⋅kg-1. The variations ranged from 5.04 to 7.47
mg⋅kg-1 of dry matter, therefore, they were low and probably resulted from the
99
difference in the granulometric composition of the soil and a higher content of
phosphorus at a farther distance of the sampling place from the landfill. In
comparison with the investigation of BIENIEK (2005) conducted in soils of a similar
granulometric composition (light loam) in the vicinity of Olsztyn, the cumulation of
copper and zinc near the municipal waste landfill in the vicinity of Czerwony Dwór
was much lower. However, like in the quoted research, the values concerning the
soil contents of the mentioned metals stayed within the limits of their natural
contents in the soil and significantly below the quality standards of soils and land
specified for agricultural land in the regulation of the Minister of Environment
(2002).
According to NIED
WIECKI et al. (2007), uncontrolled waste landfills located in
sandy areas may contaminate the soil's surface layers with heavy metals, particularly
copper and zinc. However, the quoted authors, as well as SZYMASKA and
PULIKOWSKA (2003), maintain that municipal waste is characterized by a very
varied content of heavy metals, and the intensity of environmental changes
occurring under their influence is connected with the quality of the waste, frequency
and time of storage and supply of the dump with illegal domestic sewage discharge,
particularly on uncontrolled dumps.
In the research of OLEKÓW (2007) conducted in the allotments of Wrocław
in the vicinity of industrial plants and transport routes, a higher contamination of
soils, among others – with zinc and copper, was proved, most of the soil being
contaminated with zinc (around 90%), while the contamination with copper was
higher (corresponding to scale I-V). Investigations of other authors (GAMBU,
GORLACH 2001a) show that the excess of copper in the soil occurs mostly in areas
contaminated by the copper industry and as a result of contamination with herbicides
containing copper, and in the case of zinc – as a result of coal and waste burning and
due to a storage of metal industry goods, or, like in the case of copper, is caused by
herbicides.
Lead, cadmium and arsenic
The lead content in the Polish soils amounts to 0.1 to 992.5 mg⋅kg-1 of dry
matter and is mostly dependent on the mineralogical and granulometric composition
and the origin of mother rocks. Its availability is also dependent on the soil’s
reaction and, to a lower degree, on humus and the soil absorption ability (TERELAK
et al. 1995). Besides, it is less mobile than zinc and cadmium, because it is included
in slightly soluble minerals. Environmental conditions, factor analysis (mother rock,
the character and causes of regional content differentiation and some soil properties)
and the dependence between their actual content in soils and the expected range
including numerical values after an exclusion of extreme observations resulting from
significant analytical errors or accidental contamination are also noteworthy,
particularly in the case of lead, cadmium and arsenic (DUDKA 1992).
The lead content in the surface soil layer in the area under study ranged from
10.54 to 15.57 mg⋅kg-1 of dry matter, therefore, not only did it exceed the natural
100
values but it was also lower than the permissible limit of 40 mg (GAMBU,
GORLACH 2001b).
The most lead was in the immediate vicinity of the landfill, i.e. at the distance
of 5 m (Tab. 4). As the distance from the landfill increased, the content of lead in the
soil decreased, corresponding to the values of 14.23, 12.71 and 10.54 mg⋅kg-1
of dry matter. Although the main source of environmental contamination with lead
is metallurgy and transport, it can also be released from the waste in the form
of utensils, packages and production equipment (KOC 1994). In the investigation of
KABAŁA (1995), variations in the properties of the analysed soils had a limited
range, while significant correlations between the contents of lead, zinc and copper
and the content of organic carbon and soil acidity occurred. In the findings of
BIENIEK (2005), in soils of physico-chemical properties similar to those in the
vicinity of Czerwony Dwór there was also a very low content of lead. Contrary to
this investigation, OLEKÓW (2007) proved that soils in the vicinity of Wrocław
exposed to the impact of industrial plants and transport routes were medium
contaminated with lead in 70% in the scale from degree I to III. In the research of
LASKOWSKI and TOŁOCZKO (1995), carried out near urban and industrial
agglomerations, it was proved that the concentration of lead showed a larger
dependence on the type of mother rock, granulometric composition or the content of
organic substance than on the location of research sites in the field. However, the
authors concluded that even a low content of heavy metals can be dangerous with a
severe acidification of soils due to a large share of soluble forms in their total
content.
Table 4
Cadmium, lead and aresenic content in landfill site area (mg⋅kg-1 s.m.)
Place of sampling∗
1
2
3
4
LSD p=0,05
∗see tab. 1
Pb
15,57
14,23
12,71
10,54
1,51
Cd
0,14
0,12
0,15
0,13
0,03
As
1,00
1,35
0,65
0,75
0,30
In Polish soils the content of cadmium ranges from 0.01-24.75 mg⋅kg-1 of dry
matter, 0.22 mg⋅kg-1 on average, while the permissible content in soil surface layers
ranges from 1 to 3 mg⋅kg-1 of dry matter (TERELAK et al. 1995, GAMBU, GORLACH
2001b). At the same time, this element shows a high mobility as a soil environment
component which is easily taken in by plants (TERELAK et al. 1995). Its cumulation
in soil, exceeding natural values, can be connected with the character of the
basement complex, sewage sludge application, or overfertilization with phosphorus,
etc. However, dust emissions from non-ferrous metallurgical plants and dust from
scrap materials dumps, carried by wind, constitute the major source of soil
contamination with cadmium. Environmental contamination with cadmium can also
101
be caused by the impact of municipal landfills containing industrial and energetic
waste, paint and lacquer residues (KOC 1994, GAMBU, GORLACH 2001a).
The content of cadmium in the soil of the studied landfill site near Czerwony
Dwór ranged within the limits of the natural (Oº) content and much below the soil
and land quality standards determined for agriculturally used land in the regulation
of the Minister of Environment (2002). It was very similar in the whole area under
study and ranged from 0.12 to 0.15 mg⋅kg-1 of dry matter (tab.4). Contrary to these
findings, in the investigations by OLEKÓW (2007) carried out in allotments
in Wrocław near large industrial plants and transport routes, soil contamination was
proved not only with zinc, copper and lead, mentioned before, but also with
cadmium in 87% of the studied soil, the contamination level being determined as
severe (degree I-V).
In the investigation of ROSIK-DULEWSKA and KARWACZYSKA (2004),
attention was focused on contamination with heavy metals in black earth limestone
soil lying within the reach of the impact of the ”Grundman” waste landfill in Opole.
It was found out that after a 50-year period of the use of the landfill, the content of
cadmium and lead was higher than that regarded as natural by IUNG (class I) and
stayed within the class II standard. The highest contents of heavy metals occurred in
the soil at the depth of up to 30 cm in the form of chemically stable and biologically
inactive compounds – bounded with ferric oxide, manganese oxide and organic
substance. This means that their availability to plants, soils and waters is
considerably limited.
Arsenic belongs to elements which are very common in the environment (PLAK
2007). It is used in various branches of industry and in agriculture as a component of
pesticides. Besides, it appears in small amounts in all food agents, and in larger
amounts in sea products (KOC 1994). Anthropogenic sources include, apart from
pesticides containing arsenic, also agents for wood conservation or production of
paints and lacquers, but it is non-ferrous metallurgy, particularly copper metallurgy
and liquid and solid fuel burning, which creates the greatest hazard (KABATAPENDIAS, PENDIAS 1999, PLAK 2007, ROSIK-DULEWSKA 2007). In the soil, arsenic
is absorbed by organic substances, ferric oxides, aluminium hydroxides and
manganese compounds, and its content is highly variable and ranges from 0.1 to 95
mg⋅kg-1. In soils originating from sedimentary rocks it remains at the level of 20-30
mg⋅kg-1. Arsenic appears in larger amounts in clayey soils and soils rich in organic
components, ferric, aluminium and phosphorus compounds, and in the region of the
metallurgic and chemical industry, and in large urban agglomerations its
concentration in soil can reach the values of as much as 2500 mg⋅kg-1 (KABATAPENDIAS, PENDIAS 1999).
The content of arsenic in the soil near the landfill under study stayed within the
permissible standards from 0.65 to 1.35 mg⋅kg-1 of dry matter (tab.4). However,
there was much more arsenic (around 1.5 to 2 times) at the distance of 5 and 10 m in
comparison with the remaining area under study (20 and 30 m from the landfill).
The findings indicate an impact of some waste contained in the landfill, according to
the observations of other authors (SANECKI 1995, GAMBU, GORLACH 2001b, PLAK
2007), but it does not create such a hazard as that in the investigation of MEINHARDT
102
(1995) carried out in Wrocław province and in the city of Wrocław. The author
assessed the degree of heavy metals contamination of soils with the granulometric
composition of light loamy sand with the humus content from 2.1 to 4.9% and
proved an increased content of zinc, lead, cadmium, nickel, mercury, arsenic,
sulphur, fluorine and PAH in the immediate vicinity of industrial plants, transport
routes and waste landfills (Czechnica thermal-electric power station and Siechnice
steel mill). According to this research, a high soil contamination with zinc and lead
resulted from, among others, a long-term impact of the steel mill on the
environment. In the investigation of KUSZA and CIESIELCZUK (2007), in turn, the
content of heavy metals, such as chromium, zinc, cadmium, copper, nickel, lead and
mercury in the areas of industrial plants of the Opole region showed their low effect
on the state of soils of the adjoining land. It was only in one case that a content
of lead exceeding the permissible value (56.83 mg⋅kg-1) according to soil and land
quality standards was found out. A low threat of external factors was also indicated
by GAMBU and GORLACH (2001b), who, on the basis of 3337 samples taken in the
province of Warmia and Mazury proved that in the total area of arable land the share
of soils with a natural content of heavy metals amounts to 91.5%, while soils with
the contamination degree from II to V constitute only 0.5%.
Summary
Based on the obtained data, after a 12-year's period of the use of the municipal
waste landfill near Czerwony Dwór, and in the light of standards and legal
regulations in force (Ordinance of the Minister of Environment of 9 September 2002
concerning soil quality standards and land quality standards, Journal of Law No.
165, item 1359 of 4 October 2002) and an assessment of heavy metals content in the
soil surface layer according to KABATA-PENDIAS (1999), it has been found out that
there was no exceedance of standards concerning the permissible content of the
studied elements (copper, zinc, lead, cadmium and arsenic) in the soil utilization
group B, including agriculturally used soils. The content of copper and cadmium in
the analyzed area was similar and ranged from 4.93 to 6.15 and from 0.12 to 0.15
mg⋅kg-1 of dry matter, respectively. A little more zinc (from 43.28 to 45.71 mg⋅kg-1
of dry matter) and lead (from 12.71 to 15.57 mg⋅kg-1 of dry matter) occurred in the
soil at the distance up to 20 m from the landfill, and arsenic (from 1.00 to 1.35
mg⋅kg-1 of dry matter) – closer to the landfill (at the distance up to 10 m). The
findings do not indicate any threat to the environment due to a cumulation of heavy
metals. In addition, they do not suggest any necessity of introducing restrictions in
farm production near a waste landfill.
103
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Chair of Agricultural Engineering and Raw Materials
ul. S. Okrzei 1A, 10-266 Olsztyn
e-mail: [email protected]; [email protected]
105
106
CHAPTER VII
Boena Cwalina-Ambroziak1, Jadwiga Wierzbowska2
EFFECT OF FERTILIZATION ON THE COMPOSITION OF
SOIL FUNGI COMMUNITY
Introduction
Fertility and, simultaneously, the productivity of soil are determined by, among
other things, the content of organic matter originating mainly from deficient farm
manure and liquid manure. Thus, an increasing interest may be observed in the
acquisition of organic matter from other sources, e.g. sewage sludge, municipal solid
wastes and municipal green wastes. Owing to a high content of organic matter
as well as macro- and microelement, they may be utilized for agricultural purposes
as composted organic fertilizers (SPYCHAJ-FABISIAK et al. 2002).
Natural and organic fertilization affects biotic relations in the soil, which is due
to an increased content of organic carbon in the soil, especially microbiological
carbon and, to a lesser extent, of nitrogen (LARKIN et al. 2006). As observed by
HOITINK et al. (1997), fertilization with compost should be adjusted to the contents
of macro- and microelements in the soil and to the requirements of plants, as an
excess of N, for example, has been found to promote the growth of pathogenic
factors, including: Erwinia amylovora and the genus Phytophthora. The high
concentration of ammonia N and a low C:N ratio in sewage were reported to
stimulate the development of fusarium diseases (KATO et al. 1981). Such elements
as: B, Cu, Pb, Mn, Zn, have also been implicated in affecting the structure of the soil
fungi community. PRATT (2008) claimed that in soil fertilized with organic wastes
of animal origin, as compared to the non-fertilized soil, the concentrations of P, K
and Na were significantly higher, whereas those of Mg, Cu and Zn were usually
lower and, finally, those of N, Ca, Fe and Mn – were rarely or never higher. That
author emphasized, however, that the composition of the soil fungi community
remained unchanged under the influence of the above fertilization. A similar opinion
was expressed by GÓRSKA and STPIE (2007), who claimed that the introduction of
organic additives to soil had no effect on the population numbers of hyphae fungi.
An opposite claim was made by AWAD and FAWZY (2004), who proved that
increasing doses of sewage sludge promoted the growth of bacterial and fungal
populations in soil. HOITINK et al. (1997) as well as WEYMAN-KACZMARKOWA et
al. (2002) were also convinced that composts and vermicomposts facilitated an
increase in the population numbers of soil microorganisms, thus enhancing their
107
activity and biodiversity. These authors additionally indicated the significance of the
type of organic fertilizer used and composting time in determining the structure of
the rhizosphere fungi community.
Changes in the counts and functioning of microorganisms in the soil
environment affect, among other things, plant resistance to diseases. The inhibiting
effect of manure and organic fertilizers on the growth of soil pathogens is relatively
well-documented in literature. Fertilization with bovine manure was reported to
diminish the population of Rhizoctonia solani (TSROR LAKHIM et al. 2001) and that
of Streptomyces scabies (LAZAROVITS et al. 2008) in soil, thus reducing infections
of potato tubers. The results of other investigations (GORODECKI, HADAR 1990) also
confirmed the suppressing effect of that fertilizer on the development of the
causative agent of black scurf of potato tubers (Rhizoctonia solani), and additionally
on Sclerotinia rolfsii. Organic fertilization in the form of composted plant wastes
inhibited the growth of soil pathogens (HADAR, MANDELBAUM 1986), whilst the
fresh plant wastes diminished infestation of solanacenous plants with Phytophthora
capsici, Alternaria solani and Septoria lycopersicae (KIM et al. 1997, MILLS et al.
2002) and that of pea with Aphanomyces euteiches (WILLIAMS-WOODWARD et al.
1997). LODHA and BURMAN (2000) noted a 20 – 40% reduction in the population of
Macrophomina phaseolina – a pathogenic papilionaceous plant – as affected by
fertilization with compost of plant waste. In turn, STONE ET AL. (2003) demonstrated
that organic additives (paper residues) composted with bark and those not subjected
to composting, inhibited the growth of soil pathogens (Pythium spp., Colletotrichum
lindemuthianum, Aphanomyces spp.). Numerous authors (SCHUELER et al. 1989,
DRAFT, NELSON 1996, HOITINK, BOEHM 1999) reported that composts
(e.g. composted household wastes) were likely to suppress the development of some
fungi-like organisms (Pythium spp., Phytophthora spp.) and potential pathogens
of the genus Fusarium. RINGER et al. (1997) proved that the examined types
of composts from household wastes inhibited infections with R. solani to the same
extent, but differentiated the intensity of seedlings blight by Pythium ultimum.
Composts based on municipal sewage were also found to be significant in plant
protection against soil pathogens (SERRA-WITTLING et al. 1996).
Organic fertilization evokes positive changes in the quantitative and qualitative
composition of a soil fungi community (SZCZECH 1999). A desirable phenomenon is
an increase in the count of beneficial bacteria antagonistic to pathogens. The organic
fertilizers applied may, thus, constitute potential biological protection for plants
against pathogenic factors.
In the suppression of pathogen development, great significance is ascribed to
fungi of the genus Gliocladium and Trichoderma. BULLOCK et al. (2002)
demonstrated a significantly higher count of fungi of the genus Trichoderma in the
soil subjected to organic than mineral fertilization. The above fungal species are
known for their lignolytic and cellulolytic properties. They were shown to intensify
biological processes in soil, thus increasing its phytosanitary status (ŁACICOWA,
PITA 1989, HOITINK, BOEHM 1999). Other works (BAKER, COOK 1974, NELSON
et al. 1983) report that fungi of the genus Trichoderma were colonizing sclerotia
of R. solani and S. rolfsii. In turn, CHRISTENSEN (1969) demonstrated that the
species T. harzianum was producing high quantities of CO2, ethanol and antibiotics,
108
which inhibited the growth of some fungal species, including those of the genus
Penicillium np. P. jaczewskii. High numbers of those saprotrophs of the genus
Penicillium were isolated from soil fertilized with compost from organic wastes by
DROZD ET AL. (1996). The high prevalence of those fungi in the natural environment
is explained by their high capability to adapt to environmental conditions, e.g. their
capability of exploiting various sources of food. SARAIVA et al. (2004) included
species of the genus Penicillium (apart from fungi of the genera Aspergillus and
Fusarium) in the group of microorganisms most frequently colonizing
the organically-fertilized soils.
In the present study, an attempt was made to determine the effect of different
organic fertilization compared to non-fertilized plots (control combination) and plots
with mineral NPK fertilization and fertilization with manure, on the structure
of a community of soil fungi. In addition, in vitro tests on PDA culture medium with
the addition of aqueous extracts from composts were applied to compute
the percentage index of growth inhibition of pathogen mycelium.
Study determinants
An exact field experiment was established in 2004 by the Department of
Agricultural Chemistry and Environment Protection at the Agricultural
Experimental Station in Bałcyny. Experimental plots with an area of 15m2
(at randomized complete block design, in three replications) were located on gley
luvisol soil (developed from light silty loam, complex 4 class III, characterized
by a high content of P, medium content of K and a low content of Mg, and
pH = 5.04). The following crop species were grown in a four-year rotation system:
commercial potato, spring fodder barley, winter rape and winter wheat.
The factor analyzed in the study was the type of organic fertilizer.
A phytopathological analysis was conducted over the first three years of the
experiment on the following plots: I. control (no fertilization), II. mineral NPK
fertilization, III. farm manure 10tha–1 , IV. farm manure 5tha –1*, V. “Dano”
compost 10tha –1 (compost from non-segregated municipal wastes, composted with
the “Dano” method), VI. „Dano” compost 5tha –1*, VII. green waste compost
10tha –1, and VIII. green waste compost 5tha –1.
The farm manure and composts from municipal wastes at the dose of 10t x ha–1
were applied in 2004 before potato planting (Jasia cultivar). Doses of mineral
fertilizers applied under potato were as follows: 150 kg N (34% ammonium nitrate),
65 kg P (40% superphosphate) and 166 kg Kha –1 (60% potassium salt). Mineral
fertilization on the NPK plot was applied exclusively before sowing. On the plots
with farm manure and sewage sludge, the fertilization with N was balanced to
150 kgha–1, depending on the content of total nitrogen in the fertilizers, and
completed after the main crop with ammonium nitrate. In 2005, only mineral
fertilization was applied under spring barley (Justyna cultivar): 90 kg N, 26 kg P and
100 kg Kha–1 (forms of fertilizers as above). After the harvest of spring barely and
before sowing winter rape, mineral fertilization was applied as follows: 120 kg N,
42 kg P and 134 kg K ha–1. Organic fertilization in a dose of 5 tha–1 was applied
only on plots: IV, VI and VIII. Supplementary fertilization with N up to
109
120 kgha–1 was balanced depending on the total nitrogen content of compost on the
above-mentioned plots.
In order to determine species and quantitative composition of fungi in the soil
from three sites on particular plots, constituting a given combination, soil samples
were collected at a depth of up to 10 cm. In the laboratory, the samples were mixed
and their 10-g portions were weighed into 250 ml flasks; 90 ml of sterile water were
added to the flasks which were then shaken for 20 minutes to reach a dilution
of 10–4. The culture of fungi was run on Martin’ medium at a temperature of 22ºC,
and fungal colonies grown after 5-day incubation were calculated. Results were
converted to grams of dry matter, whilst the colonies were inoculated onto agar
slants for microscopic identification of species.
A laboratory test was used to determine the effect of aqueous extracts from the
composts examined on the growth of potentially-pathogenic fungi: Botrytis cinerea,
Colletotrichum coccodes and those of the genus Fusarium (F. culmorum, F.
equiseti, F. oxysporum and F. poae). Isolates of the above species, from which
single-spore cultures were prepared for the study, originated from the experimental
soil. The aqueous extracts were prepared as follows: 2 g portions of dried material
were poured over with 100 ml of sterile water for 24 hours. After filtration, the
extracts were dosed in 2 ml portions on Petri dishes and poured over with 10 ml of
PDA medium with a temperature of 50ºC. Next, agar discs 5 mm in diameter
overgrown with 7 day mycelium of the pathogens examined were placed on the
solidified medium. The dishes with pathogen inoculum on the medium without the
aqueous extracts from composts served as a control. After 4 and 8 days, colonies
were measured alongside two perpendicular straight lines. The index of mycelium
growth inhibition was calculated from the formula: I = [(k - ) : k] x 100%,
where k and denote diameter of fungal culture in the control combination and in
the medium with compost extracts, respectively. The results obtained in the study
were elaborated statistically with the analysis of variance (STATISTICA® v.8. 200708) using the Duncan’s test to compare mean values.
Effect of mineral, natural and reduced fertilization on the composition
of a soil fungi community
Organic fertilization differentiated the number of colonies of the soil fungi only
to a negligible extent. The highest number of fungal colony forming units was noted
in the soil fertilized with “Dano” compost applied in doses of 5 tha-1 and was
significantly different as compared to the number of CFU in the soil fertilized with a
single dose of “Dano” compost (10 tha-1) and with green waste compost in both
variants of application (fig. 1).
The species composition of the soil fungi community in particular variants of
fertilization appeared to be more diversified. Amongst the isolated fungi, 49 species,
yeast-like fungi and asporogenous cultures were identified. The species of
potentially-pathogenic fungi identified in the study included: Botrytis cinerea,
Colletotrichum coccodes, Sclerotinia sclerotiorum and fungi of the genus
Aureobasidium (A. bolleyi and A. pullulans) and Fusarium (F. culmorum, F.
equiseti, F. oxysporum and F. poae). The highest prevalence of pathogens was noted
110
in the soil from the control non-fertilized plot (14% of all isolates – fig. 2); only in
that fertilization variant was their presence detected in all experimental years. In the
second year of the study, a high contribution in the fungal community was reported
for species of the genus Fusarium.
b
b
ab
ab
b
a
b
b
Dano 5 t
green waste green waste 5
10 t
t
5
4
3
2
1
0
control
NPK
manure 10t manure 5 t
Dano 10 t
Fig. 1. Number of fungal colony forming units per 1 g of soil (CFU x 105)
40
35
70
pathogens
30
25
Sclerotinia
sclerotiorum
Fusarium spp.
Aureobasidium
spp.
20
15
60
Trichoderma
spp.
saprotrophs
50
Penicillium spp.
40
Paecilomyces
spp.
Mucorales
30
20
10
Gliocladium spp.
10
5
0
saprotrophs
50
40
30
20
10
5
6
20
0
20
0
20
0
4
0
2004 2005 2006
pathogens
60
0
x from all years
a.
40
pathogens
saprotrophs
70
35
Trichoderma
spp.
Penicillium
spp.
Paecilomyces
spp.
Mucorales
60
30
Sclerotinia
sclerotiorum
25
Colletotrichu
m coccodes
20
15
Aureobasidiu
m spp.
10
50
40
30
20
0
0
2004 2005 2006
60
saprotrophs
50
40
30
20
10
5
pathogens
2004 2005 2006
10
0
x from all years
b.
Fig. 2. Fungi isolated from soil: a. without fertilization (control),
b. with mineral fertilization (%)
111
A smaller population of the pathogens was cultured in the soil fertilized with
NPK (10.6% - fig. 2b), with A. pullulans being the most frequently isolated species.
In soils subjected to organic fertilization, RITZ ET AL. (1997) reported an
increase in the count of bacteria, actinomycetes and fungi. They explained the
enhanced biological activity of the soil with an elevated content of soluble C in soil
and, to a lesser extent, with that of N. Especially favorable seem to be changes
affecting an increase in the population of beneficial microflora as specific, biological
protection of plants against pathogens (WIDNER ET AL. 1998, SZCZECH 1999). A
reduction was observed in the count of pathogens in the soil upon fertilization with
farm manure (Fig. 3 a, b) and organic fertilization (compost – Fig. 4 a-d) as
compared to the control variant and that with mineral fertilization. The contribution
of the pathogens in the fungal community ranged from 1.2% in the soil fertilized
with manure at a dose of 10 tha-1 to 9.1% in the variant with green waste compost
applied twice at a dose of 5 tha-1 each. In the other experimental variants, the
frequency of pathogen occurrence ranged from 5 to 10%. The species of pathogens
frequently isolated from the soil analyzed in the study included those mentioned
above and those belonging to the generea Aureobasidium and Fusarium, whereas the
less frequently isolated species included: B. cinerea, C. coccodes and S.
sclerotiorum.
30
pathogens
25
20
Fusarium
spp.
15
pathogens
saprotrophs
70
10
60
Trichoderma spp.
60
50
Penicillium spp.
50
Paecilomyces spp.
40
40
30
5
10
0
0
30
Mucorales
20
20
Gliocladium spp.
10
0
2004 2005 2006
2004 2005 2006
saprotrophs
x from all years
a. dose of 10 tha-1
pathogens
25
saprotrophs
pathogens
70
saprotrophs
60
20
Sclerotinia
sclerotiorum
15
Aureobasidium
spp.
10
Trichoderma spp.
50
50
Penicillium spp.
40
30
Paecilomyces spp.
20
5
0
2004 2005 2006
40
30
20
Mucorales
10
0
60
2004 2005 2006
10
0
x from all years
b. dose of 5 tha-1.
Fig. 3. Fungi isolated from the soil fertilized with farm manure (%)
112
pathogens
30
70
25
60
20
50
30
Botrytis cinerea
20
10
5
Fusarium spp.
Botrytis cinerea
10
Aureobasidium
spp.
5
Mucorales
10
x from all years
70
60
50
40
30
20
10
0
20
15
30
20
saprotrophs
pathogens
25
Paecilomyces spp.
Gliocladium spp.
2004 2005 2006
a. Dano at a dose of 10 tha-1
30
40
0
0
2004 2005 2006
60
Penicillium spp.
40
15
0
Trichoderma spp.
50
Aureobasidium spp.
10
patogeny
saprotrofy
saprotrophs
0
2004 2005 2006
pathogens
saprotrophs
Trichoderma spp.
60
Penicillium spp.
50
40
Paecilomyces spp.
30
20
Mucorales
10
Gliocladium spp.
2004 2005 2006
0
x from all years
b.Dano at a dose of 5 t x ha-1
saprotrophs
pathogens
pathogen
30
70
25
saprotrophs
60
60
20
40
40
Paecilomyces spp.
30
Fusarium spp.
10
20
5
50
Penicillium spp.
50
Sclerotinia
sclerotiorum
15
30
20
Mucorales
10
10
0
2004
2005
0
2006
pathogens
x from all years
15
saprotrophs
100
Sclerotinia
sclerotiorum
Fusarium spp.
10
80
70
60
Penicillium spp.
50
Paecilomyces spp.
Mucorales
20
10
0
20
04
20
05
20
06
saprotrophs
40
30
30
0
60
50
40
5
pathogens
Trichoderma spp.
90
25
20
0
Gliocladium spp.
2004 2005 2006
c. green waste compost at a dose of 10 tha-1
30
pathogens
Trichoderma spp.
Gliocladium spp.
2004 2005 2006
20
10
0
x from all years
d. green waste manure at a dose of 5 tha-1
Fig. 4. Fungi isolated from the soil fertilized with compost (%)
113
In the fertilization variants analyzed, the latter species occurred sparsely and
only in the first year of the study (2004) and did not colonize the soil fertilized with
“Dano” compost in either variant of application or that fertilized with farm manure
at a dose of 10 tha-1.
GORODECKI and HADAR (1990) confirmed the inhibiting effect of fertilization
with farm manure on the growth of S. sclerotiorum, and that of R. solani. In turn,
FERRAZ et al. (1999) reported on suppressed sprouting of S. sclerotiorum sclerocia
in the soil from under a tomato crop fertilized with green wastes.
In the reported study, the perpetrator of grey rot was only isolated in 2005 from
the soil fertilized with “Dano” compost in both variants of applications, and its
contribution in the fungal community did not exceed 6%. ELAD AND SHTIENBERG
(1994) demonstrated that the composts applied were effective in protecting selected
plant species against infestation with B. cinerea. Single isolates of C. coccodes
(3.1% of all isolates in this variant) were obtained in this study from the soil fed
with mineral NPK fertilizer in the second year of cultivation, i.e. soil from under
potato, whereas these isolates were not obtained from the soil fertilized with
composts.
In the soil fertilized with minerals, the content of total N is subject to increase
and, as reported by ZARZYCKA (1990), better growth of this fungus proceeds under
conditions of insufficient supply of this macroelement in the soil, which corresponds
to the results of the presented study.
Saprotrophic fungi were most often represented by species of the genera
Gliocladium, Paecilomyces and Trichoderma characterized by the antagonistic
action against pathogens, as well as by fungi of the genus Penicillium and of the
order Mucorales. They colonized all communities of soil fungi analyzed in the
study. The greatest population of these fungi was noted in the soil environment with
farm manure applied at a dose of 10 tha-1, with an especially high contribution
(33%) of species belonging to the order Mucorales (Mortierella alpina,
M. isabelina, Mucor hiemalis, Rhizopus nigricans and Zygorhynchus spp.).
DOMSCH et al. (1980) included them as permanent components of a fungal
community of the soil environment determined as a result of fertilization. In the
current study, these were the plots with that fertilizer, i.e. farm manure introduced in
a single dose and in separate doses, that were characterized by favorable dynamics
of changes in the population of those fungi, i.e. a successive increase in their count
in the subsequent experimental years. Species of the genus Trichoderma (T.
aureoviride, T. hamatum, T. harzianum, T. koningii, T. viride and T. polysporum)
were isolated in consecutive vegetative seasons from the soil in all fertilization
variants, except from the soil fertilized twice with 5 tha-1 of “Dano” compost and
from the soil fertilized with green waste compost at both variants of application.
Species of the genera Gliocladium: G. catenulatum, G. penicillioides, G. roseum and
G. salmonicolor were isolated from the soil less frequently, and their contribution in
the fungal community did not exceed 6%, except for the third year of the study in
the variants with a single administration of green waste compost and farm manure
(16.7 and 9.5%, respectively). Fungi of the genus Paecilomyces most often
colonized the soil from under the crop of spring barley in the variant with green
waste compost (double administration of the fertilizer) and that of winter rape (3rd
114
year of the study) in the variant with farm manure applied in a split dose. Ample
research studies (HOITINK, BOEHM 1999) have indicated the stimulating effect of
various organic fertilizers on the growth of fungi antagonistic to pathogens. In soil
fed with organic fertilizers, as compared to that fertilized with minerals, BULLOCK et
al. (2002) observed a higher prevalence of fungi of the genus Trichoderma.
Effect of extracts from compost on the growth of soil fungi
The aqueous extracts prepared from composted municipal wastes, subjected to
laboratory analyses, were found to suppress the growth of mycelium of six species
of pathogens isolated from soil (Fig. 5.).
80
Dano
70
Dano compost
green waste
green waste
60
LSDp=0.01)= 11.46
50
50
LSDp=0.01)= 3.30
40
50
30
30
20
10
20
10
0
10
0
20
Bc
Cc
Fc
Fe
Fo
Fp
LSDp=0.01)= 4.21
60
40
40
30
70
Bc Cc
0
Fc Fe Fo Fp
a. after 4 days of culture
Dano compost
green waste
80
70
60
LSDp=0.01)= 8.40
50
40
30
20
10
0
Bc
Cc
Fc
Fe
Fo
Fp
50
48
46
44
42
40
38
Dano compost
70
green waste
60
LSDp=0.01)= 3.10
50
LSDp=0.01)= 4.06
40
30
20
10
0
Bc
Cc
Fc
Fe
Fo
Fp
b. after 8 days of culture
Explanations: Bc –Botrytis cinerea, Cc-Colletotrichum coccodes, Fc-Fusarium culmorum, Fe-F.
equiseti, Fo-F. oxysporum, Fp-F. poae
Fig. 5. Percentage of growth inhibition of pathogen mycelium on PDA medium with aqueous
extracts from compost
The “Dano” compost was characterized by a higher biological activity than the
compost from green waste in both analytical periods; i.e. after 4 and 8 days. In the
earlier period, the most susceptible to the addition of the extracts in agar medium
appeared to be B. cinerea, C. coccodes and F. poae species, whereas in the later
period it was the C. coccodes species. This confirms the results of the field
115
experiment, because the C. coccodes species did not colonize the soil in any of the
plots fertilized with compost, whereas B. cinerea occurred in small numbers only in
the soil fertilized with “Dano” compost in both variants of its application. In
addition, analyses demonstrated the lowest index of growth inhibition of F.
culmorum mycelium.
KITA ET AL. (1996), in their in vitro research on the effect of the addition of
aqueous extract to PDA, observed poor growth of colonies and aerial mycelium of
such species as: Rhizoctonia solani and F. culmorum. In turn, STOMPOR-CHRZAN
(2001) demonstrated the susceptibility of fungi of the genus Fusarium to aqueous
extracts from manure-based vermicomposts.
Summary
The results achieved in the study demonstrate that the applied natural
fertilization with manure and organic fertilization with composts from municipal
wastes modified the qualitative composition of the soil fungi community to a greater
extent than its quantitative structure. The positive impact of this type of fertilization
was manifested in suppressing the population of pathogenic fungi in respect to the
control variant (without fertilization) and the variant with mineral fertilization. The
prevailing species isolated in the study were those belonging to Aureobasidium and
Fusarium genera. A tendency towards a stronger reduction of pathogen populations
was observed in the soil with a single administration of 10 tha-1 of organic fertilizer
as compared to the variant with a double application (5 tha-1 each). The study also
demonstrated an increase in the population of fungi of the genera Gliocladium,
Paecilomyces and Trichoderma antagonistic to pathogens. The most positive
changes in the population of beneficial fungi were observed in the plots with manure
administered both in a single dose and in a split dose. The in vitro test additionally
showed that the aqueous extracts from composts added to the culture medium
inhibited the growth of mycelium of 6 species of pathogens, though their response
was diversified. The least susceptible to the addition of extract into the medium (the
lowest index of mycelium growth inhibition) appeared to be Fusarium culmorum
species. Finally, a higher fungistatic activity was demonstrated for the extract
prepared from the “Dano” compost than for that from green waste.
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1
BoĪena Cwalina-Ambroziak
Chair of Phytopathology and Entomology
ul. Prawocheskiego 17, 10-720 Olsztyn, POLAND
2
Jadwiga Wierzbowska
Chair of Agricultural Chemistry and Environment Protection
ul. Oczapowskiego 8, 10-719 Olsztyn, POLAND
118
CHAPTER VIII
Szejniuk Boena1, Wasilewski Piotr2, Budziska Katarzyna1,
Gałzewska Beata1, Kubisz Łukasz1
EFFECT OF COMPOST FROM SEWAGE SLUDGE ON
PLANT DEVELOPMENT
Introduction
Natural use of sewage sludge is in accordance with the policy of the European
Union, which approves of introduction into soils components accumulated in biowastes, on condition of meeting the requirements contained in Directives concerning
protection of the environment and soil against contamination (GWOREK et al. 2002).
Issues related to the processing and management of sewage sludge are essential
in the present time, given an increase in the number of sewage treatment plants
established in Poland and the necessity of meeting the requirements connected to the
standards concerning environmental protection. Moreover, a constant growth in the
number of new sewage treatment plants contributes to formation of considerable
amounts of sludge which poses a serious problem. Its composition depends on the
type and origin of sewage and the technology of treatment (WŁODEK 2007). Sludge
formed in biological treatment plants usually has the content of organic substance
and nutrients which is favourable for plants (MAZUR 1996), thus it can be applied
for natural management.
The direction of sewage sludge processing depends on physico-chemical and
sanitary and hygienic properties. The excessive content of heavy metals is the factor
that decidedly limits sewage sludge application in the natural environment (HOODA,
ALLOWAY 1996, PALES et al. 1996). In the order of the Ministry of the Environment
of 1 August 2002 on municipal sewage sludge, the content of heavy metals and their
load introduced into the environment is assumed as one of the basic criteria for its
agricultural use. Exceeding the standard level of even one of all the list of heavy
metals disqualifies such material from the natural use. In the case of sewage sludge
generated in treatment plants handling the areas producing no industrial wastes most
often there is a possibility of their processing, after initial processes of thickening
and stabilization, into composts which find application to soil fertilization.
Application of compost for protection of soil structure and an increase in nutrient
availability exerts a favourable effect on the state of the environment (JAKOBSEN
1995].
119
Compost from sewage sludge applied as a fertilizer has favourable soil-forming
properties; organic substances from compost remain in soil for a longer time, which
determines the improvement of the water and gas relations of soil and leads to an
increase in fertility indexes (CORTELLINI et al. 1996; SZEJNIUK 2005). Due to its
manurial properties, the compost obtained from municipal sewage sludges which
meets quality standards shows the effect similar to that of organic fertilizers which
are applied traditionally, and is an effective source of N, P and K utilized by plants
(WARMAN, TERMEER 2005). The correct effect of a compost on the soil
environment is determined by its proper chemical composition (JAKOBSEN 1995).
According to GONDEK and FILIPEK-MAZUR (2006), analyzing the effect of compost
application on soil properties and the availability of some microelements under the
influence of those additives, they indicated a series of far-reaching positive changes
in soil, preparing this organic fertilizer for retaining or restoring fertility of
agricultural soils. Using compost as an organic fertilizer constitutes the optimal
method which allows the complete utilization of the physical and chemical
properties of this material for regeneration, fertilization and recovering of the most
essential bio-components of soil.
Favourable effect of compost is observed among others in its deacidifying
activity, a decrease in hydrolytic acidity, a growth of the contents of calcium, carbon
and organic nitrogen and a considerable increase in proportion of bio-available
forms of microelements (WARMAN, TERMEER 2005). Composts made with an
addition of rural wastes and straw constitute a valuable organic fertilizer rich in
nutrients (particularly in nitrogen and phosphorus). The effect of such a fertilizer on
plants is slower, since nitrogen compounds occur in it in humus combinations. In
this way, it can exert a favourable effect for several years, as opposed to mineral
fertilization, particularly with nitrogen and potassium (CZYYK et al. 2002).
Forming the fertility and yielding potential of soils is a long-lived process, and
changes in physico-chemical properties, both favourable and unfavourable, are
clearly noticeable only in long-term experiments. In consequence, the full spectrum
of modifying effect of composts on soil physico-chemical properties is possible to
observe and assess as a whole in studies conducted over a period of several years
(GONDEK, FILIPEK-MAZUR 2005; GONDEK, FILIPEK-MAZUR 2006).
Introducing moderate amounts of compost into soils resulted in improving its
composition, especially when the compost was applied to the surface of the soil and
after sowing the cultivated crops. The soil surface was in this way protected against
the negative impact of rainfall and fast drying at a later time. Under these conditions,
also water soaked through soil much faster, also after the application of a thin layer
of compost (JAKOBSEN 1995). In agricultural practice, spreading of manurial activity
in time can be treated as a undeniable asset of this fertilizing material, which due to
its non-invasive effect on soil and by means of an increase in sorption capacity
improves the structure and increases the water capacity of soils, and exerts a slight
influence on the chemical composition of generated effluents which, in turn,
directly translates into water environment safety (CZYYK, KOZDRA 2003).
In Poland, the trade standard BN-89/9103-09 for composts from mixed wastes,
including composts produced from sewage sludges or with an addition of sewage
sludge, was in effect for many years, which contained requirements concerning
120
macroelements, heavy metal concentration, proportion of glass, ceramics and stones,
as well as sanitary and hygienic features. After coming into force of the act on
fertilizers and fertilization, the entities launching fertilizers produced on the basis
of organic substances need an appropriate permission given by the Minister
of Agriculture. The order of 19 October 2004 concerning the execution of provisions
of the act on fertilizers and fertilization defines the scope of research and
requirements concerning the opinions which make it possible to give a permit for
launching such a fertilizer.
A favourable impact of various composts which provide the source of available
nutrients on an increase in plant yield is emphasized in the literature (EPSTEIN 1997,
AGGELIDES, BERNAL et al. 1998, LONDRA 2000, MARINARI et al. 2000). It has been
indicated that some plants cultivated on soils enriched with compost show varied
growth dynamics (KORBOULEWSKY et al. 2002). Positive effect of compost from
municipal sewage sludge meeting requirements concerning quality depends on
keeping the appropriate proportions during soil fertilization. Composts made from
sewage sludges exert an influence on the content of potassium, calcium and
magnesium in agricultural crops, which is determined by the type and rate of
compost (CIEKO, HARNISZ 2002).
Factors of the study
In order to indicate the effect of compost on the growth of selected agricultural
crops, an experiment was carried out during two growing seasons in 2005 and 2006.
The one-factorial pot experiment was established in the complete random design in
which emergence, growth and green matter yield of plants cultivated on different
substrates were evaluated. Pots of a volume of 11 litres and an area of 0.0615 m2
were filled with soil of class IVb collected from the topsoil, into which an addition
of compost from sewage sludge was introduced according to the following scheme:
P0 – soil without an addition of compost – the control
P1 – soil + compost in a ratio of 3 : 1
Compost from sewage sludge applied in the experiment was made from
activated sludge with a dry matter content of 24%, subjected to the process
of dehydration by means of the ANDRITZ press. Sewage sludge was mixed with
wood chips and burnt lime in a ratio of 1: 0.3 : 0.01. Composted material was placed
in a revolving bioreactor for 5 days and then subjected to maturation in heaps until
the moment of obtaining compost stability. The compost obtained was characterized
by a high content of phosphorus (0.56% d.m.) and potassium (0.30% d.m.), essential
for its manurial value. In addition, this fertilizer had a favourable alkaline reaction
(pH 8.2), exhibiting deacidifying effect on soil. Relatively low content of organic
substance (25.86% d.m.) and organic nitrogen (0.47% d.m.) was found, strongly
correlated with the manurial value of the tested material. Physico-chemical analysis
confirmed that the examined parameters of the compost was in accordance with the
quality standards contained in the Regulation of the Ministry of Agriculture and
Rural Development of 19 October 2004 on executing some regulations of the act of
121
fertilizers and fertilization [Dz. U. No. 236 item 2369]. The level of the heavy
metals determined (Cr, Zn, Cd, Cu, Pb) was appropriate from the point of view of
environmental protection and corresponded to the standards required for composts
applied as fertilizers.
Prepared substrates with different proportions of compost were seeded as
follows: yellow and blue lupines (13 germinating seeds per pot), oats
(62 germinating seeds per pot) and spring rye (55 germinating seeds per pot). After
sowing, until the time of full emergence, the pots were covered with a net in order
to protect them against birds. After 14 days from sowing, the assessment of
emergence was carried out by means of counting the number of plants grown in each
pot. Observations of the appearance of plants in the pots were conducted at twoweek intervals, and deformations, discolorations, the occurrence of diseases and
pests were recorded. In dry periods during the growth the plants in pots were
watered in order to obtain a substrate moisture at a level of 60% field water capacity.
Water volume was changeable depending on its transpiration through the plants
during their growth. Directly before harvesting, the average plant height of a given
species in the sample was determined (oats and spring rye – average height
of 10 plants of a given species), as well as the amount of plants of a given species in
the sample. After reaching harvesting maturity by the plants, they were harvested,
and biometric measurements and statistical calculations were carried out. Results
of study were subjected to statistical analysis in the completely random design.
Significance of differences between the averages was determined by means
of Tukey’s test.
Responses of chosen agricultural crops to addition
of compost to soil substrates
Growing deficit of humus substances in soils that occurs in Poland resulted
in a distinct growth of initiatives aiming at looking for new sources of organic
matter, which could be applied safely as alternative fertilizers. According to MAZUR
(1999), this idea is supported by the confidence that ecological balance can be
retained or restored, among others, by means of proper conditions of generated
waste utilization, which is becoming more and more apparent to the society. Correct
application of compost in order to optimize yielding conditions and to provide the
nutritional and technological quality of yield requires thorough knowledge of
responses of selected agricultural crops cultivated on soils supplied with this
fertilizer. The experiment aiming at indicating the effect of compost addition to soil
was carried out in two growing seasons and it showed a distinct influence of the
addition of compost from sewage sludge to soil on the emergence of legumes and
cereals. According to the data presented in Table 1, emergences of the tested plant
species were in most cases significantly diversified by the addition of compost to the
substrate.
122
Table 1
Effect of varied addition of compost on emergence (plants x pot-1) of plants
2005
Substrate
P0
Yellow
lupine
9.0
Blue
lupine
12.7
P1
3.7
P2
2006
45.7
Spring
rye
50.0
Yellow
lupine
11.0
Blue
lupine
9.0
8.0
32.3
37.0
0.3
8.3
8.0
34.7
36.3
P3
9.3
10.7
51.3
Average
7.6
9.8
LSDp=0.05
1.5
1.7
Oats
Oats
Spring rye
60.3
46.3
3.3
56.7
40.3
3.7
6.7
55.0
39.0
49.7
10.3
8.0
56.0
49.3
41.0
43.2
8.3
6.8
57.0
43.8
8.7
7.9
4.3
n.s.
n.s.
9.0
n.s. – differences not significant
Yellow and blue lupines responded negatively to an addition of compost to soil,
since in all the cases in 2005 and 2006 a decrease in the number of legumes
emerging was observed at higher compost rates. A difference between water
potential in the soil solution and that in seeds has a decisive impact on water
absorption during swelling of seeds. An increase in compost proportion in the
substrate probably resulted in a growth of soil solution concentration, which along
with high demand of lupine seeds for water in the course of swelling (the amount of
water absorbed in lupines amounts to 170% of seed mass, and in cereals 60 – 80%)
caused the worsening of their emergence (GRZESIUK, KULKA 1981, JASISKA,
KOTECKI 1999). Similar results of a study was presented by LEKAN and KACPEREK
(1990), who found, on the basis of the long-term experiments carried out with the
use of compost from municipal wastes, that an addition of compost inhibited plant
germination and emergence.
FILIPEK-MAZUR AND GONDEK (2003) report that an unfavourable effect
of compost observed in the first year after its application decreases in successive
years of the study. Slightly different opinion is presented by CZYYK et al. (2002),
who report that varied rates of compost applied by them did not have a significant
effect on plant emergence, which was uniform in all the combinations of the
experiment. The results obtained from the experiment are confirmed by an earlier
study by SZEJNIUK et al. (2005), where using similar methods, the weakest plant
emergence was also observed at a higher rate of compost. In the results of the
present study (Table 1) in cereal plants, the loss of oats and spring rye at the initial
stage of growth remained at a low level as compared with legumes. WINIARSKA and
LEKAN [1991) report that the diversification of plant responses to an addition
of compost to soil might be related to individual abilities of particular plants
to utilize nutrients taken up from organic fertilizers.
The height of plants depended on the substrate on which they were growing.
From the data (Table 2) it follows that yellow lupine in 2005 clearly negatively
responded to an addition of compost to soil, decreasing the height of shoots by as
123
much as 1/3 at the lowest rate applied. Further addition of organic material did not
result in the intensity of shoot reduction in this species. Also blue lupine indicated
a decrease in shoot height after fertilization with compost, yet such a response was
proved between the control and treatments with P2 and P1. Significant statistic
difference in the height of plants fertilized with compost was found in 2006 in oats
and blue lupine cultivation (Table 2). In the case of oats, the greatest plant height
was observed on treatments P2 with an addition of compost in a ratio of 6:1, where
the height was larger on average by 29%, as compared with the plants on the control
substrate P0. Significant differences in the height of oats were found between the
plants coming from pots filled with soil only (P0) and the plants from the treatments
P2 and P3.
Table 2
Effect of varied addition of compost on plant height (cm)
2005
Substrate
2006
54.8
Spring
rye
73.7
Yellow
lupine
46.4
Blue
lupine
47.1
16.5
57.3
64.3
n.d.
18.6
18.1
65.2
59.8
P3
18.3
20.4
50.3
Average
20.7
19.5
LSDp=0.05
8.4
4.5
P0
Yellow
lupine
27.2
Blue
lupine
23.1
P1
18.9
P2
n.d. – no data;
Oats
Spring rye
41.6
58.7
30.7
50.7
62.7
51.0
29.6
53.8
64.0
60.1
41.4
40.6
51.1
63.4
56.9
64.5
46.3
37.6
49.3
62.4
n.s.
6.8
n.s.
17.4
9.4
n.s.
Oats
n.s. – differences not significant
Results similar to those described above were obtained by PARADYSZ (2001),
who in his study indicated a favourable effect of compost on the growth and size of
plants. This author, however, emphasizes that stimulating effect of the addition
of compost was caused by improvement of soil properties, such as aeration and
retaining of moisture.
By contrast with the conclusions formulated by this author, however, in the own
pot study an addition of compost to soil in the cultivation of blue lupine caused
a decrease in plant height in all the experimental variants. The smallest height of
blue lupine in 2006 was observed on treatments P2 and it was significantly smaller
as compared with the value of the tested character on the control treatments (by
37%).
Different response of cereal crops and legumes to an addition of compost to soil
might result from individual needs of those plants in respect of the contents
of particular nutrients in the soil substrate and different preferences concerning
conditions of the soil environment, especially in terms of the presence of the bacteria
Rhizobium living in symbiosis with lupines (SZEMBER 2001).
124
Effect of compost on tested plant yield
Poor emergence of plants had an unfavourable effect on the green forage yield
of the tested plant species. In both growing seasons, lupines yielded significantly the
best on the substrate without an addition of compost (Table 3). A decrease in green
forage yield depending on the year of the study and the substrate was 37.1% - 63.9%
for yellow lupine and 32.3% - 94.3% for blue lupine. Yielding of cereal plants on
substrates with an addition of compost was the absolute opposite to the response
of lupines, as green matter yields of those plants were statistically significantly
higher and grew proportionally along with the growth of the amount of compost
added to the substrate.
Table 3
-1
Effect of varied addition of compost on yield (g x pot ) of green mass of tested plants
2005
Substrate
2006
58.9
Spring
rye
26.4
Yellow
lupine
72.5
Blue
lupine
51.4
20.7
80.1
47.5
0.0
28.3
24.5
89.1
48.5
P3
28.7
29.6
129.6
Average
29.3
26.4
LSDp=0.05
5.6
4.9
P0
Yellow
lupine
36.9
Blue
lupine
30.6
P1
23.2
P2
Oats
Spring rye
57.3
37.3
2.9
92.0
67.2
26.2
16.5
98.7
70.0
67.2
57.0
32.4
73.9
54.1
89.4
47.4
51.9
25.8
80.5
57.1
31.1
20.7
40.4
12.9
25.1
9.5
Oats
Similar tendencies were observed in a study of the effect of an addition of
compost from municipal wastes on the yield of rye green matter in pot experiments.
It was found that winter rye yield increased proportionally to the growth of an
addition of compost into podzolic soil and loose sand only up to a level of 2%,
whereas above this limit a decrease in plant yield occurred (SZEJNIUK 1997]. Pot
experiments carried out by CZYYK et al. (2002) confirm the favourable effect of
compost from sewage sludge and straw on maize yield, since it was observed that
the application of the highest rates of compost resulted in proportional growth in
yield of this plant. Mustard, in turn, which was the next plant tested, responded
positively to fertilization with mineral nitrogen, whereas an addition of compost did
not affect diversification of yields according to the rate size. According to
KORBOULEWSKY et al. (2002), plants grown on soil substrates enriched with
compost showed various growth dynamics, yet a general, evident, favourable
response occurred, resulting in an increase in the biomass of the tested plants grown
on soils with an addition of compost.
It follows from Table 3 that the yield of rye green matter also showed an upward
tendency in 2005 and 2006 in response to an addition of compost into the soil
substrate. FILIPEK-MAZUR and GONDEK (2003) report that the comparable yield of
125
oat dry matter was obtained after fertilization with farmyard manure and compost.
Moreover, it was proved that the manurial activity of composts is even better than
that of farmyard manure (GONDEK, FILIPEK-MAZUR 2005). Similarly, KOCH et al.
(1997), in a study of the effect of sewage sludge and composts obtained from it on
the cultivation of selected crops under field conditions, proved that considerable
differences occurred in the yield height of the tested plants, which argues in favour
of crops coming from plots fertilized with sludges and compost. BARAN et al.
(1993a), in turn, found higher plant yields on a soil fertilized with sewage sludge. In
that case, they recorded an increase in the content of total carbon and a fraction of
humins, contributing to a better availability of accessible nutrients. Favourable effect
of sewage sludge was found already at a low level of fertilization from 1 to 5% in
relation to the control groups (BARAN 1993b).
Summing up the obtained results presented in this study, it might be concluded
that the application of compost from sewage sludge has the most favourable effect
on the development of spring rye and oats.
Summary
Composting is one of methods for the natural utilization of sewage sludge.
Increasing number of sewage treated and biodegradable waste deposition at landfill
sites results in a growing interest in this way of management of wastes from sewage
treatment plants. Sewage sludge can be a valuable fertilizer which is possible to be
applied in agriculture, on condition that the processes of its initial processing and
sanitization will be carried out properly. Compost produced from sewage sludge
should be constantly monitored in order to eliminate the hazard of pathogenic
microorganisms.
The present study confirms the possibility of applying composts from sewage
sludge as fertilizers on soils of low fertility (classes IVb-VI). It also indicates
different responses of the plants – particularly legumes – to an addition of compost
to the soil or substrate. Therefore, further studies are necessary concerning the plant
response to an addition of compost from sewage sludge to soils or substrates.
Moreover, action should be taken aiming at showing farmers that these composts
can be as safe and effective in increasing the height and quality of yield as the
traditional organic fertilizers – farmyard manure, straw and other green manures.
The study indicated that the compost from sewage sludge used in the experiments
contributed to an increase in green matter yield of oats and spring rye, which proves
its considerable usefulness for soil fertilization in cereal crop cultivation.
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na plony Īyta w doĞwiadczeniu wazonowym. Ekol. i Techn. 6, 30: 25–27.
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Bydgoszcz.
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Ğciekowych i odpadów interwencyjnych na wzrost i plonowanie wybranych roĞlin
uprawnych. Zesz. Probl. Post. Nauk Roln., 506: 471 – 478.
SZEMBER A. 2001: Zarys mikrobiologii rolniczej. Wyd. AR Lublin.
USTAWA z dnia 26 lipca o nawozach i nawoeniu. Dz. U. Nr 89, poz. 991, 2000.
WARMAN P.R., TERMEER W.C. 2005: Evaluation of sewage sludge, septic waste and sludge
compost applications to corn and forage: yields and N, P and K content to crops and
soils. Bioresource Technology 96: 955-961.
WINIARSKA Z., LEKAN S. 1991: Wpływ kompostu z odpadów miejskich na plonowanie roĞlin
i właĞciwoĞci gleby w doĞwiadczeniu polowym. Wyd. IUNG Puławy: 49-70.
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1
Szejniuk BoĪena, 1BudziĔska Katarzyna, 1GałĊzewska Beata, 1Kubisz Łukasz
Department of Animal Hygiene and Microbiology of the Environment
University of Techonology and Life Sciences
ul. Mazowiecka 28, 85-084 Bydgoszcz, POLAND
2
Wasilewski Piotr
Department of Plant Production and Experimenting
University of Techonology and Life Sciences
ul. Kordeckiego 20, 85-225 Bydgoszcz, POLAND
128
CHAPTER IX
Janusz Augustynowicz1, Stefan Pietkiewicz2, Mohamed Hazem Kalaji2,
Stefan Russel3
THE EFFECT OF SLUDGE FERTILIZATION ON
CHOOSEN PARAMETERS OF CHLOROPHYLL
FLUORESCENCE AND BIOMASS YIELD
OF JERUSALEM ARTICHOKE
(HELIANTHUS TUBEROSUS L.)
Introduction
According to 2001 EU Directive related to the Promotion of Electricity
produced from Renewable Energy Sources in the internal electricity market, each
member states of European Union should reach till 2010 r. a 12 % contribution of
renewable sources energy gross use, while the whole Community 22,1 %. Polish
Development Strategy of Renewable Energy Sector adopted by the Parliament of the
Republic of Poland (2001) promotes development of renewable energy sources in
our country and indicates basic goals as well as conditions of renewable energy
development in Poland till 2020. There is an assumption to increase of the share
of energy from renewable sources in the whole country fuel-energy balance up to
7,5% (340 PJ) in 2010 and to 14% in 2020. It means three times increase as
compared to 1999 (2,5% – 105 PJ). Converting biomass is one of a processes to get
renewable energy by the use of energy crops such as e.g. Jerusalem artichoke,
Virginia mallow, Giant Knotweed, Common osier, Reed Canarygrass, Miscanthus
Giganteus and others (AUGUSTYNOWICZ et al. 2008).
During the last few years an increase interest of energy crops Jerusalem
artichoke (Helianthus tuberosus L.), also known as topinambura is observed. This
plant originates from Northern America and belongs to the family Asteraceae
(MAJTKOWSKI 2003). Stems of topinambur of up to 3 cm diameter are 2 – 4 m tall.
Analyzed species forms underground stolons, bearing at their tips tubers (same as
with potato). The raw material for energy purposes are both tubers, which can be
used for bioethanol or biogas production as well as above ground parts: fresh or
fermented – for biogas production, dry – for direct combustion of fragmented mass
or to produce briquettes and pellets (MAJTKOWSKI 2003, STOLARSKI 2004).
Topinambur is a crop of very high production potential (KAYS, NOTTINGHAM
2007). Its yielding is determined first of all by genotype, but the soil culture and
content has also a significant effect. Same as with root and tuber crops the best for
129
topinambur cultivation are soils midloosed, aerated, rich in mineral nutrients and
having enough moisture. It can be also cultivated in the worse site, less profitable for
potatoes. There is no way to cultivate this crop in marsh and acid soils. All crops can
serve as a forecrop for it, even some small weeded fallows, nevertheless it requires
deep ploughing (20-30 cm). Topinambur can be cultivated in the same stand for 3-4
years (STOLARSKI, 2004). In one of the national experiments total biomass yield was
ca 110 tha-1: above ground mass 75,6 t.ha-1, tubers 32,4 t.ha-1 (STOLARSKI, 2004).
The yielding in this experiment was much higher on polish soil bonitation class III
as compared to class IVb. Under polish conditions average field of topinambur on
dry matter basis is 10-16 t d.m..ha-1 (MAJTKOWSKI 2003, STOLARSKI, 2004). High
yielding potential, easy cultivation, low cost of launching plantation and big abilities
of adaptation to soil conditions, speak for further dissemination of this crop, this
time however, as an energy crop, in Poland (MAJTKOWSKI 2003). A real possibility
to obtain high yielding, being the costs low, is the principal cause of increased
interest in this crop GRADZIUK 2003). The goal of research on energy crops aim to
elaborate such a way of its cultivation to reach maximum biomass increase. There
are two ways of cultivation: traditional, providing nitrogen from such its
conventional sources as mineral fertilizers or new, using as its source such
inconvenient for environment waste, as sludge (AUGUSTYNOWICZ et. al. 2008).
Restoration to soil mineral nutrients from sludges seems to be adequate not only
from economic point of view, but also is necessary to maintain and renew ecological
homeostasis. Mineral and organic composition of residuals from municipal
wastewater treatment plant is proximal to a soil organic substance – a humus
(BCZALSKA 1998). So, there is possible natural, with this agricultural use, of the
Sludges. The sue of the latter in non industrial way should meet needs concerning
their chemical composition and sanitary state. Heavy metal contents in sludges for
non industrial use is limited due to their toxic influence on living organisms and
ability for bioacumulation (BIE 2002).
Plant physiology offers many physiological parameters to be exploited in
different plant science fields. Some physiological issues such as the photosynthetic
efficiency of plants have started to be investigated to evaluate the performance of
crops when growing under various environmental and growth conditions. One of
these parameters is chlorophyll fluorescence which indicates the capacity of plants
to convert light energy to biochemical energy during the photosynthetic process. The
advantages of this technique are that, it is non-invasive, non-destructive and rapidly
measured using highly portable equipment. Many scientists have already applied
this technique in their researches as a biomarker or bioindicator and proved that it
provides reliable information about plant photosynthetic efficiency which is highly
correlated to plant vitality (KALAJI I and GUO 2008).
The aim of the paper was to analyze the effect of fertilization with sludges from
municipal wastewater treatment plant on activity of photosynthetic apparatus of
Jerusalem artichoke (Helianthus tuberosus L) as crop claimed to be the most
promising energy crops for Poland.
130
Research conditions
In 2007, in the Institute of Land Reclamation and Grassland Farming Falenty
near Warszawa a two factorial field experiment in randomized blocks design with
three replicates was conducted. Each replicates involved 9 plants. The following
treatments were used:
1. control (no nitrogen fertilization) („0”),
2. 100% N sludge, 0% N mineral fertilizer (100% sludge),
6. 0% N sludge, 100% N mineral fertilizer (0% sludge).
An equivalent of 170 kgha-1 pure nitrogen was used as fertilization in
treatments with sludges. The used nitrogen rate was established according to the
acceptable maximum resulting from the Act on Fertilizers and Fertilization (2000,
2004). The sludge was provided by communal wastewater treatment plant Falenty.
The sludge meets all requirements concerning possibility its use in agriculture. The
chemical indices that characterize used sludge presented in table 1.
Table 1
Characteristics of sludge chemical composition
Analyzed index
pH in water
Unit
pH
Result
12.7
Dry matter content
%
22.73
N total
% dry matter
2.93
P2O5
% dry matter
3.79
K2O
% dry matter
0.41
Cd
mg kg dry matter
0.94
Cr
mg.kg-1 dry matter
16.7
Cu
mg kg dry matter
Ni
mg kg dry matter
Pb
mg kg dry matter
Zn
Hg
.
-1
-1
104
-1
15.6
-1
19.8
-1
mg kg dry matter
603
mg.kg-1 dry matter
0.57
.
.
.
.
Each additional treatment, 0 treatment involving, was enriched with mineral
potassium fertilizer 240 kg K.ha-1. Its rate was established on the basis of the crops
nutritional needs and potassium content in sludge.
131
The experiment was performed on the soil classified as black earth degraded.
Parameters that characterized this soil presented in table 2.
Table 2
Characteristics of soil for Jerusalem artichoke cultivation in Falenty
pH
Total in %
C:N
Level
Depth
[cm]
H2 O
KCl
C
N
A
0-30
5.16
4.71
1.38
0.09
15.33
C
30-60
5.05
4.8
2.38
0.15
15.86
Ak
60-80
5.83
5.18
1.01
0.04
25.25
C
80-150
5.84
5.16
0.47
0.02
23.50
Measurements of physiological indices that characterize photosynthetic
apparatus of topinambur performed in 3 terms July 2, August 2 and October 17. The
following indices of physiological activity of the apparatus were used: index of
photosystem II (PSII) functioning and vitality (Performance Index, P.I.) and
maximum quantum yield of PSII (FV/FM). These indices are recommended by Kalaji
and Łoboda (2007, 2009) as the best indicators of photosynthetic apparatus
efficiency measured with detection and analysis of chlorophyll a fluorescence signal
technique. There were determined with the use of HandyPEA fluorimeter
(Hansatech Instruments, King’s Lynn, Norfolk, UK).
The above mentioned parameters of chlorophyll a fluorescence were measured
for three layers of the canopy (upper, middle and lower) in 3 replicates for each
treatment. At harvest, on 25 XI biomass field of Jerusalem artichoke was
determined. The results were statistically analyzed with the use of SAS 9.1. version.
Indices of Jerusalem artichoke photosystem II functioning
Figure 1 presents data which characterize changes of global Performance Index
PSII (P.I) of topinambur fertilized with sludge during vegetation. The highest P.I.
values were reached in middle period of vegetation of the analyzed crop – the index
is about 5 units. In July the sludge used caused a decrease of analyzed index as
compared with that of 100% share of mineral fertilizer. In August the value of P.I.
increased for treatments with sludge, the highest values being for treatments with z
50% and 75% share of sludge nitrogen. At the end of vegetation the sludge clearly
stimulated an increase in activity of photosynthetic apparatus of studied crop.
Similar data reported AUGUSTYNOWICZ et al. (2008).
132
6
P.I. [r.u.]
0
4
100% sludge
75% sludge
2
0
50% sludge
25% sludge
02.07
02.08
0% sludge
17.10
Date of measurement
Fig. 1. Changes of global Performance Index (P.I.) of Jerusalem artichoke fertilized with
a sludge during vegetation
In the case of upper layer of canopy the values of Performance Index (P.I) on
2.07 showed the use sludge increased activity of photosynthetic apparatus for
treatment with 75% share of sludge nitrogen in the applied nitrogen rate (Fig. 2A).
The data of measurements made one month later (2.08) indicated a similar tendency
of sludge affecting on the index of photosynthetic apparatus efficiency. The highest
values P.I. were found for treatment with 50% share of sludge nitrogen.
Measurements in autumn (17.10) showed significant decrease in activity
of photosynthetic apparatus for treatment 100 % share of mineral nitrogen as
compared to those where the sludge without mineral nitrogen was used.
Data obtained on 2.07 for central layer of the canopy layer showed the use
of sludge caused a decrease in Performance Index of PSII when compared to the
treatment containing the sole mineral nitrogen (Fig. 2B). On 02.08 activity
of photosynthetic apparatus of Jerusalem artichoke was the lowest for the control
treatment (a slight above 3 relative units). In this term the highest value of P.I. was
found for the crop fertilized with 75 % share of sludge nitrogen. It was almost
3 relative units higher than in the control. In October (17.10) a similar tendency as
during August were observed, the only exception was less intensive activity of
photosynthetic apparatus, while the highest value of analyzed index was found for
treatment „100% sludge”.
An analysis of Performance Index PSII (P.I.) for lower layer of canopy leaves
made 2.07 showed the use of sludge caused no visible effect (Fig. 2C). The highest
value of Performance Index of Photosystem II was found for treatment without
sludge nitrogen in applied rate of nitrogen. Data of measurements taken a month
later (2.08) indicated significant effect of sludge used on the activity of
photosynthetic apparatus of the crops. The highest value of analyzed P.I. was fund
for the treatment 25% share of sludge nitrogen in the applied nitrogen rate. At the
end of vegetation the activity of photosynthetic apparatus of Jerusalem artichoke
upper layer of canopy leaves was low and the crop started to wilt, senesce and die
The works of other authors imply the activity of photosynthetic apparatus of various
plants, crops seems to be lower in autumn than in springtime or during the summer
(KALAJI et al. 2004 a,b, KALAJI 2004).
133
P.I. [r.u.]
6
0
4
100% sludge
75% sludge
2
0
50% sludge
25% sludge
02.07
02.08
17.10
0% sludge
Date of measurement
Fig. 2A. Changes of PSII Performance Index (P.I.) for Jerusalem artichoke upper layer
of canopy leaves fertilized with a sludge during vegetation
P.I. [r.u.]
8
0
100% sludge
6
4
75% sludge
50% sludge
25% sludge
2
0
02.07
02.08
17.10
0% sludge
Date of measurement
Fig. 2B. Changes of PSII Performance Index (P.I.) for Jerusalem artichoke central layer
P.I. [r.u.]
8
0
100% sludge
6
4
75% sludge
50% sludge
25% sludge
2
0
02.07
02.08
17.10
0% sludge
Date of measurement
Fig. 2C. Changes of PSII Performance Index (P.I.) for Jerusalem artichoke upper layer
134
Changes in maximum quantum yield of the Jerusalem artichoke
Photosystem II
Numbered studied confirm the parameter FV/FM showed potential efficiency of
PSII and can be used as a reliable indicator of photochemical activity of
photosynthetic apparatus (KALAJI and ŁOBODA 2009). For a majority of plants in
full development and under non stressed conditions its maximum value is 0,83
(ANGELINI et al. 2001). A decrease in the parameter shows an analyzed crop was
earlier imposed to the activity of stress factors, which deteriorated functions
of PS II, thus decreasing the efficiency of the electron transport (HE et al. 1996).
Figure 3 presents the changes of maximum quantum field of Photosystem II
(FV/FM), for each individual date of measurement (global values), in Jerusalem
artichoke fertilized with a sludge, for the whole vegetation period. The highest value
of the analyzed index were noted in August (2.08), while the lowest ones by the end
of vegetation. Noteworthy is that during the first months of vegetation no
stimulating effect of a sludge on the analyzed index as compared to the applied sole
mineral fertilizer was found. In October (17.10) the sludge, especially in treatment
75% sludge, affected considerably an increase in quantum field of Photosystem II.
fv/fm
0,85
0
100% sludge
75% sludge
50% sludge
25% sludge
0,8
0,75
0,7
02.07
02.08
17.10
0% sludge
Date of measurement
Fig. 3. Changes of PSII maximum quantum yield (FV/FM) for Jerusalem artichoke upper
layer of canopy leaves fertilized with a sludge during vegetation
An analysis of changes in quantum yield of Photosystem II for upper layer of
Jerusalem artichoke canopy leaves performed in July (2.07) showed the sludge only
for treatments 100% sludge and 75% sludge stimulated increase in the analyzed
index (Fig. 4A). A month later the highest values of FV/FM were obtained for the
treatment with 50% sludge nitrogen in the applied rate of nitrogen, while
a comparable for a treatment containing only mineral nitrogen. At the end
of vegetation the higher values of maximum quantum yield than in the control were
found only for treatment 75 % sludge. The lowest value were found for treatment
with 100% share of mineral nitrogen in the applied rate of fertilizer.
Data obtained for July (2.07) in the respect of quantum yield of Photosystem
II values for central layer of Jerusalem artichoke canopy leaves showed no
significant differences for analyzed treatments (Fig. 4B). In August (2.08)
significantly higher value of PSII quantum field were found for treatment involving
135
a sludge only. At the end of vegetation (17.10) the higher values of the analyzed
index were observed for treatment with a sludge than for that with mineral nitrogen.
Leaves of the Jerusalem artichoke lower layer canopy leaves showed no
significant differences in quantum field of Photosystem II values, both in the view
of date of measurement and treatment (Fig. 4C).
fv/fm
0,85
0,8
0,75
0,7
0,65
02.07
02.08
17.10
0
100% sludge
75% sludge
50% sludge
25% sludge
0% sludge
Date of measurement
Fig. 4A. Changes of PSII maximum quantum yield (FV/FM) for Jerusalem artichoke upper
fv/fm
0,9
0,85
0,8
0,75
0,7
02.07
02.08
17.10
0
100% sludge
75% sludge
50% sludge
25% sludge
0% sludge
Date of measurement
fv/fm
Fig. 4B. Changes of PSII maximum quantum yield (FV/FM) for Jerusalem artichoke central
1
0,8
0,6
0,4
0,2
0
02.07
02.08
17.10
0
100% sludge
75% sludge
50% sludge
25% sludge
0% sludge
Date of measurement
Fig. 4C. Changes of PSII maximum quantum yield (FV/FM) for Jerusalem artichoke lower
136
Biomass yields of Jerusalem artichoke fertilized with a sludge is presented in
table 3. It results from this data the highest biomass was found for the crop of the
treatment, containing 100% share of sludge nitrogen in the applied nitrogen rate.
Jerusalem artichoke is a crop of high production potential. Cultivated in fertile
soils and with the abundance of water topinambur can field to 200 tha-1 fresh weight
(green above ground parts and tubers, together), with this tuber yield even 90 tha-1
(KOWALCZYK-JUKO 2003). On the other hand Kruczek (1995) reported tuber yield
15-30 tha-1 and green above ground mass: 50-70 tha-1. Comparison of these data
with the ours resulted in finding the obtained biomass yield seems to be relatively
low. However, the presented now data are from the first growing season after
launching the plantation.
Table 3
Biomass field of Jerusalem artichoke fertilized with a sludge
Treatment
Biomass [t . ha-1]
0
21.94
100% sludge
33.89
75% sludge
31.87
50% sludge
30.08
25% sludge
25.25
0% sludge
28.50
Summary
Summing up, at the beginning of vegetation period an upper layer of Jerusalem
artichoke (topinambur) canopy leaves of the treatment with 75% share of sludge
nitrogen in applied nitrogen rate showed significantly higher Performance Index
of Photosystem II. It was accompanied by higher maximum quantum field PSII for
this layer, both for treatments 75% sludge and 100% sludge. In August there was an
increase in activity of photosynthetic apparatus of the analyzed crops. This
relationship especially strongly represent data obtained for the treatment with
a sludge, where the values of P.I. were the highest for all studied layers of the
canopy of the crops. At the end of vegetation period higher values both P.I. and
maximum quantum yield of Photosystem II were observed both in crops fertilized
with a sludge as compared to the treatments 0 and 0% sludge.
Differentiated nitrogen fertilization caused a diversification of processes in
photosynthetic apparatus during Jerusalem artichoke vegetation. The rate
170 kg N ha-1 based on 75% nitrogen provided by a sludge is an optimum rate for
functioning of the photosynthetic apparatus, while the highest fresh mass field is
provided when the nitrogen fertilization is completely based upon a sludge.
Difference found for biomass yield for the above treatments was about 2 t fresh
weightha-1.
137
References
ANGELINI G., RAGNI P., ESPOSITO D., GIARDI P., POMPILI M.L., MOSCARDELLI R., GIARDI
M.T. 2001. A device to study the effect of space radiation on photosynthetic organisms.
Physica Medica - Vol. XVII, Supplement 1, 1 st International Workshop on Space
Radiation Research and 11th Annual NASA Space Radiation Health Investigators’
Workshop Arona (Italy), May 27-31, 2000.
AUGUSTYNOWICZ J.,, S. PIETKIEWICZ, M. KALAJI, S. RUSSEL. 2008A. Wpływ preparatów EM
na wybrane parametry fizjologiczne i produkcjĊ biomasy przez roĞliny energetyczne na
przykładzie słonecznika bulwiastego (topinambura). Wielokierunkowo bada
w rolnictwie i lenictwie, UR w Krakowie, 2: 9 – 24.
BCZALSKA D. 1998. Ocena moĪliwoĞci składowania skratek pochodzących z Grupowej
Oczyszczalni ĝcieków we Włocławku na miejskim wysypisku komunalnym, Mat. Konf.
Nauk. – Techn. Osady ciekowe w praktyce, Czstochowa – Ustro.
BIE J. B., 2002. Osady Ğciekowe. Teoria i Praktyka, Politechnika Czstochowska,
Czstochowa.
DYREKTYWA 2001/77/EC Parlamentu Europejskiego i Rady z dnia 27 wrzeĞnia 2001 r.
w sprawie promocji energii elektrycznej ze Ĩródeł odnawialnych na wewnĊtrznym rynku
energii elektrycznej.
GRADZIUK P. 2003. Biopaliwa, Akademia Rolnicza w Lublinie - Instytut Nauk Roln.
w Zamociu, „Wie Jutra”. Warszawa.
HE J., CHEE C.W., GOH C.J. 1996. Photoinhibition of Heliconia under natural tropical
conditions: the importance of leaf orientation for light interception and leaf
temperature. Plant Cell Environ. 19: 1238-1248.
KALAJI M. H., RYKACZEWSKA K., PIETKIEWICZ S., KOTLARSKA-JAROS E.2004A. Wpływ
dolistnego nawoĪenia siarkowo-azotowego na aktywnoĞü i rozwój roĞlin ziemniaka
[okreĞlany] metodą fluorescencji chlorofilu a. Zesz. Probl. Post. Nauk Roln. 496:
367-374.
KALAJI M.H. 2004. Chlorophyll fluorescence a: A new tool to be exploited in plant breeding
programs. Workshop: Improvement of tolerance to environmental stress and quality in
cereals. CICSA. IHAR, Radzików, Polska. 25-27.03.2004: 14-15.
KALAJI M.H. Łoboda T. 2009. Chlorophyll fluorescence to in plants’ physiological state
researches. Publisher: Warsaw University of Life Sciences -SGGW, Warsaw 2009.
KALAJI M.H., GUO P. 2008. Chlorophyll fluorescence: A useful tool in barley plant breeding
programs. In: Photochemistry Research Progress (Eds. A. Sanchez, S. J. Gutierrez). Nova
Publishers, NY, USA: 439-463.
KALAJI M.H., ŁOBODA T. 2007. Photosystem II of barley seedlings under cadmium and lead
stress. Plant, Soil Environ., 53: 511-516.
KALAJI M.H., WOŁEJKO E., ŁOBODA T., PIETKIEWICZ S., WYSZYSKI Z. 2004B.
Fluorescencja chlorofilu - nowe narzĊdzie do oceny fotosyntezy roĞlin jĊczmienia,
rosnących przy róĪnych dawkach azotu. Zesz. Probl. Post. Nauk Roln. 496: 375-383.
KAYS S., NOTTINGHAM S.F. 2007. Biology and Chemistry of Jerusalem Artichoke:
Helianthus tuberosus L. Taylor and Francis Group, Boca Raton, Florida.
KOWALCZYK-JUKO A. 2003. Topinambur, W: Kocik B. (red.), Roliny energetyczne, Wyd.
AR Lublin: 96-110.
KRUCZEK S. 1995. Dla kogo Topinambur? Top Agrar Polska 4/95.
MAJTKOWSKI W. 2003. Potencjał upraw energetycznych. W: Badania właĞciwoĞci i
standaryzacji biopaliw stałych. Mat. Seminar. Europ. Centrum Energii Odnawialnej.
IBMER, 36 – 44.
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STOLARSKI M. 2004. Produkcja oraz pozyskiwanie biomasy z wieloletnich upraw roĞlin
energetycznych. Probl.In. Roln., 45: 47-56.
STRATEGIA ROZWOJU ENERGETYKI ODNAWIALNEJ. 2001.USTAWA z dnia 26 lipca 2000 r. o
nawozach i nawoeniu. (Dz. U. Nr 89, poz. 991) i poprawki do ustawy (2004).
1
Janusz Augustynowicz,
Department of Rural Sanitation
The of Land Reclamation and Grassland Farming
Falenty, Al. Hrabska 3, 05-090 Raszyn, POLAND
2
Stefan Pietkiewicz, 2Mohamed Hazem Kalaji
Department of Plant Physiology, Agriculture and Biology Faculty
Warsaw University of Life Sciences
ul. Nowoursynowska 159, 02-776 Warszawa, POLAND
3
Stefan Russel
Free Standing of Microorganism Biology, Agriculture and Biology Faculty
Warsaw University of Life Sciences
ul. Nowoursynowska 159, 02-776 Warszawa, POLAND
139
140
CHAPTER X
Wojciech Dbrowski
TREATMENT AND FINAL UTILIZATION OF SEWAGE
SLUDGE FROM DAIRY WASTE WATER TREATMENT
PLANTS LOCATED IN PODLASKIE PROVINCE
Introduction
From the beginning of the 90-s of the latest century, there is observed the
development of dairy processing plant on the north east part of Poland. The
accession of Poland to European Union has had the impact on this process. On the
one hand, it causes economic growth of the region, but on the other, it increases the
danger for natural environment caused by industrial plants. According to researches
conducted in the period of 1998-2000, the amount of treated sewage in Podlaskie
province reached about 138 000 m3d-1 among others - 9070 m3d-1 was treated by
individual dairy systems. The amount of sewage sludge generated during the whole
year in Podlaskie province, reached 19600 tons d.m., among others 1200 tons d.m.
were produced by dairy plants (Boruszko and others, 2000). While analysing
problems connected with the amount of sewage and sewage sludge in Podlaskie
province in 2008, there was observed the increase of dairy sewages, which are
treated in individual dairy waste water treatment plants in the province. According
to the data of the author, this amount reached 12 000 m3d-1. While assessing the
quantity of dairy sewage, it is necessary to take into account the fact that during last
years the rate of the used water and generated sewage decreased in relation to the
amount of processed milk. The changes, which were observed in individual dairy
waste water treatment plants, are proved by such parameters like personal equivalent
(P.E.) or the amount of sludge produced during sewage treatment. The quantity of
sludge in dairy waste water treatment plants rose from 1140 tons d.m. in 1998 to
almost 3700 tons d.m. in 2008. In the biggest plant located in the town
of Wysokie Mazowickie (Mlekovita Diary Cooperative) there was noticed the
increase of generated sludge from 600 to almost 2200 tons d.m. in analogous period
of 10 years. Sewage sludge is the byproduct in the process of sewage treatment, the
way of its finale utilization depends on many factors among others physico-chemical
composition of sewage which is put through the treatment process and the method of
its processing. On the account of the law, sewage is the waste, however while
meeting the criteria (ROZPORZDZENIE, 2002) it can be the essential product, which
will come back to the environment in safe form. The quantity of sewage sludge
among others dairy sludge, will rise together with the load of sewage.
141
Figure 1 shows the increase of the stream of sewage sludge predicted in National
Sewage Treatment Programme.
Fig. 1 The forecast of sewage sludge quantity [Mg d.m. y-1] in Poland up to 2015
according to National Sewage Treatment Programme
The quantity of sludge in Poland in a period of 2000-2010 will rise almost twice
that is why its treatment and final utilization will be the main current problem within
the next years. It considers also the sludge produced in dairy wastewater treatment
plants. It is differentiated by physico-chemical composition in relation to the sludge
formed in municipal waste water treatment plants.
Dairy waste water treatment plants, characteristics of research base
In Podlaskie province there exist nowadays nine plants using individual waste
water treatment systems. On the one hand, the quantity of treated sewage is not high
in relation to the quantity of municipal sewages but on the other hand, taking into
account pollution load, there is easily seen the impact of them on the state of surface
waters which are receiving waters of treated sewage. Sewage sludge produced
during the process of dairy sewage treatment is used to fertilize soils.
Table 1 shows the basic parameters of chosen systems of dairy sewage treatment
plants in Podlaskie province according to the data from 2008. The analysis covered
the largest dairy objects, which use individual wastewater treatment plants. The size
of these plants is proved by the fact that 7 from 9 analysed objects work on the basis
of integrated pollution prevention permission (IPPC).
Apart from the quantity of sewage and sewage sludge there was given personal
equivalent (P.E.) characteristic for each object describing the level of load, which is
treated by dairy waste water treatment plants. The average BOD5 in dairy sewage is
142
about 6 to 10 times higher than in case of municipal sewage. It is proved by the own
research and also by literature (B.A.T., 2005, RUFFER 1998).
Table 1
Characteristics of chosen dairy W.W.T.P-s in Podlaskie province
Plant
Sewage
quantity
3
m d
-1
P.E.
Sludge
amount
-1
Mg d.m. y
Wysokie
Mazowieckie
5500
277000
2200
Bielsk
Podlaski
700
9800
230
Grajewo
1800
41300
420
Kolno
730
31800
220
Zambrów
697
17000
90
Sejny
800
7100
130
Moki
600
15000
80
Pitnica
1300
35100
250
Suwałki
700
8000
25
Source: own researches
143
Waste water and excess
sewage sludge treatment
Intensive biological and chemical
removal of C,N,P. Aerobic sewage
sludge stabilization in separate chamber,
filter press dewatering
Aerobic activated sludge system, aeration
ditches, chemical phosphorus removal.
Simultaneously aerobic stabilization,
gravitational thickening
Sludge activated system (Promlecz) ,
chemical phosphorus removal.
sewage sludge dewatering with mobile
centrifuge
Aerobic activated sludge system, aeration
ditches. Simultaneously aerobic
stabilization, sewage sludge dewatering
with mobile centrifuge
Aerobic activated sludge system.
Separated aerobic stabilization, sewage
sludge dewatering with mobile centrifuge
Activated sludge in aerobic system,
aeration ditches, chemical phosphorus
removal. Simultaneously aerobic
stabilization, sewage sludge dewatering
with mobile centrifuge
filter press dewatering
filter bed dewatering
Activated sludge system (Promlecz),
chemical phosphorus removal.
sewage sludge dewatering with filter bed
Dairy waste water treatment plants, which worked to the middle of 90-s of the
last century, used the method of activated sludge mainly in the form of two stage
chambers of activated sludge (the chamber of high and low loaded activated sludge).
These systems were not initially adapted to intensive nitrogen and phosphorus
removal, because there was not required by current law restrictions. In the process
of treatment there were not used intensive biological and also chemical methods.
The original Polish solution is an activated sludge chamber of “Promlecz” type and
“Potap” aerations – the project of the Office of Studies and Investments Realisation
in Dairy Industry (patented in Poland and abroad) (Piotrowski, 1982). Sludge
stabilization process was done simultaneously thanks to long periods of aeration and
low load of activated sludge chambers. Dairy plants working in 70s and 80s
of the last century were characterised by high production changeability during the
whole year. On account of heavy decrease of production in a period from fall to
spring, dairy waste water treatment plants used only a part of appliances to treat
sewage and sewage sludge. In the 90s of the XX-th century the situation changed,
the plants did not register the sharp fall of production except of summer time. To the
end of the XX-th century, sludge dewatering was conducted only with the use
of filter bed, which effectiveness depended; to the large extend, on atmospheric
conditions. Taking into account high increase of sludge, filter beds are used
nowadays to gather treated sludge before its final use. The utilization of natural
methods of sludge dewatering needs very large surface, while dairy waste water
treatment plants are usually located near localities. Offensive smell connected with
sewage sludge dewatering has negative impact on the level of habitants’ life, the use
of mechanical systems gives the possibility of the utilization of more commonly
seen deodorising devices of sludge draft. Dairy waste water treatment plants which
used Polish solutions from the 70s of the last century, still work beside of new
systems created for intensive removal of carbon, nitrogen and phosphorus
compounds from sewage. The potential of dairy waste water treatment plants
working on Podlaskie province are much differentiated. The oldest dairy waste
water treatment plant located in Bielsk Podlaski has been working over 30 years,
within the modernization there has been introduced only chemical phosphorus
removal from sewage. It is essential to underline that this object complies with
binding regulations, but also very significant is the experience of users who are able
to conduct effective sewage treatment and sewage sludge utilization. The dairy
waste water treatment plants which are analysed in Table 1, use many different ways
of sewage and sewage sludge treatment. The shared feature is the utilization
of activated sludge method to sewage treatment and aerobic stabilization of excess
sludge. Among nine analysed waste water treatment plants, only one object can be
an example of modern system of sludge treatment worthy the XXI century. Dairy
waste water treatment plant in Wysokie Mazowieckie is the biggest one of this
object type in Poland but also in Europe. In summer period this plant processes over
two millions of litres of milk per day. To 2000, this waste water treatment plant
worked according to typical system of Promlecz with simultaneously stabilization
and dewatering sludge with filter beds, which were commonly used in the beginning
of the 90-s of the last century. After modernization and introduction of intensive
biological and chemical sewage treatment, the amount of sewage sludge rose over
144
twice. The increase of sludge reached 5200 kg of sludge dry matter per day on
average reaching the increase rate on the level of 0,46 kg d.m.kg-1 BOD5 (Kajurek
2005). There were used separate chambers in aerobic stabilization, which processed
mechanically thickened sludge. Stabilization time ranged between 5 to 8 days, the
process is exothermic and the stabilization temperature reached 30-36 ºC. In order to
limit the temperature increase in chambers and to provide suitable air change, under
the cover of each chamber there is pressed air in amount of 2.5 thousands m3h-1.
After stabilization process, sludge is dewatered with filter press. There is possibility
of additional lime stabilization. At the moment, on account of production increase in
this plant, it is necessary to modernize both sewage treatment line, but also sewage
sludge treatment. In Podlaskie province there has not been used anaerobic
stabilization of dairy sludge. These solutions were used in the largest municipal
waste water treatment plants in the region, which is connected with the energy
harvesting from produced biogas. The course of sludge stabilization process is
influenced by the content of organic matter and the composition of reject water
(DBROWSKI, 2006, 2008). In case of dairy waste water treatment plant in Wysokie
Mazowieckie it is possible to use anaerobic sewage reactor, as the first stage of
treatment, and later typical aerobic system. It is also connected with the possibility
of energy recovery and total change of the way of sewage sludge treatment. The use
of sewage treatment processes and sludge treatment with energy recovery must be
stimulated by appropriate regulations and must be economically explained.
However, it will not change the final sludge use – nowadays sludge from dairy
waste water treatment plants owned by Mlekovita and Mlekpol (the two biggest
producers in Poland) is used to fertilize soil. Farmers who are the members
of cooperative society or those who have contracts on milk delivery use stabilized
sludge as a fertilizer. This kind of sludge utilization gives the guarantee that it is
safely used. Sludge producer is obliged to make not only periodical sludge
examinations but also soils tests before and after fertilization by sludge
in accordance with the order on municipal sewage sludge.
Research conditioning
Sewage sludge composition and its sanitary state are two basic elements, which
decide about agricultural utilization of municipal and industrial sludge in accordance
with current order on municipal sewage sludge. Equally essential is the composition
of soils on which sludge from dairy waste water treatment plant can be used as
beneficial fertilizer. In analysed sludge samples, there was determined the content
of lead, copper, cadmium, nickel, zinc and chromium but also nitrogen, phosphorus,
magnesium and calcium. Organic matter content was also determined in order to
assess the level of diary sludge stabilization. The range of metals research shown in
table 2, 3, and 5 is connected with the order on municipal sewage sludge mentioned
above, which describes research range on account of stabilized sludge management,
as raw material, not as waste (ROZPORZDZENIE 2002). Sludge was mineralised
with the use of microwave system Mars 5 in accordance with EPA 3015 and EPA
3051 procedures. Metals determination, except of mercury, was done with the use of
emission spectrometry with inductively stimulated plasma, mercury was determined
145
by atomic absorption spectrometry on AMA-254 analyser. Macro elements
determination in sludge was done in accredited laboratory in accordance with PNEN ISO 11885 standard, with the use of optical emission spectrometer with
inductively stimulated plasma – spectrometer of Varian Vista MPX Company.
The results shown in Table 2 were compared with the permissible limits. In case
when sewage sludge is used naturally and also agriculturally, the strictest criteria
apply to chosen heavy metals. There was shown also the composition of chosen
fertilizers used in agriculture (tab. 4). The examinations were conducted during
working out of environmental impact statement on sludge management from dairy
waste water treatment plant in Wysokie Mazowieckie in 2001 (DBROWSKI, 2003).
In a period of 1998 – 2000 within the project “Water, sewage and sewage sludge in
waste water treatment plants in Podlaskie province”, there were carried out the
researches of municipal sewage sludge. Regional Environmental Protection Fund in
Białystok financed this project. It covered all waste water treatment plants in
Podlaskie province. The research results of municipal sludge were compared with
the examinations of sludge quality from two biggest meat-processing plants in
a period of 1999-2001, which had individual waste water treatment systems and
from dairy waste water treatment plants in a period of 1998-2002. The comparison
of sludge research results from dairy, municipal sewage plants and chosen natural
fertilizers prove the usefulness of dairy sludge utilization to fertilize or make
reclamation of soils on the area of Podlaskie province.
Characteristics of sewage sludge
In dairy plants of Podlaskie province, the production but also sewage load and
sewage sludge increases. In the largest analysed plant in Wysokie Mazowieckie,
there was observed the rise of sewage load determined by BOD5 by 30% in a period
of 2004-2008. The increase of load influences mainly the rise of following
parameters: BOD5 and COD5 and the quantity of produced sewage in small extend.
It is typical for sewage plants in Podlaskie province, taking into account the decrease
of individual water use per product unit. According to authors’ examinations water
use rate was within the range of 1.3 to 4.2 m3m-3 of processed milk in a period
of 2004-2005. The amount of sewage fluctuated between 1.8 to 4.3 m3m-3 of milk.
These rates did not change sharply in 2008. On the basis of the analysis of dairy
waste water treatment plants in Podlaskie province, the quantity of generated sludge
amounted from 0.13 to 0.45 kg d.m. per 1 m3 of treated sewage. According to the
data from 2005, in case of municipal waste water treatment plants, this indicator
reached 0.247 kg d.m. per 1 m3 of sewage.
The results of the analysis presented in Table 2 show that heavy metals content
in sludge produced in diary waste water treatment plant is low, sharply below the
limit values, which allows to use sludge as fertilizer in some crops. Lead content in
all analysed sludge ranged between 3.2 to 19.9 mg Pbkg-1 d.m. alongside the limit
quantity of 500 mg.kg d.m. if sludge is used as fertilizer. In comparison, the average
lead content in Polish soils used agriculturally amounts 13.6 mgkg-1 d.m. ( Mocek
2002) alongside the range of 3.6 to 42 mgkg-1 d.m. (Łukowski, 2009). In case
of zinc, higher level of this element was observed in sludge from milk plant in
146
Suwałki. This situation can be caused by the fact that rain watere comes to dairy
waste water treatment plant from the area around the plant. Moreover, the higher
zinc content in dairy and municipal sludge is caused by industrial installations made
of zinc-plated steel. The content of the rest of metals like copper, chromium,
cadmium, nickel and mercury, was similar in all sludge from analysed plants. The
low heavy metals content was in dairy sludge, which was analysed in Lugo province
in Spain. The average heavy metals content in sludge (mgkg-1 d.m.) amounted: in
case of chromium 15.99, nickel 11.04, copper 58.55, zinc 289.74, cadmium 0.11,
mercury 0.08 and lead 10.05 (MOSQUERA LOPEZ M.E., 2000). In table 5 there were
shown the research results of municipal sludge together with the previous sludge
examinations from industrial sewage plants of Podlaskie province. It was stated that
heavy metals content of sludge in municipal waste water sewage plants was sharply
higher than the values characteristic for dairy sludge shown in table 2. In case
of meat plants, which have their own waste water treatment plants, there was stated
similar metals content in sewage sludge, only zinc content was on lower level.
Tanning plants, which use chromium technology of hide tanning, caused high
chromium content in sludge from municipal waste water treatment plants. Moreover,
in case of municipal sewage plants, there was observed high zinc content in sludge.
It is typical for sludge from large municipal waste water treatment plants in
Podlaskie province (Białystok, Łoma, Suwałki). Substantial impact on high
contamination of municipal sludge by zinc has the fact that municipal sewage plants
take rain water in case of combined sewerage system.
Low metal content is one of criteria, which conditions the possibility
of recycling of dairy sludge to environment. The legislator determined also the
characteristics of soils on which can be used sludge divided into light, medium and
heavy. Permissible dose must be determined on the basis of examinations and
counting, the limit value is the utilisation of 5 tons d.m. of sludge in a period
of 10 years. Except of determination of metals content in sludge and soils before its
utilization, it is necessary to monitor soils also after its use. Dairy sewage sludge
contained similar quantity of heavy metals in comparison with natural fertilizers
(Table 4). In Table 3 there were shown research results of theses metals in sludge
like nitrogen, phosphorus, magnesium or calcium. These are essential parameters,
which prove the fact that sewage sludge can be used to fertilize or soils reclamation.
Filipek and Fidecki presented similar examination results while analysing sludge
from dairy waste water treatment plants. According to their researches, magnesium
content fluctuated between 4.5 to 6.2 gkg-1 d.m., while calcium from 3.0 to
46.9 gkg-1 d.m. (FILIPEK 1999).
The separate element conditioning sewage sludge utilization is their sanitary
state. The legislator determined the necessity of sludge examinations on pathogenic
bacteria in a type of Salmonella and the quantity of alive eggs of following
helminths: Ascaris sp., Trichuris sp., Toxocara sp. Precise examinations conducted
in two dairy waste water treatment plants in Podlaskie province in 1998 showed that
sludge from dairy waste water treatment plants can be safely used in agriculture in
process of fertilisation and soils reclamation.
147
The installation of sanitary sewage sludge stabilization in Wysokie Mazowieckie
was done but after 2000, there was no necessity to use it. In case of the majority
of dairy waste water treatment plants these devices are not installed.
Table 2
Heavy metals content in sewage sludge, from dairy W.W.T.P-s
Plant
Pb
Quantity of heavy metals mg.kg-1 d.m.
Zn
Cu
Cd
Ni
Cr
Hg
Wysokie
Mazowieckie
Bielsk Podlaski
Grajewo
Kolno
Zambrów
Sejny
Moki
Pitnica
Suwałki
10.2
170
22.40
0.52
3.10
4.60
0.18
5.8
19.9
12.6
8.1
10.0
3.2
7.1
9.0
163
207
139
234
240
150
410
675
20.00
22.50
27.00
28.00
26.00
20.00
62.10
7.70
0.40
0.45
2.30
0.60
0.80
0.15
0.84
0.50
3.30
17.70
13.90
9.10
1.90
6.20
14.00
3.70
4.30
13.30
14.20
9.60
2.10
9.60
8.80
8.50
0.19
0.32
0.16
0.26
0.06
0.20
0.10
0.03
Maximum accepted
for agriculture reuse
500
2500
800
10
100
500
5
Table 3
Biogenic compounds content and organic substances in sewage sludge from dairy
W.W.T.P-s
Chosen characteristic parameters
Plant
Wysokie
Mazowieckie
Bielsk Podlaski
Grajewo
Kolno
Zambrów
Sejny
Pitnica
Moki
Suwałki
N-total
g kg-1 d.m.
P-total
g kg-1 d.m
93.6
17.0
26.9
31.0
71.0
93.5
69.0
62.7
60.0
20.8
1.9
10.4
2.5
48.8
2.0
36.0
8.2
5.3
.
.
Ca
g kg-1 d.m
Organic
substances
%
3.9
28.0
82.1
6.8
1.2
5.9
5.7
4.2
24.7
2.1
4.5
61.9
24.8
42.3
41.3
18.0
73.3
18.0
47.8
74.2
67.0
31.2
72.0
61.0
82.8
64.0
74.2
Mg
g kg-1 s.m
.
.
Dairy production is connected with many sanitary obligations, material (milk)
and water is examined, treatment process is monitored on account of sanitation. It is
translated into sanitary quality of sewage and later on sewage sludge quality.
Different situation is observed in municipal dairy waste water treatment plants
where sewage sludge goes through sanitary decontamination, more often there are
used also the processes of thermal processing of sludge to limit its capacity and
provide sludge stabilization.
148
Contents of heavy metals in chosen organic fertilizers (mgkg-1 d.m.)
Fertilizer
Cow liquid manure
Swine liquid manure
Manure
Pb
11
11
17
Cd
0,46
0,82
0,1
Cr
5,4
9,0
22,0
Cu
45
294
27
Ni
3,8
11,0
16,0
Table 4
Hg
Zn
0,05
0,04
0,10
222
896
190
Table 5
Heavy metals contents in municipal and industrial sludge from W.W.T.Plants of Podlaskie
province - max. value 1996-2002
Type of W.W.T.P,
research period
Dairy waste water
treatment plants,19982002
Meat industry waste
water treatment plants,
1999-2001
Municipal waste water
treatment plantsPodlaskie province,
1998-2000
Heavy metals content (mgkg-1 d.m) – maximum value
Pb
Zn
Cu
Cd
Ni
Cr
Hg
19,0
48
26
0,80
12,0
19,0
0,38
7,0
80
136
1,4
19,0
21,0
0,2
94
1436
136
4,9
25
1000
5,15
Summary
Sewage sludge produced in analysed dairy waste water treatment plants are
exposed to recycling. They come back to environment in a form of fertilizer,
because they comply with the requirements of the order of 2002. According to
European hierarchy waste management, the most preferable is prevention of their
forming, reuse and recycling.
In case of dairy or municipal sewage sludge, the prevention consists in limit of
produced sludge. On account of the fact that sludge quantity depends on the capacity
of sewage load and technology of their treatment, it is difficult to limit their amount
on the level of treatment process. According to recommendations shown as the Best
Available Technology (B.A.T.) for food industry, in case of milk plant the most
important is high quality and product safeness. Less important are activities, which
limit water use and sewage production in comparison with processed material unit,
which means milk.
The current order on water quality carried to receiving water, demands high
requirements from treated sewage, which translates into intensive technology of
removal of carbon, nitrogen and phosphorus compounds. Moreover the use of
modern intensive treatment methods causes the increase of sludge quantity. On
account of low metals content in dairy sludge, high content of nitrogen, phosphorus,
149
calcium and magnesium and the lack of sanitary danger, there is no alternative for
sludge recycling to the environment in a form of for example fertilizer. In Podlaskie
province 96,91 % of soils agriculturally used have natural heavy metals content,
while only 0,03% of them can be classified to second degree of contamination,
which also proves the necessity of sludge recycling coming not only from dairy
waste water treatment plants (Terelak, 2001). There is no explanation for the
utilization of expensive thermal processes in sludge treatment, which are more often
used in Poland, in case of municipal waste water treatment plants. The situation can
be changed after introducing methods of anaerobic treatment of dairy sewage, which
can take place in case of the largest objects. Among analysed plants, this situation
can consider the ones in Wysokie Mazowieckie, Grajewo and Pitnica.
The experiences gathered from introducing of new technologies can be used by the
whole dairy industry and similar food plants, which have individual waste water
treatment plants.
References
BORUSZKO D.,DBROWSKI W., MAGREL L. 2000: Bilans Ğcieków i osadów Ğciekowych
w oczyszczalniach Ğcieków województwa podlaskiego, Fundacja Ekonomistów
rodowiska i Zasobów Naturalnych., Białystok: pp.44.
DBROWSKI W. 2003: Rolnicze wykorzystanie osadów Ğciekowych na przykładzie
województwa podlaskiego, Gospodarka Odpadami Komunalnymi, Kołobrzeg-KopenhagaOslo: 191-199.
DBROWSKI W. 2006: Management and utilization of sewage sludge from dairy industry
wastewater treatment plants; IWA Specialized Conference: state of the art , challenges
and perspectives, Moscow: 734-737.
DBROWSKI W. 2008: Oczyszczanie odcieków z przeróbki osadów w oczyszczalni Ğcieków
mleczarskich, Inynieria i Ochrona rodowiska, Wydawnictwo Politechnika
Czstochowska, 11, 1: 115-122.
FILIPEK T., FIDECKI M. 1999: Ocena przydatnoĞci do nawoĪenia osadu Ğciekowego
z mleczarni w Krasnymstawie. Univ. Agric. Stetinesis 200. Agricultura 70: 87-92.
KAJUREK M. 2007: Studies on heavy metals contents changes during treatment of sewage
from dairy wastewater treatment plant, Polish Journal of Environmental Studies, 16, 2A,
part III: 665-668.
LOPEZ-MOSQUERRA M.E., MORION C., CARRAL E. 2000. Use of dairy sludge as a fertilizer
for grasslands in northwest Spain: heavy metals levels in the soil and plants, Resources,
Conservation and recycling 30: 95-109.
ŁUKOWSKI J. 2009. Wpływ odpadów organicznych i mineralnych na mobilnoĞüi
biodostĊpnoĞü metali ciĊĪkich w Ğrodowisku glebowym, rozprawa doktorska, Uniwersytet
Technologiczno Przyrodniczy w Bydgoszczy: 58-68.
MOCEK A. 2002. Stopnie skaĪenia gleb Polski metalami ciĊĪkimi. Journal Res. Apel. Agr.
Eng. 47: 29-34.
NAJLEPSZE DOSTPNE TECHNIKI (BAT). 2005. Wytyczne dla branĪy mleczarskiej.
Ministerstwo rodowiska: 1-46.
PIOTROWSKI J., PASTERNAK T. 1980. Oczyszczanie Ğcieków mleczarskich w kraju
i za granicą, Przegld Mleczarski , 1: 23-27.
ROZPORZDZENIE MINISTRA RODOWISKA z dnia 1 sierpnia 2002r. W sprawie komunalnych
osadów ciekowych, Dz. U. Nr 134, poz. 1140.
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RUFFER H., ROSENWINKEEL K. H. 1998. Oczyszczanie Ğcieków przemysłowych, ProjprzemEKO, Bydgoszcz: 164-178.
TERELAK H., MOTOWICKA- TERELAK - T., STUCZYSKI T., PIETRUCH CZ. 2000. Pierwiastki
Ğladowe w glebach uĪytków rolnych Polski. Inspekcja Ochrony rodowiska, Biblioteka
Monitoringu rodowiska, Warszawa: 1-69.
Wojciech Dąbrowski
Department of Technology and Environmental Protection
Technical University of Bialystok
ul. Wiejska 45B, 15-351 Bialystok, POLAND
151
152
CHAPTER XI
Joanna Kostecka
SELECTED ASPECTS OF THE SIGNIFICANCE OF
EARTHWORMS IN THE CONTEXT OF SUSTAINABLE
WASTE MANAGEMENT
Introduction
According to MILLENNIUM ECOSYSTEM ASSESSMENT (2005) submitted by the
General Secretariat of the United Nations, the state of about 2/3 of services provided
by the World’s ecosystems to Man is deteriorating. It has happened as a result of
over exploitation and a loss in the variety of species which would otherwise
guarantee the stability of ecosystems (some of the consequences of the deterioration:
the decline in fish stock, loss of soil fertility, decreased number of pollinating
insects). One of the main reasons for the ecological problems of the World is the
result of pollution and the presence of various toxic substances in water, soil and air.
In ecosystems the concentration of some of those substances is too high due to either
natural processes or high anthropogenic pressure. The changes of ecosystems are
related to soil acidification, the pollution of groundwater, the eutrophication of
surface water, radiation and the greenhouse effect. The depletion of resources,
excessive noise and electromagnetic field have a negative effect on Man’s situation
which has this far been stable (ALBISKA 2005, SIEMISKI 2007, DOBRZASKA et al.
2008).
Nowadays people have started to realise that it is no longer enough to undertake
only superficial actions in order to compensate for the damage caused by the
implementation of various communication, urban or industrial projects. There is a
need for complex system solutions which are also important for future generations.
This way of thinking is fundamental for the concept of sustainable development. Its
supporters call for significant civilizational changes on an ecological, social and
economic level. A wide scope of those changes gives one the right to formulate a
postulate which will state that a new vision of development can reach the status of a
revolution which can be compared to the revolutions so often mentioned in the
history of mankind, i.e. the agricultural, scientific and industrial revolutions
(Table 1).
PAWŁOWSKI (2009) suggests that a fundamental discussion about sustainable
development should be enriched by ethical, technical, legal and political aspects as
well as by the hierarchization of problematic groups out of which morality is
153
considered to be a fundamental problem as without it the sustainable development
revolution will not be successful.
Nowadays, at the turn of the century, we need to consider a number of local,
regional and global problems that are related to a social, economic and
environmental sphere. The solution to those problems is very complicated and
requires the co-operation of politicians, economists, sociologists, biologists,
entrepreneurs representing different countries as well as every citizen. On the other
hand, SKUBAŁA (2008) thinks that the concept of sustainability has its origins in the
idea that we borrowed the Earth from our grandchildren.
Table 1
Critical phases in the process of the development of mankind
(based on PAWŁOWSKI 2009 - changed)
1
2
The development phase
Period
Hunting and picking
The agricultural
revolution
Upper Palaeolithic
The beginning: about 9000 years ago in Asia, in Europe
about 4000 years later
The symbolic beginning: the publication of “On the
Resolution of the Heavenly Spheres” by Nicolaus Copernicus
(1543); the explication,
the publication of “Mathematical Principles of Natural
Philosophy” by Isaac Newton (1687)
The symbolic beginning: a significant modification of a
steam engine by Watt (1769). The next phase (1860-1914):
the beginning of using crude oil (an internal combustion
engine) and electricity
Significant events: the U’Thant’s speech (1969), the
definition of sustainable development (Bruntland’s
definition) introduced by the United Nations (1987), the
United Nations Conference “Earth Summit” in Rio de Janeiro
(1992), the announcement of the United Nations Decade of
Education for Sustainable Development (2005-2014)
3
The scientific revolution
4
The industrial revolution
5
The sustainable
development revolution
In Poland, we can consider the following as some of the most important
achievements in the implementation of the ideas of sustainable development: the
STATE ENVIRONMENTAL POLICY (1990), the CONSTITUTION OF THE REPUBLIC OF
POLAND (1997) which accepted the regulation that the ideas of sustainable
development would be implemented (the article 5), the binding multidimensional
strategic plan Poland 2025 (RZDOWE CENTRUM STUDIÓW STRATEGICZNYCH,
MINISTERSTWO RODOWISKA 2000) as well as constantly improved and updated
environmental protection policy (GRUSZECKI 2008). The survey shows that despite
the announcement of the United Nations Decade of Education for Sustainable
Development, the idea of sustainable development is still under-acknowledged
(KOSTECKA 2007, 2009).
154
Sustainable development and the problem of waste management
Among the issues described above there is another problem which was noted in
the last decade of the 20th century – a constant increase in the volume of municipal
waste which can be explained by the rapid development of civilization and a higher
standard of living (ROSIK-DULEWSKA 2007). As waste has a negative effect on the
environment, rational and pro-environmental ways of waste management need to be
urgently considered and implemented. Otherwise, all the elements of ecosystems,
i.e. soil, water, atmosphere and consequently Man will continue to suffer
(KOZŁOWSKI 2000, SIEMISKI 2007). The redirection of the waste management
strategy towards environmental protection must be a characteristic for the 21st
century (KEMPA 2001, BARAN, DROZD 2004). It is due to the statistical data which
shows that Poland alone produces 9354 tonnes of municipal waste annually
(CONCISE STATISTICAL YEARBOOK 2008). However, according to the authors of the
National Waste Management Plan (2010), the real amount of waste is at least 10%
larger.
Every citizen should contribute to the implementation of the pro-environmental
waste management. In accordance with the WASTE MANAGEMENT ACT (Journal of
Laws 2001, no. 62, pos. 628) every citizen must take broad actions in order to avoid
generating waste. Resulting waste will need to be sorted at recycling centres
provided by appropriate authorities (in Poland, at the moment, it is only possible to
sort metal, plastic, glass and paper). There should also be facilities for the separation
of hazardous waste from a waste stream. Apart from the legal regulations (the
WASTE MANAGEMENT ACT, art. 10 – a legal requirement to segregate waste), the
difference in rates for having non-segregated and segregated waste collected might
be another factor that would motivate citizens, who are not aware of ecological
issues, to segregate waste. However, in order to fully implement the proenvironmental principles of waste management we need to take more responsibility
for the education and the ecological culture of all citizens.
The presence and the disposal of waste has been a problem for some time.
Nowadays it is far more serious as it is related to growing urbanization and overconsumption. In Poland, modern waste management was neglected for a long time
both by the representatives of the central and provincial offices, and by the
communal authorities responsible for environmental protection and citizens. That is
why an urgent reconstruction of the waste management system towards sustainable
solutions is an important issue.
In order to reduce the amount of waste, the Poles need to considerably change
their attitudes and behaviour, e.g. they should gradually eliminate purchases of
products whose production, exploitation and then resulting waste has a negative
effect on the environment. In order to achieve this we need to constantly make
customers aware and there must be cooperation between producers and those
responsible for waste processing. In households we need to avoid products with
unnecessary packaging or instead use returnable packaging. We also need to
compost organic waste as well as some forms of packaging. The reduction of the
amount of waste in industrial factories requires long-term actions. A cleaner
production means not only an investment, as in many cases there would be a need
155
for the whole technological lines to be changed (PRZYWARSKA 2005) but also, as it
has been shown in the survey, changing inappropriate attitudes of people which
usually accounts for 2/3 of obstacles in the implementation of the principles of
environmental protection (SKALMOWSKI 2007).
Nowadays, we know that a suitable location of a well equipped dump site can
considerably limit its negative effect – the pollution of surface and ground water and
the pollution of soil and reduce the danger of microbiological, dust and odour
contamination. However, the problem that still exists in Poland is not only the lack
of acceptance of our participation in the waste management system but also the lack
of acceptance of the necessity of waste segregation which makes it possible to reuse
it (KOZŁOWSKI 2000, BARAN, DROZD 2004, JDRCZAK 2007, ROSIK-DULEWSKA 2007).
Why you should not dispose of organic waste at landfill sites and illegal
dump sites
The WASTE MANAGEMENT ACT (Journal of Laws 2001, no. 62, pos. 628)
formulates the principles of dealing with waste, which guarantees the protection of life
and health of Man as well as environmental protection in accordance with the ideas of
sustainable development. The regulations included in the act are to prevent the
production of waste, force limits on its amount and its negative influence on the
environment and make recycling and neutralizing easier.
The main aim of the new regulations, which meet the UE standards within the scope
of waste management, is to limit the prevailing method of neutralizing waste which
is its disposal at landfill sites (KEMPA 2001, GRUSZECKI 2008).
In Poland, similarly to the majority of developed countries, we can observe a
constant increase in municipal waste production. According to the NATIONAL
DEVELOPMENT PLAN for the period of time 2007-2013 and the STATE
ENVIRONMENTAL POLICY for the period of time 2009-2012, the weight of waste
produced by every citizen reaches 300 kg annually. However, it is still half as large
as in the richest EU countries.
The vast majority of municipal waste (in 2007 – 9609 tonnes) (STATE
ENVIRONMENTAL POLICY …) is still disposed of in landfill sites. A significant part
of that mass (at least 30% on average) is biodegradable which, in the form of waste,
is a serious threat to all the elements of the environment. On the other hand
however, it is a potential material for the production of compost which increases soil
fertility.
We need to remember that landfill sites which contain organic substance have
perfect conditions for all living organisms – especially for those traditionally
regarded as pests to Man: rodents, insects and some species of birds. As bioton gets
heated, such animals not only find food but also warm hiding places all year round.
Their large populations are very mobile, they move long distances from a place to
place spreading pathogenic elements.
In the situation where Poland, as well as other EU countries, has a duty to obey the
LANDFILL DIRECTIVE 99/31/EC, which requires the gradual reduction of organic
waste mass stored at landfill sites until 2025, every idea of neutralizing such kinds
156
of waste (including in places where they are produced) should be recognized,
widespread and constantly improved.
As numerous studies show, one of the successful methods of neutralizing
organic waste is vermicomposting. The use of vermiculture biotechnology in
relation to selected biowaste makes it possible not only to reduce the volume of
organic waste disposed of at landfill sites but also to gain two important products:
organic fertilizer – vermicompost and earthworm body walls whose nutritional value
is high. Vermicomposting is a widely used method and it is still multidimensionally
examined all over the world including the search for the importance of the use of
various species of earthworms (EDWARDS 1998, DOMINGUEZ et al. 2001, BORGES et
al. 2003, DICKERSON 2004, KOSTECKA 1994; 2004A, PARVARESH et al. 2004,
SELDEN et al. 2005, SHARMA et al. 2005; GARG et al. 2006, KOSTECKA, PCZKA
2006).
From the facts presented above we can tell that actions towards the organic
waste segregation and the pro-environmental neutralization of waste should become
more important in the future strategy for sustainable waste management. According
to the LANDFILL DIRECTIVE UE 99/31/UE, it must be a common responsibility of all
citizens and communal authorities. However, as studies show, not even all decision
makers are aware of it (KOSTECKA et al. 2007).
After segregating biowaste from a waste stream, the most justifiable way of its
neutralization is anaerobic fermentation with biogas recovery, aerobic composting
(or vermicomposting) and, when it is required, combustion with energy recovery.
Neutralizing segregated organic waste can take place on various scales: at an
appropriate installation, at a municipal, communal compost sites or household
compost areas (KOSTECKA 1998, 2000, KASPRZAK 2001, JDRCZAK 2007).
In a situation where waste is stored as the disposal of bioton at communal
installations or having compost areas in a garden or an allotment is not possible,
some citizens could be convinced to use a specific (unconventional) form of the
organic waste management. “Earthworm ecological boxes” would allow the
management of wastes in a place where they are generated (as vermiculture on a
small scale, in boxes in a “handy” place e.g. a balcony, a kitchen or a basement)
(APPELHOF 1982, 1993, KOSTECKA 2000). It is also possible to vermicompost office
wastes (KOSTECKA 2003).
The conditions of a study
The study of earthworm ecology and the use of earthworms in the neutralization
of organic waste have been conducted at the centre in Rzeszów since 1986
(previously the branch of University of Agriculture in Kraków and nowadays the
University of Rzeszów). The aim of the publication is to present those parts which
refer to the sustainable waste management against the background of the selected
studies conducted in Poland and abroad.
In the individual study (conducted in a laboratory as well as on a semi-technical
and technical scale) of the neutralization of organic waste in vermiculture, we
included agricultural wastes (cattle and horse manure as well as post-harvest
residues), sewage sludge from several municipal sewage treatment plants, cellulose,
157
office and household wastes. A concentrated population of earthworm Eisenia fetida
fetida (Sav.) has been used in the study. Their life functions, observed in substrate which
contained the above mentioned wastes, made it possible for coprolitic fertilizer, also
known as vermicompost, to form. The results of the study have also been used in broad
educational actions.
Selected aspects of issues concerning modern vermiculture and its
importance
The conditions of Northern Europe and Poland are ideal for a concentrated
population of the above mentioned earthworm Eisenia fetida fetida (Sav.) to be used
in the vermicomposting of organic waste (KOTOWSKI 1989, KOSTECKA 1994,
KASPRZAK 1998). We can find this geopolitical species in a soil surface layer where
organic waste accumulates. It was bred on the American continent in the 1950s
(BOUCHE 1987; EDWARDS, BOHLEN 1996). Nowadays, we can breed it on a
technical scale (EDWARDS, BOHLEN 1996, KASPRZAK 1998, KOSTECKA 2000) or in
households (APPELHOF 1982, 1993, KOSTECKA 1994, YGADŁO 2002, JDRCZAK
2007, ROSIK-DULEWSKA 2007). In some countries the whole system solutions for
the organic waste management with the use of vermiculture are created
(FREDERICKSON, HOWELL 2003, BLOUIN et al. 2006, GARCIA-ORTEGA, OLIVARESGONZALES 2006, FREDERICKSON et al. 2007).
In the process of vermicomposting other species of earthworms are potentially
suitable; in a temperate climate: Eisenia fetida andrei, Dendrobena veneta,
Dendrobena rubida and Lumbricus rubellus, in a tropical climate: Eudrilus eugeniae,
Perionyx excavatus and Pheretima elongata (EDWARDS 1988, 1998, DOMINGUEZ
2004, SINGH et al. 2004).
Vermiculture is relatively new biotechnology (EDWARDS, BOHLEN 1996) which
takes place in controlled conditions and comprises of the breeding of concentrated
populations of earthworms in various organic waste. In order for a process to be
called vermiculture there should be over 100 specimens of earthworms per 1 dm3
(LAC 1991, KOSTECKA 2000, 2004A, ZHENJUN 2003, GARG et al. 2006). Thanks to
its life functions such a concentrated population very quickly and successfully
transforms various organic wastes into a fertilizer (vermicompost) of excellent
quality (KALEMBASA 1998, ZABŁOCKI, KIEPAS-KOKOT 1998, KOSTECKA 1999A,
SZCZECH, SMOLISKA 2001, ARANCON et al. 2003, 2004, EDWARDS et al. 2004,
HURY 2008, ALI et al. 2007, ZALLER 2006, 2007, GUTIERREZ-MICELI et al. 2007). In
the situation where soil in Europe is low in humus and where organic waste is a
threat to the environment, the vermiculture biotechnology makes it possible to
neutralize biowaste in a pro-environmental way.
One of the requirements for a vermiculture technique is feeding earthworms with
thin layers, providing the next ones only after earthworms have finished eating the
previous layers (GADDIE, DOUGLAS 1977). For the proper vermicomposting of
organic waste with the presence of Eisenia fetida, pH of soil in breeding beds needs
to be kept within 6.7-7.5 and, as far as possible, the temperature should be regulated
(the ideal soil temperature for the life functions of E. fetida is 12-28o C). Humidity
158
(about 70%) as well as aeration is also very important. Currently the main directions
of the use of vermiculture are considered to be:
• Neutralizing segregated organic waste; through the production of vermicompost
which results in the possibility of supplementing the lack of nutrients in plants
and making microorganisms in soil more active,
• The production of earthworms body biomass which is rich in protein (about 5871% of dry mass); biomass can be used as a food supplement for fish, poultry,
pigs and other animal e.g. zoo animals,
• Gaining additional populations of earthworms whose introduction to soil will
improve the process of reclamation,
• Using enzymes and other substances included in lymph in cosmetic, wine, beer,
textile and medical industry (ZHENJUN 2003; DOMINGUES 2004; DOMINGUES,
EDWARDS 2004, EDWARDS et al. 2004).
In the process of vermicomposting it is possible to neutralize various organic
wastes e.g. sewage sludge, post-production wastes of agricultural processing
industry, post-harvest residues, green wastes, cotton wastes, wastes generated during
coffee production, kitchen wastes, wastes from supermarkets and restaurants, wastes
from slaughterhouse, bones and feathers from poultry slaughterhouse, excrement
(poultry, pig, cattle, sheep and horse), excrement of fur-bearing animals (foxes,
minks and rabbits), post-champignon wastes, wastes from beer and paper factories
(EDWARDS 1988, BUTT 1993; OROZCO et al. 1996, KASPRZAK 1998, KOSTECKA
1999A, 2000, 2004B, BENITEZ et al. 2002, ZAJC 2002, FREDERICKSON, HOWELL
2003, DOMINGUEZ 2004, DOMINGUEZ, EDWARDS 2004, LOH et al. 2005,
GAJALAKSHMI et al. 2005, GUPTA et al. 2005, NOGALES et al. 2005, GARG et al.
2006, HUERTA et al. 2006, AIRA, DOMINGUEZ 2007, SUTHAR 2007, BRANDON et al.
2008).
It must be stressed that this pro-environmental biotechnology is only common to
a lesser extent. Its rare occurrence is usually associated with the fact that people
breed earthworms for their own use and that earthworms are used to escalate the
process of organic matter breakdown only on a small scale (ROCISZEWSKA et al.
1998), whereas in other countries such as Germany, France (YGADŁO 2002), Spain,
Canada, Sweden and the USA the process takes place far more often and is far more
common (GADDIE, DOUGLAS 1977, KOSTECKA 2004A).
As it has been mentioned before, the breeding of earthworms can take place on
various scales: small and „handy” or large , e.g. at sewage treatment plants.
KOSTECKA (2000) conducted a positive trial of vermicomposting of sewage sludge
at a sewage treatment plant in Brzesko, Poland. After the preliminary trial in a
laboratory and the preparation of a station for vermiculture at the plant, earthworms
were introduced onto 6-month sludge of stabilized parameters which was also
modified with sawdust. After 3 months earthworms transformed the sludge into a
fertilizer of tubercular structure.
We can vermicompost organic material in solid beds e.g. made of concrete (an
optimum size: 2 m wide, 10 m long and 0.5 m high, which is equivalent to 10 beds
according to the US norms) (GADDIE, DOUGLAS 1977). On the other hand,
earthworms can also transform organic waste in beds made from waste materials
(e.g. wood, brick, rubber transmission belts) (KOSTECKA 2000). We can also
159
organize systems of a high technical level (BOUCHE 1987, RIGGLE, HOLMES 1994,
BLOUIN et al. 2006, GARCIA-ORTEGA, OLIVARES-GONZALES 2006, FREDERICKSON
et al. 2007).
Vermiculture and agricultural households
Providing permanent access to organic fertilizers and feed and, at the same time,
reducing their purchase cost is known to be an important issue in the agricultural
sector. Which is why neutralizing animal excrement, post-harvest residues as well as
home organic wastes through vermiculture can be of paramount significance (Figure
1). The drawing shows that in vermiculture we can use any segregated organic
wastes (household wastes, wastes generated by herbivorous and omnivorous animals
as well as any post-harvest residues). Vermicompost (which can used e.g. for
vegetables, orchards or field cultivation) and the protein of an earthworm body wall
(which can be used as a feed supplement for omnivorous animals and fish) are the
products of vermiculture.
Presently, in some parts of Poland, there is a growing interest in vermiculture but
mainly on a small scale in order to produce vermicompost. Vermicompost derived
from cattle manure has good fertilizing qualities. Experiments on the use of
vermicompost have been conducted for many years now (JARECKI, MAKOWSKI
1992, MURAWSKA et al. 1992, KOSTECKA, KOŁODZIEJ 1995, SŁAWISKI, SONGIN
2001, HURRY 2008).
Although the harvest of potatoes grown on vermicompost is smaller (JARECKI,
MAKOWSKI 1992) or the same (SADOWSKI, NOWAK 1990) in comparison to the
harvest of potatoes grown on manure, the conducted experiments also show a
positive influence of vermicompost on harvest. It transpires that the proportion of
potato tubers that can be consumed is higher (KOSTECKA et al. 1996, SADOWSKI,
NOWAK 1990) and the health of plants is considerably better. In the study conducted
by KOSTECKA and co-authors (1996), tubers grown on vermicompost were
sporadically affected by Phytohthora infestans during harvest as well as after 7
months of storage. It is confirmed by SZCZECH i BRZESKI (1994) who consider that
vermicompost functions as a plant protection agent.
The content of nitrate and heavy metals has been determined in vegetables
grown on vermicompost (KOSTECKA, BŁAEJ 2000). They have higher nutritional
value (the content of nitrate, lead and cadmium was lower) in comparison to
vegetables fertilized with minerals. Favourable properties of vermicompost as an
organic fertilizer have been scientifically proven which is also confirmed by
allotment owners who notice that plants are healthier and crops yields are higher.
Vermicompost has a positive influence on soil fertility (EDWARDS et al. 2004),
however, some authors are of the opinion that soil fertility is stimulated more by the
growth of a microorganism population rather than vermicompost (DOMINGUEZ
2004, TOGNETTI et al. 2005). Other authors also proved that vegetables (potatoes,
cabbage, lettuce, tomatoes) and e.g. strawberries grown of that organic fertilizer (of
examined content, produced from cattle manure) grew faster (ATIEYH et al. 2002,
ARANCON et al. 2003, 2004; HASHEMIMAJD et al. 2004, GUTIÉRREZ-MICELI et al.
2007) and were healthier (SZCZECH, SMOLISKA 2001, SZCZECH et al. 2002;
BŁAEJ, KOSTECKA 1998) than those fertilised with minerals or compost.
160
Vegetables (potatoes, tomatoes, cucumbers) fertilized with vermicompost absorbed
less amount of nitrate and heavy metals in comparison to those fertilised with
minerals (KOSTECKA, BŁAEJ 2000). In the study conducted by EDWARD et al.
(2004) it was shown that vermicompost significantly slowed down the growth of
pathogenic fungi such as Pythium, Rhizoctonia i Verticillium and vegetables
contained smaller amount of heavy metals. Agricultural use of vermicompost
escalated the growth of a root and the absorption of nutrient (PADMAVATHAMMA et
al. 2008).
Source: author’s own work
Dark arrows show organic wastes which are feed for earthworms, light arrows identify the
possibilities of using the products of vermiculture: vermicompost (right) and earthworm
biomass (left)
Fig. 1. Vermiculture in an agricultural household
Taking into consideration the multifunctional development of rural areas and the
need to create additional sources of income for country dwellers, the importance of
another project – earthworm mass – needs to be emphasised. Not only can
earthworms be a feed supplement but also mature specimens can be sold e.g. to
fishermen or as a pedigree material for other vermicultures.
As mentioned earlier, the existence of vermiculture in agricultural households
creates the possibility to use the protein of an earthworm body. For many years,
several publications have shown (MC INROY 1971, SABINE 1983) that earthworm
biomass is an attractive feed due to its high levels of aminoacids such as lysine,
methionine, cystine, tryptophan and threonine. In periods of long-lasting drought,
during winter and when it becomes cold, the possibilities of finding biomass by
birds bred in yards are limited. In such cases, earthworm biomass derived from
vermiculture can be a cheap source of valuable protein reserves. Many authors have
conducted the analysis of the content, nutritional value and the vitamins of
earthworms which demonstrated that earthworms are an attractive feed for fish,
poultry, pigs and zoo animals (ZHENJUN 2003; VIEIRA et al. 2004; SOGBESAN et al.
2007A i B).
161
In an individual study (KOSTECKA, DEJNEKA 1998), it was proved that poultry
using a free yard (French white ducks and hens of a general use were fed) had an
interest in the biomass of earthworm E. fetida. According to KORELESKI et al.
(1994), we can introduce earthworms into mixed fertilizer in the form of powder or
granulate. The need of constant rejuvenation of an earthworm population favours the
idea of the removal of earthworm biomass from breeding stations on a regular basis.
It results in the possibility of starting new stations or using earthworms as feed.
Earthworms used in household vermiculture need to be protected from scratching
birds (e.g. with plastic net) to prevent parasitic diseases of poultry from spreading.
Feeding scratching birds with biomass must take place in controlled conditions.
Vermiculture and individual households and public institutions
Food leftovers are generated not only in the kitchen – we produce them in great
amounts at schools, universities, offices, canteens, street markets. It has been shown
that vermicompost derived from home and office wastes is very rich in nutrients,
however, its salinity may be high (KIEPAS-KOKOT, SZCZECH 1998, KOSTECKA
2000). Vermiculture which takes place in small boxes (at home, school, hospitals,
canteens, work places) makes recycling of organic waste of high quality possible
(selected organic wastes) in places where they are produced. In order to draw
attention towards pro-environmental actions of those who decide to start
vermicomposting, such a solution has been called “an ecological earthworm box”
(KOSTECKA 1999B, 2000).
„An ecological earthworm box” (at home, school, work place etc.) makes it
possible for anyone to contribute to the sustainable waste management; it reduces
the ecological footprint (BEST FOOT…) which is associated with transport of waste
to landfill sites and it prevents negative consequences of biowaste deposits at landfill
sites as well as illegal dump sites.
We need to follow the following rules in order for ecological boxes to function
properly:
• A feed layer should be no thicker than 10 cm every time a new one is added. We
should not include too much meat waste, we need to add acid waste very
carefully bearing in mind that earthworms in small boxes are not able to retreat
to a remote and safe place. The practice showed that feed should not cover the
whole surface of a breeding box. The observation of earthworm behaviour and
their gathering in new waste can indicate the acceptance of new conditions,
• We need to thin an earthworm population on regular basis by taking away of about
25% of specimens several times during a vegetation season,
• The humidity in boxes should be kept on a permanent level of 70% and other soil
parameters should also be taken considered ,
• As a box is being filled up, we need to gradually remove the resulting
vermicompost (KOSTECKA 1994, 2000).
Oxygen is also an important factor for the proper functioning of the earthworm
biology. Soil bed should be loose and porous. An important aspect of the functioning of
ecological boxes can be the presence of additional fauna e.g. the population of
162
Enchytraeidae (white worms). Their large populations can slow down the growth of E.
fetida because of the excretion of substances that are toxic to earthworms. That is why,
despite large amounts of waste, earthworms can starve and consequently die when white
worms are present in boxes (MAKULEC 1996). Earthworms prefer feed without the
excretions of white worms. The number of earthworms was three times larger in waste
without Enchytraeidae’s excretion (KOSTECKA, ZABOROWSKA-SZARPAK 2001).
A serious problem of ecological boxes with a small volume of soil bed is the
concentration of a breeding population which is too high. When this happens specimens
have a negative effect on one another (the worsening of the conditions of earthworms
and the lack of copulation) (KOSTECKA 2000). In such conditions the life strategy of
earthworms means moving energy distribution from procreation towards growth (AIRA
et al. 2007). Too high a concentration can lead to the predominance of male
specimens which do not produce ova (anisopary effect) (MEYER, BOUWMAN 1997).
Consequently such a process leads to the slowing down of the growth of a population as
earthworms do not reproduce. Such an unfavorable process can be slowed down by
reducing the amount of waste containing cellulose in the ratio 1:1/2 (KOSTECKA 2000).
Vermiculture and businesses
According to the principle of constant development, the knowledge of the
environmental impact assessment (EIA) needs be considered as one of the most
fundamental issues that should be known to owners and managers of large as well as
small businesses. EIA refers to the localization of businesses, the realization of their
aims, the exploitation of equipment and installation together with the functioning of
the product or service provided. The pro-environmental management of businesses
leads to the assessment of the impact that production/service processes have on the
environment and then taking action in order to gradually reduce negative effects
(e.g. reducing the volume of produced waste; putting less pressure on the
environment owing to smaller electric and thermal energy consumption, using less
water and producing less sewage; which reduce the emission of hazardous
substances to the atmosphere and surface water).
The redirection of businesses towards pro-environmentalism often requires only
small investment. Moreover, as far as waste management is concerned, businesses
can reduce their ecological footprint by using vermiculture (Table 2).
Spreading the information about so many potential advantages of an ecological
earthworm box may increase an interest in possessing one. The questionnaire
conducted among random workers of Huta Stalowa Wola shows (KOSTECKA et al.
2007) that 63% of those questioned would agree to have an ecological box in their
surroundings. Their preferences, as far as the location of a box is concerned, are
presented in the Figure 2.
Those willing to have an ecological box would probably have to face a serious
problem – the difficulties in finding a pedigree population. Nowadays the number
of earthworm breeders has drastically decreased. The last inventory of vermiculture
was conducted in the period of 1995-1997 and showed the disappearance of the vast
majority of 209 farms (ROCISZEWSKA et al. 1998). Nowadays we know of only
a few breeders with sizeable vermicultures. It seems, however, that they are able to
163
expand their farms if interest in purchasing pedigree populations increases.
Table 2
Examples of actions that lead to the reduction of anthropogenic impact of businesses within
the scope of waste management (including organic waste management) (based on
KOSTECKA, KOSTECKI 2006, changed)
Sector
WASTE
management
Actions which lead to the reduction of anthropogenic
impact and to financial savings
-defining areas where it is possible to limit the production of waste, analyzing
its amount and content in order to determine purchase policy (e.g. cleaning,
office and food products),
-the purchase of products (company supplies) in big or returnable packaging,
-conducting the analysis of costs of packaging,
-preparing a waste (materials) recycling programme,
-reducing packaging of toilet articles, introducing soap dispensers,
-reducing the amount of informative leaflets for clients and printing on
recycled paper,
-composting of organic wastes in a place where they are generated in „an
ecological earthworm box”a (reducing the costs of having waste collected,
producing vermicompost),
-using electronic mail (the reduction of paper use).
a - based on Kostecka, 1999a, 2003
% of interviewed
a)
100
80
60
40
20
0
in a room
women
% of interviewed
b)
in a corridor
men
outside
the building
100
80
60
40
20
0
1
women
2
3
4
5
men
a) at a work place:
b) in household conditions: (1) at home (2) on a balcony (3) in a basement (4) at an allotment (5) in the garden
Source: KOSTECKA et al. 2007
Fig. 2. The preferences of the location of boxes with earthworms
164
Summary
Vermicomposting is one of the solutions to the problem of neutralizing organic
waste. It can reduce the volume of waste at landfill sites and its negative impact on the
environment due to the possibility of producing biowaste in places where waste is
generated. Vermicomposting can also promote the idea of getting closer to nature. It also
favours the sustainable waste management which results in economic savings on
a home, local, regional, national continental and even World scale.
Vermiculture can be popularised in rural areas. It offers a successful transformation
of household organic waste into fertilizer. Using vermicompost for vegetable crops
allows the achievement of high biochemical value. It is also possible to use earthworms
for biomass.
The topic of vermiculture should be included in education within the scope of ideas
and problems concerning environmental engineering, the formation of the environment,
environmental protection or sanitary engineering. It can also be useful in other courses.
As the United Nations Decade of Education for Sustainable Development (20052014) has been announced, scientists should promote access to the aspects discussed
above among a large number of students within the scope of a general subject or the
humanities.
It also needs to be highlighted that due to the growing interest in proenvironmental actions and the popularization of the results of studies on
vermiculture, taking further actions e.g. the economics of the vermiculture
phenomenon, which is understood in broad terms, is justified and even required.
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Joanna Kostecka
The Chair of Natural Theories of Agriculture and Environmental Education
University of Rzeszow
ul. Cwiklinskiej 2, 35-601 Rzeszów, POLAND
170
CHAPTER XIII
Halina Dbkowska-Naskrt, Agata Bartkowiak, Jacek Długosz,
Szymon Róaski
THE QUALITY OF SOIL TARE FROM THE SUGAR
PLANT WITH REGARD TO ITS UTILIZATION FOR SOIL
FERTILIZATION
Introduction
One of the major waste from sugar industry is soil tare from the beet cleaning
(soil tare) (MIZERSKA 2007). Soil tare is on the list of wastes, which the sugar plant
can utilize 2006 (MINISTER OF ENVIRONMENT ORDER 2006). Soil tare has the
highest share this group of waste. Under waste utilization regulation 27.04.2001
(LOW OF WASTE 2001) waste, the formation of which is impossible to prevent, and
for which there exist technologically and economically feasible grounds to ensure
proper recycling, under pertinent regulations of environmental protection, should be
recycled in the first place.
Soil loss problem due to root crop harvesting is significant if we consider
impoverishment in nutrients and organic matter. When harvesting root crops such as
sugar beet, potato, carrot or leak, significant amount of soil retained by the root
furrow is taken out from the field. Data from the intensive sugarbeet production
show that the mass of wet soil sticking to the root (soil tare) may stand for up to
11% of the mass of the raw material that is delivered during the campaign (ORUC,
GUNGOR 2008).
Soil material is retained on the storage roots and its amount is related to the
shape of the root, the amount of lateral roots, depth of rooting, the composition of
soil and the amount of water in it as well from the technique of harvesting. Soil mass
taken out from the field depends also on the adhesive properties of soil and its water
capacity, increasing with the increase of clay fraction and water contents in soil
during the harvesting (LARYMERS, STRÄTZ 2003).
The amount of soil material brought with the beets to the sugar plant is related to
morphology of root and its size. Large beet roots with a smaller or no furrow contain
less soil. Moreover, the depth of rooting (fig. 1) that depends on the variety of beet
and the technology of harvesting also influence the soil tare (VERMIEULEN, KOOLEN
2002).
Particularly large amounts of nutrients are taken out from soils rich in organic
matter, with high water capacity. It was reported that a significant amount of
nutrients is lost from soil; for example annual loss of phosphorus is up to 3.0 kg P
171
ha-1 and nitrogen 30 kg N ha-1 during the sugarbeet production. Li and co-workers
(LI et al. 2006) estimated that the largest losses of soil are reported during the potato
and sugarbeet grow.
Fig. 1. Different depths of beet rooting: shallow (1-2) medium (3-5), deep (6-7).
RUYSSCHAERT et al. (2008) reported that soil losses during the root crop
harvesting is comparable to soil degradation due to water or wind erosion. Soil
losses at sugarbeet harvesting ranges between 1.2 to 1.9 tha–1 yr–1, and 0.2 to 0.3
tha–1 yr–1at potato harvesting. The process of soil losse due to sugarbeet production
in Bavaria (Germany) ranges between 4.5 and 7 tha–1yr-1 (MAIER, SCHWERTMANN
1981).
Taking into account long term cultivation of root crops it is necessary to regard
the decrease of soil profile depth as a result of soil adherence to the root surface.
A new parameter was used for characterization of soil erosion processes when
harvesting such crops as sugarbeet (Beta vulgaris L.) SLCH (Soil Losses due to
Crop Harvesting) (RUYSSCHAERT et al. 2007) or SLRH (Soil Loss due to Root crop
Harvesting) (POESEN et al. 2001).
Long term study of SLCH for sugarbeet in Belgium showed that in the years
1968 – 1996 the annual values of this parameter equal 18.7 – 20.4 tha–1 in rainy
years and 4.2 - 4.6 tha–1 in dry years.
Calculated annual mean value for SLCH was 5.0 tha –1 (POESEN et al. 2001).
The above data indicate that negative effects of the process is related to soil loss and
its degradation and also to the increased costs of soil transport with the crop to the
sugar plant.
Transportation of soil with the crop should be reduced as it causes
environmental problems and the increase of costs of the final product (KOCH 1996).
In Poland the problem of soil tare utilization is not well recognized. It is not
only of concern for working sugar factories but also those which are closed and have
left waste and byproducts to be utilized. It is very important particularly in the light
172
of environmental protection programs in regions and provinces where over 40 % of
the total industrial waste come from sugar industry (MANAGEMENT OF WASTE
DISPOSAL 2003).
Conditions of the study
One of the working sugar plants is Glinojeck S.A. located in the Ciechanów
district, Mazowieckie province. Daily, the plant uses 12000 tons of beets. In the year
2007, 12 mln tons of sugarbeet were processed and corresponding amounts of soil
tare plant was deposited in the vicinity of the plant.
The study of soil tare partly mixed with lime in the vicinity of the sugar plant in
Glinojeck was undertaken. Soil tare was sampled from near the Glinojeck plant.
Waste material has been collected for the last 20 years and consists of soil that was
washed out from the beets and defecation lime (another by-product of sugar
production).
Total area of the pile was 0.5 ha, with the irregular shape 132 x 81 m (Fig. 2).
Prior to sampling, preliminary drilling was made and the studied area was divided
into 14 plots. From each plot a collective sample consisting of 10 –15 sampling was
taken. Material sampled from the depth 20 – 60 cm was analyzed.
1
2
3
4
5
6
7
9
8
10
11
13
12
14
Fig. 2. Localization of samples taken for the analysis.
Soil material in the laboratory was dried in the room temperature; there was no
plant residuum in it. There were no coarser fragments of stones except for several
clums of lime stone.
Samples were sieved through a 1 mm sieve and the following analyses were
performed:
1. pH in H2O was determined potentiometrically on pH-meter Radiometer PHM.
2. Organic carbon was determined according to Tiurin`s (LITYNSKI 1976).
3. Content of CaCO3 according to Scheibler (LITYNSKI 1976).
173
4. Texture was determined according to Boycouse – Cassegrande method in
Prószyski modification.
5. Total phosphorus using molibdeniane method.
6. Total contents of macroelements (K, Mg, Na) and trace elements Zn, Cu, Ni, Pb,
Cr and Cd) after mineralisation in concentrated acids (HF and HClO4) (CROCK
and SEVERSON 1980) was determined using AAS technique on Philips PU
9100X spectrometer.
7. Available forms of P and K according to Egner – Riehm method.
8. Available Mg according to Schatchabel`s.
9. Contents of S-SO4 according to BARDSLAY – LANCASTER (1960).
10. Total content of Hg was determined using AMA 256 spectrometer.
11. Contents of available Zn, Cu, Ni, Pb and Cd according to LINDSAY and
NORVELL (1978) after the DTPA extraction, on AAS spectrometer.
Soil tare composition
Analysis of soil tare samples (Table 1) showed that their pH was neutral or
slightly alkaline with pH in H2O in the range between 7.05 – 7.49 and the mean
value at 7.26.
Table 1
Physico-chemical properties of the studied material
Soil No
pH in H2O
1
2
3
4
5
6
7
8
9
10
11
12
13
14
Mean
Range
7.05
7.17
7.13
7.20
7.25
7.19
7.39
7.21
7.28
7.31
7.24
7.49
7.40
7.37
7.26
7.05-7.49
C org.
[g kg-1]
5.9
10.1
9.6
9.8
8.3
13.0
10.9
14.3
10.9
9.2
12.4
12.7
10.9
9.7
10.6
5.9-14.3
ø<0,002mm
[%]
17.77
1
24.73
5
21.12
2
55.48
2
21.23
10
20.69
4
24.88
4
27.97
4
22.56
4
17.43
2
29.94
7
24.13
5
17.68
5
29.23
5
25.35
4
17.43-55.48
1-10
CaCO3
Texture*
LS
SL
SL
SL
SL
SL
SL
SL
SiL
SL
SL
SL
SL
SL
* - LS - loamy sand, SL - sandy loam, SiL - silt loam (USDA)
Total organic carbon contents ranged from 5.9 to 14.3 gkg–1 (with the mean
value 10.6 gkg-1). Such amounts are characteristic for soils from sugarbeet fields
usually rich in organic matter.
174
Content of calcium carbonate differentiated from 17.43 to 55.48 % with the
mean value 25.35 %. High amounts of CaCO3 in the analyzed soil material came
from another byproduct (defecation lime) in sugar technology; at present changed
technology allows to separate lime from the soil material.
Texture of the studied material is mainly loamy, with the clay fraction (ø < 0.002
mm) content in the range 1 – 10 % (Table 1). Fine size clay particles dominate in
tare soil from the sugar plant (MIZERSKA 2007). This fraction presents highest
sorption capacity for nutrients.
Adhering soils removed from the surface layer of beet fields contain appreciable
amounts of essential plant nutrients such as total phosphorus (except samples 1 and
10) and potassium, calcium and magnesium: 0.73 gkg–1, P; 2.4 gkg–1,
80.67 gkg–1 Ca and 2.3 gkg–1 Mg (Table 2).
The content of available forms of potassium was high and very high (PN-R04022 1996) and ranged between 158.0 and 282.0 mg kg –1 of soil. Phosphorous and
magnesium contents were high (PN-R-04023 1996; PN-R-04020 1994) and were in
the range from 153.0 to 192.0 mgkg –1 and 215.0 – 515.0 mgkg–1 respectively.
High contents of macroelements in soil tare can be a source of essential plant
nutrients. Moreover, the addition of lime is a source of calcium in soil tare and such
enrichment improves physical properties of soil and soil pH (MIZERSKA 2007).
Table 2
Total contents of macroelements
Soil No
1
2
3
4
5
6
7
8
9
10
11
12
13
14
Mean
Range
P
K
0.07
0.75
0.85
0.72
0.36
0.75
0.75
0.88
0.65
0.20
1.60
1.11
0.75
0.75
0.73
0.07-1.60
5.4
2.6
2.6
2.7
1.7
2.3
2.0
1.9
2.2
1.6
1.7
2.7
2.2
2.0
2.4
1.6-2.7
Mg
(gkg-1)
1.7
2.1
2.0
3.9
1.9
2.7
2.1
2.5
2.1
1.9
2.2
2.6
2.1
2.4
2.3
1.7-3.9
Ca
Na
53.7
83.2
86.7
90.4
82.6
84.7
91.6
85.7
87.9
73.2
59.4
89.2
79.2
81.9
80.67
53.7-91.6
0.3
0.3
0.3
0.3
0.4
0.3
0.3
0.3
0.3
0.2
0.4
0.4
0.4
0.4
0.3
0.2-0.4
The content of S-SO4 ranged from 76.6 to 157.0 mgkg–1 (mean 117.0 mgkg –1)
which is characteristic of the top soil material rich in clay fraction. The observed
sulphur contents in S-SO4 form are high but typical for the natural level of this
element (Table 3) – TERELAK et al 1998.
175
Table 3
Content of available forms of selected macroelements
Soil No
1
2
3
4
5
6
7
8
9
10
11
12
13
14
Mean
Range
P
K
192.0
188.0
175.0
175.0
153.0
179.0
183.0
181.0
177.0
175.0
183.0
172.0
175.0
172.0
177.0
153.0-192.0
158.0
203.0
141.0
166.0
170.0
224.0
208.0
212.0
282.0
212.0
212.0
282.0
299.0
232.0
214.0
158.0-282.0
Mg
S-SO4
215.0
385.0
315.0
275.0
305.0
335.0
260.0
515.0
315.0
295.0
340.0
315.0
305.0
320.0
321.0
215.0-515.0
157.0
108.5
127.6
110.6
76.6
157.2
106.8
138.0
98.2
113.6
102.1
124.6
110.6
106.7
117.0
76.6-157.0
(mgkg-1)
Total contents of microelements in the soil material were typical for soils having
loamy and silty texture. Samples contained relatively high amounts of
phytoavailable zinc and copper (Table 4). The contents of nickel, lead, cadmium and
mercury were on the levels of natural and fulfilled all the requirements needed for its
agricultural application (KABATA-PENDIAS, PIOTROWSKA 1987). Similarly, the
properties of phytoavailable forms of these metals in analyzed soil tare were low
(Table 4).
The contents of analyzed metals were in the acceptable range, and stemmed
from the variability of composition of soil brought from the beet fields, and the
amounts of lime added during the process of sugar production.
176
Table 4
Total contents and available forms of selected microelements
No
1
2
3
4
5
6
7
8
9
10
11
12
13
14
Mean
Zn*
22.7
22.22
28.62
19.22
24.6
27.7
31.87
23.27
22.65
28.02
41.02
65.77
64.85
58.85
34.38
19.22Range
65.77
Zn** Cu* Cu** Ni*
1.02
1.19
1.26
1.37
1.23
1.59
1.47
1.37
1.63
1.72
1.74
1.52
2.26
2.24
1.54
1.022.24
3.3
4.01
3.51
2.96
3.47
3.17
2.9
3.06
3.02
2.69
3.29
3.15
2.65
2.85
3.15
2.694.01
0.4 9.62
0.42 9.86
0.46 9.37
0.44 9.3
0.53 9.42
0.55 9.46
0.71 9.19
0.73 9.0
0.58 9.1
0.47 9.2
0.78 8.94
0.74 9.04
0.64 9.05
0.43 8.71
0.56 9.23
0.4- 8.710.78 9.86
Ni** Pb*
[mgkg-1]
0.73 18.66
0.81 10.17
0.74 12.71
0.8 10.75
0.53 10.3
0.62 15.45
0.75 12.46
0.7 7.95
0.65 11.32
0.73 13.02
0.7 19.1
0.74 10.56
0.8 13.29
0.82 11.25
0.72 12.64
0.53- 7.950.82 19.1
Pb**
Cd*
Cd**
Cr*
1.96 0.12
0.04 8.45
0.3
0.15
0.06 10.24
0.54 0.25
0.08 11.36
0.45 0.26 <0.02 11.01
0.26 0.44
0.04 10.42
0.9
0.69
0.08 10.95
0.7
0.55
0.05 11.27
0.27 0.67
0.12 10.92
0.4
0.72
0.12 10.64
0.6 <0.02 <0.02 12.15
0.5 <0.02 <0.02 13.72
0.46 0.17
0.07 14.55
0.77 0.09
0.07 14.66
0.28 0.15
0.11 13.85
0.6
0.31
0.07 11.73
0.26- <0.02- <0.02- 8.451.96 0.69
0.12 14.66
Hg*
0.012
0.012
0.019
0.015
0.012
0.02
0.015
0.02
0.015
0.013
0.016
0.016
0.016
0.015
0.015
0.0120.02
* - total content, ** - content of DTPA extractable forms
Summary
The results of the chemical and physico-chemical analysis indicate that the
waste from sugar plant (soil tare) is a valuable material for the fertilisation of sandy
soils, with acid pH values, and poor in nutrients for plant.
Agricultural application of the soil tare is not hazardous for the environment as
regards the contents of heavy metals such as Pb, Hg, and Cd.
Thus, soil tare, - the waste which comes from cleaning and washing of sugar
beets is proper for the enrichment of fields in nutrients, also as the additive of other
waste such as composted sewage sludge.
Having taken into consideration the fact that bulb and root plants (beet) left
relatively low amounts of plant residue in the soil, and even then the residue
undergoes swift mineralisation, it would be beneficial to apply soil tare on the beet
fields, from which the most valuable components were removed together with the
crop. Such a supplementation would decrease the value of the SLCH indicator, and
the losses linked with it, and comporable to the losses during erosion process.
Soil tare is also recommended for landscaping during such investments
as construction of highways, sodding of artificial embankments, slopes or pits.
177
References
CROCK J.G., SEVERSON R. 1987. Four reference soil and rock samples for measuring
element availability in the western energy regions. Geochemical Survey Circular.
KABATA-PENDIAS A., PIOTROWSKA M. 1987. Pierwiastki Ğladowe jako kryterium rolniczej
przydatnoĞci odpadów. IUNG, Puławy, Seria P33: 46.
KOCH H.J. 1996. Possibilities and limits for reducing soil tare of sugarbeet through tillage
density, N-fertlizer supply variety and cleaning. Proc. 59 th Il RB Congress: 483 – 497.
LARYMERS P., STRÄTZ J. 2003. Progress in soil tare separation in sugarbeet harvest.
Journal of Plant Nutrition and Soil Science. 166, 1: 126 – 127.
LI Y., RUYSSCHAERT G., POESEN J., ZHANG Q.W. 2006. Soil loss due to potato and
sugarbeet harvesting. Earth Surface Processes and Landforms. J. Wiley Sons. 31, 8:
1003 – 1016.
LINDSAY W.L., NORVELL W.A. 1978. Development of a DTPA soil test for zinc, iron,
manganese, copper. Soil Sci. Soc. Am. J., 42: 421-428.
LITYSKI T. 1976. Analiza chemiczno-rolnicza. Wyd. PWN Warszawa.
LOW OF WASTE on 27 april 2001. [Journal of Low No 62, 628 as amended].
MAIER J., SCHWERTMANN U. 1981. Das Ausmass des Badenabstrags in einer Lösslandschaft
Niederbayerns. Bayerisches Land wirtschaftiches Jarbuch, 58 (2): pp 194.
MANAGEMENT OF WASTE DISPOSAL kujawsko-pomorskie province 2003. Typescript, Toru.
MINISTER OF ENVIRONMENT ORDER on 21 april 2006 [Journal of Low No 75, 527]
MIZERSKA D. 2007. Gleba i inne odpady uzyskane podczas procesu oczyszczania buraków –
metody zagospodarowania, przepisy prawne. Gazeta Cukrownicza 10: 330-332.
ORUC N., GÜNGÖR H. 2003. A study on the soil tare of sugarbeet in Eskisehir.
PN-R-04020. 1994. Chemico-agricultural analysis. The determination of available
magnesium in mineral soils. Polish Committee of Normalization.
potassium in mineral soils. Polish Committee of Normalization.
phosphorus in mineral soils. Polish Committee of Normalization.
POESENJ., VERSTRAETEN G., SOENSEN R., SEYNAEVE L. 2001. Soil losses due to harvesting
of chicory roots and sugar beet: an underrated geomorphic process? Catena, 43: 35 – 47.
RUYSSCHAERT G., POESEN J., NOTEBAERT B., VERSTRAETEN G., GOVERS G. 2008. Spatial
and long term variability of soil loss due to crop harvesting and the importance relative to
water erosion: a case study from Belgium. Agriculture, Ecosystems and Environment.
126: 217 – 228.
RUYSSCHAERT G., POESEN J., WANTERS A., GOVERS G., VERSTRAETEN G. 2007. Factors
controlling soil loss during sugarbeet harvesting at the field plot scale in Belgium.
European Journal of Soil Science. 58, 6: 1400 – 1409.
TERELAK H., MOTOWICKA-TERELAK T., PASTERNACKI J., WILKOS S. 1998. ZawartoĞü siarki
w glebach mineralnych Polski. Pam. Puł., supl.891, 1-59.
VERMEULEN G.D., KOOLEN A.J. 2002. Soil dynamics of the origination of soil tare during
sugarbeet lifting. Soil and Tillage Research. 65, 2: 169 – 184.
Halina Dąbkowska-NaskrĊt, Agata Bartkowiak, Jacek Długosz,
Szymon RóĪaĔski
Department of Soil Science and Soil Protection
University of Technology and Life Sciences
Bernardyska 6, 85-029 Bydgoszcz, POLAND
178
179
180
The Faculty of Environmental Management and Agriculture of University of Warmia and Mazury
in Olsztyn unites tradition and modernity. The result of its 60 year history is the inheritance of the spirit
of numerous scholars of polish agriculture sciences. From the very beginning, it has educated
specialists, who irrespective of the political and the economic situation, have tried to develop modern
farming in the Warmia and Mazury region and outside its borders. The multidisciplinary education
made the offered programs of studies and laid the foundation for new faculties of technology, natural
sciences and economics based on the growing scientific resources. What should also be stressed is the
contribution of the Faculty staff to the preparatory work concerning organizational and curriculum
planning issues of the University. Today the Faculty, which is at the top national rankings, offers
candidates the widest range of areas of studies and specializations. It enjoys the great interest of youth
from the north-east of Poland. There are always more candidates than the University can admit. The
Faculty offers bachelor and master’s studies, both regular and extramural.
The Faculty of Environmental Management and Agriculture is a University unit which promotes
people of both collective and individual success. The exceptional staff of academic teachers guarantees
a standard of education which is comparable, and sometimes even higher than that offered at similar
academic institutions. Those admitted every year in the Faculty of Environmental Management and
Agriculture (c. 800 students).
There 252 persons of faculty and staff:
• 32 full professors,
• 31 associate professors,
• 98 doctors (assistant professors),
• 4 assistants,
• 4 lecturers,
• 87 supporting persons (administrations and technicians) and others.
The main topics of research work:
• biological and agrotechnical determinations of crops productivity and landscape management,
• methods and techniques of crop nutrition and crop protection as regards productions of safe foods
and the needs of environmental protection,
• methods and techniques of land reclamation,
• methods of growing and technological value of yield of vegetable and fruit crops,
• methods of sewage and waste utilization of estimation of alternative crops and development of seed
production of alternative crops,
• creation of new genotypes of crop plants,
• multifunctional development of rural areas.
The main fields of education:
• Agronomy,
• Engineering and systems of agricultural production,
• Environmental protection,
• Horticulture,
• Landscape architecture.

university of warmia and mazury in olsztyn

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