CHAPTER 1 GENERAL INTRODUCTION

CHAPTER 1 GENERAL INTRODUCTION
CHAPTER 1
GENERAL INTRODUCTION
Beneficial use of sewage sludge on agricultural lands is a very well known
practice around the world. The benefits include; a source of essential crop
nutrients (Muse et al., 1991), improvements in soil structure (Ojeda et al., 2007),
and minimisation of soil erosion and runoff (Muse et al., 1991; Ojeda et al.,
2003). Nutrients applied above a crop’s nutrient requirement, however, can be
detrimental to plant growth (Brady, 1974) and will ultimately pollute water bodies
(Neal et al., 2002). In addition, waste products from cities and industrial areas
contain pathogens, toxic elements and organic contaminants which can pose a
serious health hazard. Therefore, many countries have developed sewage
sludge guidelines to optimize agricultural benefits without compromising
sustainability.
Beneficial agricultural use accounts for 28% of the total sludge produced from
South African wastewater treatment plants. This is despite the enormous
pressure on South African wastewater treatment plants to dispose of or utilize
their sludge in an environmentally sustainable way. The poor public perception of
sludge use and the lack of thorough local field scale studies have resulted in a
rigid guideline with single upper limit for all cropping systems. This situation
exacerbated the current low usage rates in agriculture.
26
Generally, sludge with acceptable quality for agricultural use is applied according
to crop N requirements (Mile and Graveland, 1972; Dotson, 1973). However,
sludge application according to crop N demand causes P build up in the soil
profile (Kelling et al., 1977; Pierzynski, 1994; Maguire et al., 2000a,b), although
its availability can be affected by sludge treatment processes in the water care
works (Soon et al., 1978a; Kirkham, 1982; McCoy et al, 1986; Jokinen, 1990;
Kyle and McClintock, 1995; Frossard., 1996; Maguire et al., 2001).
A large fraction of the N in sewage sludge is in the organic form, whereas plants
absorb N in the form of NH4 and NO3. The availability of N for plants depends
therefore on the rate of mineralization of the sludge (Kelley et al., 1984) and the
losses of inorganic nitrogen through volatilization, denitrification, and leaching. All
the above mentioned processes are influenced by soil type, availability of water,
soil temperature, and soil pH. Furthermore, the type of sludge treatment process
affects both mineralization and volatilization, while sludge application methods
largely influence ammonia volatilization (Henry et al., 1999). Therefore, direct
extrapolation of studies conducted in a specific soil type, climate, and sludge type
can compromise both the environment and crop yield.
A range of studies have shown that an increase in sludge application rate
increased grain and stover yield of agronomic crops (Binder et al., 2002, Cooper,
2005; Lavado et al., 2006; Bozkurt, 2006) as well as dry matter yield of pasture
grasses (Sullivan et al., 1997; Zebarth et al., 2000; Cogger et al., 2001). Crop
27
response to sludge application rate is, however, influenced by the availability of
water. For instance studies conducted by Soon et al. (1978b) showed a positive
maize growth response up until a sludge N supply of 200 kg ha-1 under dryland
as opposed to
400 kg N ha-1 under irrigation (Binder et al., 2002). Sludge
applications of 400 kg N ha-1 under dryland cropping resulted in the build up of
nitrate in the soil profile compared with similar rates under irrigation. In contrast,
pasture grass response to sludge application rate increased until a total N supply
of just above 800 kg ha-1, but such a high rate was associated with nitrate and P
accumulation in the soil profile (Cogger et al., 2001). The above shows that crop
nutrient demand is dynamic and is influenced by the cropping system and the
availability of water.
Despite this, the former South African sewage sludge guideline had a single
upper limit for all cropping systems of 8 Mg ha-1 yr-1 which was based on studies
conducted in other countries. It was recently updated to 10 Mg ha-1 yr-1 based on
local laboratory incubation studies and very few short-term (1- 2 years) local field
trials on specific crops. Such a guideline may be too inflexible for a range of
different cropping systems, sludge types, nutrient contents, climates, and soil
types. Both guidelines called for local field scale medium to long-term studies to
be carried out (Water Research Commission, 1997; Snyman and Herselman,
2006). This study was initiated after the 8 Mg ha-1 yr-1 guideline was released,
and was the first of its kind in South Africa to include agronomic benefits and
environmental impacts of different sludge rates on contrasting cropping systems.
28
The objectives of this study were
•
To investigate the dynamic nature of sludge application rate across a
range of management practices and cropping systems.
•
To investigate the agronomic benefits and sustainability of using municipal
sludge according to crop N demand
•
To evaluate the sustainability of high surface loading rates above crop
requirements for turfgrass sod production.
•
To adapt the N sub routine from the CropSyst model and include it into the
SWB model, and to test its reliability as a decision support tool for sewage
sludge management on agricultural lands.
In order to meet these objectives the following hypotheses were tested.
Agronomic crops:
1. Sludge application above the 8 Mg ha-1 yr-1 limit will increase dryland maize
and irrigated maize-oat rotation grain and forage yield.
2. Under both dryland and irrigated conditions, more N can be exported in
forage than grain.
3. Sludge applications that produce the highest yield under irrigated and dryland
conditions will not cause an accumulation of N in the soil profile and are not
susceptible to excessive nitrate leaching.
4. Sludge application according to crop N demand results in the accumulation of
total and plant available P in the soil profile.
29
Dryland pasture (Weeping lovegrass)
1. Sludge application above the 8 Mg ha-1 yr-1 limit will increase weeping
lovegrass hay yield, crude protein content, and water use efficiency.
2. The ideal sludge application rate to satisfy weeping lovegrass N demand is
dynamic and could exceed the 8 Mg ha-1 yr-1 sludge limit.
3. Under high hay yield productin conditions, N supply from double the norm can
fully be utilized and such systems are not prone to excessive nitrate leaching.
4. Sludge application according to crop N demand results in the accumulation of
total and plant available P in the soil profile.
Turfgrass
1. High sludge surface loading rates well above recommendations based on
crop removal
a. Are possible without reducing turf growth and quality.
b. Do not cause an accumulation of N and P below the active root zone.
c. Can minimize soil loss through sod harvesting, and
d. Do not cause unacceptably high nitrate and salt leaching.
Nitrogen and phosphorus availability from sludge is influenced by sewage sludge
treatment processes, climatic and edaphic factors. At the same time, crop
nutrient uptake is dependent on crop type, water availability, nutrient availability,
climate, soil type, management practice, and farming intensity. Ideally, specific
studies would need to be conducted for each sludge type, cropping system,
30
management practice, and climate, but this is impractical. Therefore this project
seeks to add a mechanistic N sub routine into an existing mechanistic soil water
balance and crop growth decision support model. The model is tested for
accuracy against the independent data sets collected from various cropping
systems described in this study. The predicting capability of the Model appear
promising and could be utilized by agronomists and sludge guideline developers
as a decision support tool for sustainable use of sludge across various types of
cropping systems, agro-ecological zones, and soil types. Phosphorus dynamics
from sludge, however, is not modelled in this study due to the complex nature of
P availability from sludge treated with Fe and Al salts, a subject that remains
contentious in the scientific literature.
The thesis is presented in the following order:
Chapter 2 is the literature review. It provides a general background followed by
discussion on sewage sludge types and characteristics, as well as beneficial
agricultural use. It contains a detailed description of nitrogen modelling before
concluding with motivation and rationale for the study.
Chapter 3 the Materials and Methods, presents information on field site
description, sludge characteristics, cropping systems and treatments, rainfall and
irrigation, plant and soil sampling, chemical analysis and statistical methods.
31
Chapter 4 presents the effect of various sludge application rates on grain and
forage yield under dryland (maize) and irrigated conditions (maize-oats rotation).
This is followed by measurement of grain and forage N uptake, soil profile N
mass balance, residual nitrate and ammonium, and nitrate leaching. Finally the
total P mass balance and residual Bray-1P of the cropping systems under
investigation is presented.
Chapter 5 covers dryland pasture (weeping lovegrass) and the effect of various
sludge application rates on hay yield, crude protein content, N uptake and water
use efficiency. Soil profile N mass balance, nitrate leaching, residual nitrate and
ammonium, total P mass balance and residual Bray-1P are presented.
Chapter 6 is on turfgrass. This chapter addresses the agronomic benefits and
disadvantage of very high sludge application rates and the associated negative
environmental impacts such as N and P accumulation below active root zone,
nitrate and salt leaching and soil loss through sod harvesting.
Chapter 7 deals with modelling. In this chapter the model is initially calibrated
using data collected from field studies and the literature. The calibrated model is
tested against independent data sets collected during the study period.
Chapter 8 summarizes important findings and forwards recommendations for
further studies.
32
REFERENCES
BINDER, D.L., DOBERMANN, A. SANDER, D.H. & CASSMAN, K.G., 2002.
Biosolids as nitrogen sources for irrigated maize and rainfed sorghum. Soil
Sci. Soc. Am. J. 66,531-543.
BOZKURT, M.A., AKDENIZ, H. KESKIN, B. & YILMAZ, I. H., 2006. Possibilities
of using sewage sludge as nitrogen fertilizer for maize. Acta. Agr. Scand.
B-S P. 56,143-149.
BRADY, N.C., 1974. The nature and properties of soils, 8th edn, Macmillan
Publishing Company, Inc., New York:
COGGER, C.G. BARY, A.I. FRANSEN, S.C. & SULLIVAN, D.M., 2001. Seven
years of biosolids versus inorganic nitrogen applications to tall fescue. J.
Environ. Qual. 30,2188-2194.
COOPER, J.L., 2005. The effect of biosolids on cereals in central New South
Wales, Australia: 1. Crop growth and yield. Aust. J. Exp. Agric. 45,435443.
DOTSON, G.K., 1973. Some constraints of spreading sewage sludge on
cropland. In proceedings of conference on land disposal of municipal
effluents and sludges. Mar. 12-13, 1973. Published by Rutgers University,
New Brunswick.
FROSSARD, E., SINAJI, S. & DUFOUR, P., 1996. Phosphorus in urban sewage
sludges as assessed by isotopic exchange. Soil Sci.Soc. Am. J. 60,179184.
HENRY, C., SULLIVAN, D. RYNK, R. DORSEY, K. & COGGER, C., 1999.
33
Managing nitrogen from biosolids. [Online]. Available at http://www.ecy.
wa.gov/pubs/99508.pdf (accessed 30 Mar. 2007; verified 04 Feb. 2008).
Washington.
JOKINEN, R., 1990. Effect of phosphorus precipitation chemicals on
characteristics and agricultural value of municipal sewage sludges. Acta
Agric. Scand. 40,141-147.
KELLING, K.A., PETERSON, A.E. WALSH, L.M. RYAN, J.A. & KEENEY, D.R.,
1977. A field study of the agricultural use of sewage sludge: I. Effect on
crop yield and uptake of N and P. J. Environ. Qual. 6,339-345.
KELLEY, W.D., MARTENS, D.C. RENEAU, R.B. & SIMPSON, T.W., 1984.
Agricultural use of sewage sludge: A literature review. Virginia Water
Resource Research Center, Virginia Polytechnic Institute and State
University, Blacksburg.
KIRKHAM, M.B., 1982. Agricultural use of phosphorus in sewage sludges. Adv.
Agron. 35:129-163.
KYLE, K.A., & MCCLINTOCK, S.A., 1995. The availability of phosphorus in
Municipal wastewater sludge as a function of the phosphorus removal
procedure and sludge treatment method. Water Environ. Res. 67,282-289.
LAVADO, R. S., RODRÍGUEZ, M. ALVAREZ, R. TABOADA, M.A. ZUBILLAGA
M.S., 2006. Transfer of potentially toxic elements from biosolid-treated soil
to maize and wheat crops. Agric. Ecosys. Environ. 118,312-318.
MAGUIRE, R.O., SIMS, J.T. & COALE, F.J., 2000a. Phosphorus solubility in
34
biosolids-amended farm soils in the Mid-Atlantic region of the USA.
J.Environ. Qual. 29,1225-1233.
MAGUIRE, R.O., SIMS, J.T. & COALE, F.J., 2000b. Phosphorus fractionation in
biosolids-amended
soils:
Relationship
to
soluble
and
desorbable
phosphorus. Soil Sci. Soc. Am. J. 64,2018_2024.
MAGUIRE, R.O., SIMS, J.T. DENTEL, S.K. COALE, F.J. & MAH, J.T., 2001.
Relationships between biosolids treatment process and soil phosphorus
availability. J. Environ.Qual. 30,1023-1033.
MCCOY, J.L., SIKORA, J. & WEIL, R.R., 1986. Plant availability of phosphorus
In sewage sludge. J. Environ. Qual. 15,403-409.
MILE, R. A. & GRAVELAND, D.N., 1972. Sewage sludge as a fertilizer
Can. J. Soil Sci. 52,270-273.
MUSE, J. K., MITCHELL, C.C. & MULLENS, G. L., 1991. Land Application Of
Sludge. Environmental quality. Agriculture and nature resources 607.
Alabama Cooperative Extension System. Auburn University. Auburn.
NEAL, C., JARVIE, H.P. WADE, A.J. & WHITEHEAD, P.G., 2002. Water quality
functioning of lowland permeable catchments: inferences from an
intensive study of the River Kennet and upper Thames. Sci. Tot. Environ.
282, 471–490.
OJEDA, G., ALCAŇIZ, J.M. & LE BISSONNAIS, Y., 2007. Differences in
aggregate stability due to various sewage sludge treatments on a
Mediterranean calcareous soil. Agric. Ecosyst. Environ.125,48-56.
PIERZYNSKI, G.M., 1994. Plant nutrient aspects of sewage sludge. In C.E. Clapp
35
et al. (ed.). Sewage sludge: Land utilization and the environment. Soil Sci.
Soc. Of Am., Madison, Wis.
SNYMAN, H.G. & HERSELMAN, J.E., 2006. Guidelines for the utilisation and
Disposal of wastewater sludge: v. 2 requirements for the agricultural use
of wastewater sludge. WRC rep. TT 262/06. Water Research Commission.
South Africa.
SOON, Y.K., BATES, T.E. BEAUCHAMP, E.G. & MOYER, J.R., 1978b. Land
application of chemically treated sewage sludge: I. Effects on crop yield
and nitrogen availability. J. Environ. Qual. 7,264-268.
SOON, Y.K., BATES, T.E. & MOYER, J.R., 1978a. Land application of
chemically treated sewage sludge: II. Effects on plant and soil
phosphorus, potassium, calcium, magnesium, and soil pH. J. Environ.
Qual. 7,269-273.
SULLIVAN, D.M., FRANSEN, S.C. COGGER, C.G. & BARY, A.I., 1997. Biosolids
and dairy manure as nitrogen sources for prairie grass on a poorly drained
soil. J. Prod. Agric. 10,589-596.
WATER RESEARCH COMMISSION., 1997. Permissible utilisation and disposal of
sewage sludge. 1st ed. WRC, Pretoria TT85/97.
ZEBARTH, B.J., MCDOUGALL, R. NEILSEN, G. & NEILSEN, D., 2000.
Availability of nitrogen from municipal biosolids for dryland forage grass.
Can. J. Plant Sci. 80,575-582.
36
CHAPTER 2
LITERATURE REVIEW
2.1 Background information
Population growth in cities and the expansion of industries is resulting in a rapid
increase in the volume of waste products that need to be either beneficially used
or disposed of in some way. Similar to other countries, South African wastewater
treatment plants are under enormous pressure to dispose of or utilize their
sludges and effluents in environmentally sustainable ways. The daily total
wastewater flow emanating from South African wastewater treatment plants was
estimated at 5400 Ml d-1 (Marx et al., 2004). Considering the potential benefits of
using sludge on agricultural lands, South Africa, like all other countries, have
developed sewage sludge guidelines for beneficial agricultural use (WATER
RESEARCH COMMISSION, 1997; Snyman and Herselman, 2006). Despite this,
only 28% of the total South African sludge is used for beneficial agricultural
purposes.
The first South African sludge guideline (WATER RESEARCH COMMISSION,
1997) was developed based on studies conducted in other countries and it
allowed a maximum annual application rate of 8 Mg ha-1 yr-1. The guideline was
updated in 2006 (Snyman and Herselman, 2006) based on local sludge-soil mix
incubation studies and few single year field scale trials on specific crops and
allows a maximum annual application rate of 10 Mg ha-1. These studies showed
37
the ideal mineralization rates and the potential for nitrate leaching under optimal
soil water and temperature conditions (Snyman and Van der Waals, 2004), which
is rare under field conditions. Actual sludge mineralization and ammonium
nitrification on the field is, however, variable depending on the sludge
composition, edaphic and climatic factors. Consequently, due to the lack of
sufficient local medium to long term field scale studies, there has been calls for
further study by both the former and current South African sewage sludge
guidelines. This study investigates the agronomic benefits of various sludge
application rates and the possible potential environmental impacts associated
with various sludge application rates on four major cropping systems: irrigated
maize-oat rotation, dryland maize, dryland pasture, and turfgrass.
In general South African sludges have mean N content of 3.85% and organic
matter content of 55% (Snyman and Herselman, 2006). Large fraction of the N in
sewage sludge is organic and should first be transformed to inorganic N before it
could be available for crop uptake. The processes involved in the transformation
of N from organic to inorganic are influenced by factors such as soil pH, soil
texture, bulk density, soil water, and soil temperature (Gilmour and Gilmour,
1980; Sims and Boswell, 1980; Parker and Sommers, 1983; Artiola and Pepper,
1992; Barbarick et al., 1996; Janssen, 1996; Leiros et al., 1999). In addition all
major processes involved in the nitrogen cycle of agricultural lands including crop
nutrient uptake are influenced by water availability, soil temperature, climatic
factors, soil type, management practices, crop type, and farming intensity.
38
The complex interaction between the different factors involved in the nitrogen
cycle of the soil – plant system makes it difficult to extrapolate field studies
across different agro-climatic conditions and soil types. Moreover, conducting
medium to long term field studies across various ecological zones, soil types,
sludge types, and cropping systems is not only expensive but also logistically
impractical. Nevertheless, a validated model could play a significant role in
extrapolating field data. Nitrogen subroutine was added to the existing SWB
model as part of the study. The potential of the new SWB model with N
subroutine to extrapolate to other locations will be tested using data collected
from this study.
2.2. Sewage sludge types, characteristics, and agricultural use
2.2.1 Sewage sludge types
The characteristics of sewage sludge vary depending on its treatment. Primary
sewage sludge is the result of the primary settling of solids from wastewater, and
it has not yet undergone any treatment process. It has unpleasant characteristics,
is full of pathogens and is not recommended for land application (US EPA, 1995;
Spinosa and Vesilind, 2002).
Secondary sewage sludge is generated through the biological treatment of
primary sludge followed by clarification. Tertiary sewage sludge is generated
39
through further processing by chemical precipitation (using aluminium, iron, lime,
or organic polymers) and filtration (US EPA, 1995). The most common sludge
treatment processes with their corresponding effects on sludge properties and
land application practices are summarized in Table 2.1.
40
Table 2.1 Effects of sewage sludge treatment processes on sludge properties and land application practices
(Adapted from US EPA, 1984).
Sludge
category
Primary
Secondary
Treatment process
Thickening
Digestion (Anaerobic and
Aerobic)
Alkali stabilization
Conditioning
Tertiary
Dewatering
Composting
Heat drying
Effect on sewage sludge properties
Increases solids concentration by removing water,
thereby lowering sewage sludge volume
Reduces the volatile and biodegradable organic
content and the mass of sewage sludge by
converting it to soluble material and gas. Reduces
pathogen levels and odour.
Raises sewage sludge pH. Temporarily decreases
biological activity. Reduces pathogen levels and
controls putrescibility and odour.
Improves sewage sludge dewatering characteristics.
May increase the mass of dry solids to be handled.
May improve sewage sludge compactability and
stabilization.
Increases solid concentration of organic sewage
sludges to 15% - 40%, and for some inorganic
sewage sludges to 45% or more. Improves ease of
handling by converting liquid sewage sludge to damp
cake. Some nitrogen and other soluble materials are
removed with the water.
Lowers biological activity. Can destroy most
pathogens. Degrades sewage sludge to humus-like
material. Increases sewage sludge mass due to
addition of bulking agent.
Disinfects sewage sludge. Destroys most pathogens.
Slightly lowers potential for odour and biological
activity.
41
Effect on land application
practice
Lowers sewage sludge
transportation costs.
Reduces sewage sludge quantity.
Therefore lowers sludge
transportation costs
High pH immobilizes metals as
long as pH levels are maintained,
thus reducing heavy metal
leaching and crop uptake
Polymer-treated sewage sludge
may require special operational
considerations at the land
application site.
Lowers sewage sludge
transportation costs.
Excellent soil conditioning
properties. Significant storage
space usually needed. May
contain lower nutrient levels than
less processed sewage sludge.
Greatly reduces volume of sewage
sludge. Therefore lowers
transportation cost.
2.2.2 Nitrogen and sewage sludge
Most of the nitrogen in sewage sludge is in organic form, whereas plants absorb
N in the form of NH4 and NO3. The availability of N for plants depends therefore
on the rate of mineralization of the sludge (Kelley et al., 1984) and the losses of
inorganic nitrogen through volatilization and denitrification (Fig. 2.1).
Sewage sludge treatment and handling affects the characteristics of N in various
ways.
i) Dewatering reduces the amount of water soluble inorganic N (NH4 and
NO3)
ii) Heat or air drying and lime treatment reduce NH4 through volatilization in
the form NH3, but it does not affect NO3,
iii) Aerobic treatment facilitates the conversion NH4 to NO3, while anaerobic
conditions inhibit oxidation of NH4 to NO3.
a) Nitrogen mineralization
Mineralization is the transformation of N from its organic state to the inorganic
forms of NH4+ or NH3 by heterotrophic soil organisms consuming nitrogenous
organic substances as a source of energy (Lutz, 1965; Jansson and Persson,
1982). The release of ammonium during mineralization is driven by the need of
micro-organisms for carbon (Sprent, 1987).
42
N2
NO
N2O
Biological N fixation
(legume plants)
Industrial
N fixation
(inorganic
fertilizers)
Natural N fixation
Denitrification
NH3 volatilization
Manure
Plant residues
Organic matter
Immobiliz
ation
Mineralization
NH4+
Fixation to
clay minerals
Plant assimilation
Nitrification
NO3Nitrate leaching
Figure 2.1 Simplified nitrogen cycle in terrestrial plant-soil system.
The organic N pool is usually divided into three pools based on the potential rate
of mineralisation. These are i) a rapid turnover pool made up of amino acids and
43
proteins, ii) a slow turnover pool, and iii) a highly resistant pool that is often not
detected in short period soil incubation trials (Henry et al., 1999; Smith et al.,
1997). The size of each pool depends on the type of sludge and its treatment.
The mineralisation of each pool is influenced by environmental factors
(temperature and water content) and edaphic factors (soil pH, texture, structure,
and soil fauna) (Leiros et al., 1999; Janssen, 1996; Parker and Sommers, 1983;
Trindade et al., 2001).
The variation in mineralisation rates between different sludge types during the
year of application is mainly as a result of the size of the rapidly mineralisable
organic nitrogen pool, assuming similar environmental and edaphic factors
(Henry et al., 1999, Table 2.2). Mineralisation also continues at a slower rate for
the following 3 years, after which it is considered negligible (Table 2.3). Results
from many long-term studies have shown that four or more years after sludge
application, 20 to 50 percent of the organic nitrogen remains in the soil as stable
organic matter (Henry et al., 1999).
Organic matter decomposition (nitrogen mineralization) increases as the soil
water content increases from permanent wilting point to field capacity and as the
temperature increases from cold to warm (Vinten and Smith, 1993). Under
controlled warm and moist laboratory conditions, the rapidly mineralisable pool
decomposes in two to four weeks. In situations where either the soil water or
temperature is not ideal, it extends from two to six months. The slowly
44
mineralisable and highly resistant pools, however, take months or even years for
complete decomposition (Henry et al., 1999).
Table 2.2 Estimates of nitrogen mineralization for various sludge treatment
methods in the year of application (percent of initial organic N) (adapted from
Henry et al., 1999).
Treatment method
Mineralization (% of initial organic N)
Liquid
20 – 40
Anaerobically
Dewatered
25 – 45
digested
Heat-dried
25 – 45
Aerobically digested
30 – 50
Lagooned
10 – 30
Lime-stabilized
30 – 60
Composted
0 – 30
Drying bed
15 – 40
Oxidation ditch
30 - 50
Table 2.3 Nitrogen mineralization rate estimate ranges for all types of sludge for
years following the application year (percent of the remaining organic N)
(adapted from Henry et al., 1999).
Year following application Mineralization rate
year
(% of remaining organic nitrogen)
1
5 – 12
2
2–6
3
1-2
45
b) Ammonia volatilization
Ammonia volatilization is the gaseous loss of NH3 from the soil surface to the
atmosphere (Haynes and Sherlock, 1986). Volatilization takes place as a result of
the difference in vapour pressure gradient between ammonia in solution and the
ambient air (Freney et al., 1983) and can be very large. Fine et al. (1989)
reported a loss of 87% of the mineralised nitrogen through ammonia volatilisation
from activated sewage sludge. Other studies conducted by BEAUCHAMP et al.
(1978), indicate that about 20% of total nitrogen applied was lost as ammonia
during the first week of decomposition from anaerobically digested liquid sludge.
Ammonia volatilization is highly dependant upon the sludge treatment process,
and whether the sludge is left on the surface or incorporated (Table 2.4).
Volatilization is increased by higher soil water content, air temperature, and wind
speed (Henry et al. 1999) and suppressed at low pH, cool temperatures, and low
wind speeds (Freney et al., 1983).
Ammonia volatilization is high from initially wet soils (below saturation), which are
allowed to dry slowly (Fenn and Escarzaga, 1977). This is because evaporative
loss of water promotes ammonia volatilization by increasing or maintaining the
concentration of ammonia in the soil solution over time (Haynes and Sherlock,
1986). On the other hand, there is no ammonia volatilization from very low soil
water content soils (Nelson, 1982). It is also well documented that ammonia
volatilization is high from surface application of NH4+ compound fertilizers
46
compared with incorporated fertilizers (Fenn and Kissel, 1976; Quemada et al.,
1998).
Table 2.4 Ammonia volatilization rates from Northwest Biosolids applied in
western Washington (maritime climate) (adapted from Henry et al., 1999).
Volatilization rate
(%
Treatment method
of
initial
ammonia lost)
Liquid (incorporated)
20 – 40
Agronomic rate
14 – 50
Double
Anaerobically
digested
Dewatered
agronomic rate
25 – 49
Lime amended
45 – 134
Incorporated Reduced pH
12 – 28
Surface applied
51 - 127
Aerobically digested (incorporated)
6 – 12
Lagooned (incorporated)
4 – 20
Lime-stabilized (incorporated)
14 – 22
Drying bed (incorporated)
2–5
Oxidation ditch (incorporated)
9 - 23
c) Denitrification
Denitrification is the gaseous loss of nitrogen in the form of nitrous oxide (N2O)
and dinitrogen (N2), mediated through microbial action (Fig. 2.1) (Delwiche, 1981;
Tisdale et al., 1985). The microorganisms responsible for the denitrificaction
need nitrogen in the form of nitrate, soluble organic carbon and water-logged
(anaerobic) conditions (Henry et al., 1999). Sewage sludge land application
47
provides two of these three conditions (nitrate and soluble carbon) thereby
making denitrification highly susceptible to wet conditions. It was clearly
illustrated by Henry et al. (1999) that denitrification is greater under irrigated
systems compared with dryland conditions (Table 2.5).
Table 2.5 Suggested denitrification values for sludges applied to agricultural
lands in the Pacific Northwest, USA (adapted from Henry et al., 1999).
Farming system
Denitrification rate (% of inorganic N lost)
Non-irrigated
0
Irrigated
0 -15
2.2.3 Phosphorus and sewage sludge
The most common forms of inorganic P in wastewater aqueous solutions are
orthophosphates and polyphosphates. Orthophosphates are available for
biological metabolism without further breakdown while polyphosphates usually
should undergo hydrolysis and revert to orthophosphates (Crites et al., 2005).
Generally the concentration of total P in municipal wastewaters may range
between 4 to 12 mg L-1 of which 1 to 4 mg L-1 is organic. Secondary wastewater
treatments can only remove 1 to 2 mg L-1 P (Metcalf and Eddy, 2003). This is in
contrast to the soil solution P concentration benchmark of 1 mg L-1 for
wastewater discharge to rivers and streams (Sims and Pierzynski, 2000).
Therefore, direct discharge of wastewater effluents after secondary treatments
without further P removal processes could cause eutrophication of surface
waters. Wastewater treatment plants, therefore, use either chemical or biological
48
phosphorus removal methods to bring down the concentration below the
benchmark before they could discharge it to rivers and streams. Chemical
precipitation is used to remove the inorganic forms of phosphate by the addition
of coagulants and mixing of wastewater and coagulant. The most commonly
used coagulants are calcium, aluminium, and iron (Tchobanoglous et al., 2003).
In the biological phosphorus removal method, the P accumulating organisms are
encouraged to grow and consume P and are subsequently removed (Storm et
al., 2004).
Sewage sludge treatment methods and chemical or biological nutrient removal
processes influence the availability of P (Frossard et al., 1996; Maguire et al.,
2001; Penn and Sims, 2002; Kirkham, 1982; McCoy et al., 1986). The
predominant form of P in sludges that have undergone tertiary treatment is
inorganic P (McLaughlin, 1984). Chemicals used in tertiary treatments such as Al
or Fe salts, decrease the plant available P fraction (Elliott et al., 2002; Häni, et
al., 1981; Kyle and McClintock, 1995). Therefore, the percentage of total P found
in the easily soluble fraction is higher in sludges not treated with Fe or Fe + Al.
The application of lime to the Fe or Fe + Al treated sludges, however, increased
the concentration of the easily soluble P fraction (Penn and Sims, 2002).
In work done by Penn and Sims (2002), a significant increase in total soil P was
observed in sludge amended silty clay loam and sandy loam soils compared with
their control (no sludge). The labile forms of soil P (M3-P, M1-P, FeO-P, and
49
water soluble P) were especially high for soils that received biological nutrient
removal sludge, compared to the control and the soils receiving Fe and lime
treated sludge treatments. Soils that received sludge treated with Fe and Al salts,
slightly increased the labile soil P, but soils which received sludge treated with Fe
and lime resulted in an intermediate labile P increment (Penn and Sims, 2002).
Long term studies conducted on wastewater applied to a sandy soil for 30 to 50
years, showed an increase in the inorganic form of P, that was dominated by Albound phosphate (Beek et al., 1977). Other studies conducted by Soon and
Bates (1982) on soils amended with Ca, Al, and Fe treated sludges show that
significant increase in soil P was in the Al- and Fe-bound P fraction, irrespective
of the sludge type used. Therefore, it is of the utmost importance to consider the
type of sludge used when quantifying sludge application rates. This can help to
optimise crop harvests by minimising the environmental impacts.
2.2.4 Utilising sewage sludge on agricultural lands
Beneficial use of sewage sludge on agricultural lands is a common practice in
various regions of the world. This can be clearly observed from Table 2.6 which
presents the percentage of sewage sludge applied to agricultural lands for 15
European Countries, USA, Australia, and South Africa. Beneficial use accounts
for 28% of the total sludge produced from South African wastewater treatment
plants, which places South Africa 11th in the list of Table 2.6. The remaining 72%
is accounted for through land fill (3%), non beneficial land application (47%),
50
accumulation at plant (20%), and unspecified (2%) (Du Preez, et al. 2000). A
combination of factors including low sludge quality due to pathogen and
pollutants as well as lack of sufficient information on the beneficial use of sludge
contributes to the relatively low sludge usage in agriculture.
Table 2.6 Annual sewage sludge produced and the percentage applied to
agricultural lands for 15 European Countries and USA. (USA and EU (AEA
Technology Environment, 2002); Australia (Priestley, 1991); South Africa
(Lötter and Pitman, 1997))
Country
Austria
Belgium
Denmark
France
Germany
Greece
Ireland
Italy
Luxembourg
Holland
Portugal
Spain
Sweden
United Kingdom
Switzerland
USA
Australia
South Africa
Annual
dry
sludge % of sludge produced applied to
production (Mega tons)
agricultural lands
0.320
13
0.075
31
0.130
37
0.700
50
2.500
25
0.015
3
0.024
28
0.800
34
0.015
81
0.282
44
0.200
80
0.280
10
0.180
45
1.107
55
0.215
50
6.900
41
0.300
9
0.310
28
The major reason for applying sludge is as a source of plant nutrients. Optimal
agricultural production with minimal environmental impacts can be achieved
through proper sludge application and management practices. These practices
include sludge application according to cropping system nutrient demand and
proper sludge application timing and methods. Since most of the nitrogen in
51
sewage sludge is bound in organic form, it is released slowly compared to
inorganic fertilizers, as can be seen in Figure 2.2. The organic matter in sewage
sludge is also a good source of energy for soil micro flora (Muse et al., 1991).
Figure 2.2 Nitrogen requirement of maize during the growing season and
nitrogen availability from fertilizer compared with sludge (adapted from Muse et
al., 1991).
The organic matter in sewage sludge applied to agricultural lands also improves
the soil’s physical characteristics and structure, thereby improving soil porosity,
stability of aggregates and water retention (Ojeda et al., 2007). It can play a
significant role in improving the water holding capacity of sandy soils, aeration
and water movement in clay soils, eases plant root penetration, and reduces
runoff and soil erosion (Muse et al., 1991). It also improves soil chemical
characteristics by increasing cation exchange capacity and also as a source of
macro and micro nutrients for plants (Mininni and Santori, 1987).
52
2.2.5 Sludge application rates on agricultural lands
Generally, application rates of sewage sludge to agricultural lands are dictated by
the nitrogen content of the sludge (Mile and Graveland, 1972; Dotson, 1973). In
most US States, the sewage sludge application rate to agricultural land is based
on the N requirements of crops and the concentrations and loading rates of other
trace elements, as defined in the US EPA 503 rule (US EPA, 1994). However,
phosphorus based management is considered in areas with high soil P test (Sims
et al., 2000).
Previous research by Kelling et al., (1977); Pierzynski (1994), Peterson et al.
(1994) and Maguire et al. (2000a,b) indicate that continuous sludge applications
based on nitrogen demand will cause soil P to accumulate to levels above those
needed for optimum crop production. Phosphorus availability from sludge
amended soils, however, depends on the type of treatment and processes which
the sludge went through in the water care works (Kyle and McClintock, 1995;
Maguire et al., 2001; Jokinen, 1990; Soon et al., 1978a; Kirkham, 1982; McCoy
et al., 1986; Frossard et al., 1996).
The province of Ontario, Canada, also monitors the soil for possible P
accumulation at sludge disposal sites. They recommend that bicarbonate
extractable P levels should not exceed 60 kg P ha-1, the concern being P
contamination of waters by soil erosion or runoff (McLaughlin, 1984). Sims and
Gartley (1996), classified soil P status based on M1-P analyses as follows: 0 –12
53
mg kg-1 = low, 13 –24 mg kg-1 = medium, 25 – 50 mg kg-1 = optimum and > 50
mg kg-1 = excessive. Considering such a classification for agricultural lands
receiving sludge could minimise ground and surface water contamination through
leaching and runoff, thus improving the sustainability of sludge use on agricultural
lands.
Similar to most US States, the current South African guideline is based on the
nitrogen needs of crops (sludge suitable for agricultural use) (Snyman and
Herselman, 2006). However, unlike the US guideline, it does not consider
phosphorus based management.
Sludge application timing and methods
It is advisable to schedule sludge applications on agricultural lands around the
time of tillage or planting. However, it depends on the type of sludge, type of soil,
crop, climate, time, and method of application (Shepherd, 1996). Correct sludge
application timing is essential for efficient use of nutrients and to minimise
possible ground or surface water pollution (Evanylo, 1999).
Sludges are commonly applied to agricultural lands either on the soil surface
(through spraying or spreading) or incorporated into the soil (through injection or
agricultural implements). Surface application is most commonly used on
pastures, range and forest lands. Sludge incorporation, on the other hand, is
most commonly used for agronomic crops (Evanylo, 1999).
54
Surface application of liquid sludge is conducted using tractor drawn tank
wagons, special applicator vehicles equipped with flotation tyres, or irrigation
systems. It is usually restricted for use in areas with slopes less than 7 percent.
The disadvantages of spraying liquid sludge on the surface are mainly potential
odour problems and the reduction in the aesthetic value of the application site. To
avoid the risk of runoff losses and excess leaching below the root zone, liquid
sludge should preferably be applied in split rather than a single big application
(Evanylo, 1999).
Liquid sludge can also be injected below the soil surface. This method minimizes
odour problems, reduces ammonia volatilization, minimizes runoff losses and can
be used in areas with slopes of up to 15 percent. Liquid sludge injection can be
conducted using tractor-drawn tank wagons with injection shanks or tank trucks
fitted with flotation tyres and injection shanks. Dewatered sludges are usually
surface applied to crop lands using equipment similar to that used for applying
limestone, or animal manures. The sludge is then incorporated into the soil by
ploughing (Evanylo, 1999).
Consequences of very high sludge applications
Excess nutrients are detrimental to plant growth (Brady 1974), and pollute water
bodies (Kumm, 1976; Keeney, 1973). As water moves over the soil surface it
carries along decomposing organic matter, fertilizers, pesticides, as well as
55
sediments. The nutrient loads of such runoff waters result in nutrient enrichment
of water bodies such as lakes and dams. This nutrient enrichment is the cause of
undesirable proliferation of aquatic plants (eutrophication). Nitrogen in water
bodies is oxidized into nitrite and nitrate, which depletes levels of dissolved
oxygen (Neal et al., 2002; Sparks, 1995) resulting in the death of fish and other
aquatic creatures (Cameron and Haynes, 1986).
Water that percolates beyond the root zone may also carry nutrients in a
dissolved form, thereby contaminating groundwater. Contaminated water
presents a hazard for humans and animals, as well as for plants. A consequence
of groundwater nitrate pollution is methemeglobinemia or blue baby syndrome,
which can cause death in infants (Croll and Hayes, 1988).
2.2.6 Classification of sludge for use on agricultural lands
Waste products from cities and industrial areas contain pathogens, toxic
elements and organic contaminants which can pose serious health hazards.
Some of the pathogens that can possibly be found in sewage sludge before its
treatment and the diseases they cause are presented in Table 2.7 (U.S. EPA,
1995).
Heavy metals of concern that can be found in sewage sludge include As, Cd, Cu,
Pb, Hg, Mo, Ni, Se, Zn, and Cr. Organic contaminants that can be found in
sewage sludge include pesticides (chlordane, endrin etc.), herbicides (2.4-D,
56
2.4.5-TP Silvex), volatiles (benzene, carbon tetrachloride etc.), and semi volatiles
(O-Cresol, m-Cresol etc.) (U.S. EPA, 1995). Sewage sludge must therefore be
used within certain guidelines, and many countries have published sewage
sludge regulations for land application (U.S. EPA, 1995; Wang, 1997; Krogmann
et al., 2001; Snyman and Herselman, 2006).
Table 2.7 A few of the pathogens that could potentially be present in municipal
sewage sludge and the diseases or symptoms they cause (adapted from U.S.
EPA, 1995).
Pathogen
Salmonella sp.
Escherichia coli
Shigella sp.
Hepatitis A virus
Echoviruses
Entamoeba
histolytica
Giardia lamblia
Ascaris sp.
Trichuris trichiura
Toxocara canis
The
South
African
Disease/symptoms
Salmonellosis (food poisoning), Typhoid fever
Gastroenteritis
Bacillary dysentery, severe gastroenteritis
Infectious hepatitis
Meningitis, paralysis, encephalitis, fever,
symptoms, diarrhoea, etc.
Amoebic dysentery
“flu-like”
Diarrhoea, abdominal cramps, weight loss
Digestive and nutritional disturbances, abdominal pain,
vomiting, restlessness, coughing, chest pain, and fever
Abdominal pain, diarrhoea, anaemia, weight loss
Fever, muscle aches, neurological symptoms
Sewage
Sludge
Guideline
classifies sludge
under
microbiological, stability, and pollutant classes (Snyman and Herselman, 2006).
Tables 2.8 and 2.9 present the microbiological and pollutant classes of the
current South African guideline, compared with those of the USA and EU and
Table 2.10 presents the stability classes of the current South African guideline.
57
Table 2.8 South African preliminary classification: microbiological class (Snyman
and Herselman, 2006) compared with the USA (US EPA, 1995); (US EPA,
2003).
South African microbiological class
B
C
Sample
One or
Two of
that failed
All three
three
doesn’t
more of the
samples
samples
exceed the
samples
comply
comply
following
exceed
A
Organism
Faecal
coliforms
(CFU/gdry)
Helminth
ova (Total
viable
ova/gdry)
Salmonela
sp.
(CFU/4gdry)
Enteric
viruses
(PFU/4gdry)
< 1000
< 1 × 10
6
< 1 × 10
7
>1 × 10
7
US microbiological class
B
Geometric
Representative
mean of 7
sample at time samples at
of use or
time of use/
disposal
disposal
A
< 1000
< 2 × 10
< 0.25 or
(one
viable
ova/4gdry)
N/A
<1
4
>4
<1
N/A
N/A
N/A
N/A
<3
N/A
N/A
N/A
N/A
N/A
<1
N/A
CFU - colony forming units
PFU - plaque-forming units
6
N/A – not applicable
The current South African sewage sludge guideline has similar pollutant class
ranges with US guideline for Cr, Cu, Pb, Ni, Zn, As, Cd, and Hg. The EU
guideline is, however, very conservative compared to both the South African and
US guidelines. The US guideline adds Se to the list of pollutants, which is not
considered in either the South African or EU guidelines. The South African and
US microbiological class “A” ranges for faecal coliforms and total viable helminth
ova are similar. The US guideline, however, adds salmonella and enteric viruses
to the list.
58
Table 2.9 South African preliminary classification: pollutant class (Snyman and
Herselman, 2006) compared with the US land application pollutant limits (US
EPA, 1995) and proposed EU maximum permissible limits in sludge in mg kg-1
(IC Consultants, 2001).
Element
As
Cd
Cr
Cu
Pb
Hg
Ni
Zn
Se
Pollutant limits (mg kg-1)
South African pollutant class
US land application pollutant limit
a
b
c
Concentration
Ceiling
in bulk and
concentration in
bagged
all sludge that
sewage sludge
is land applied
<40
40 – 75
>75
41
75
<40
40 – 85
>85
39
85
<1200
1200 –
>3000
1200
3000
3000
<1500
1500 –
>4300
1500
4300
4300
<300
300 –
>840
300
840
840
<15
15 – 55
>55
17
57
<420
420
>420
420
420
<2800
2800 >7500
2800
7500
7500
N/A
N/A
N/A
36
100
EU proposed
maximum
limit
N/A
10
1000
1000
750
10
300
2500
Bagged sewage sludge is sold or given away in a bag or other container for application to land (US EPA, 1995)
N/A
N/A - not applicable
The current South African sewage sludge guideline considers all three categories
(microbiological, stability, and pollutant classes), together to assess the
appropriateness of using a given type of sludge for agricultural purposes.
According to the microbiological classification, a sludge with microbiological class
“A” can be applied without restriction at agronomic rates, class “B” may not be
appropriate to use for some crops with edible parts below the soil surface, and
class “C” can only be used for agricultural application if stability class 1 or 2 is
achieved and this is restricted for certain crops (Snyman and Herselman, 2006).
59
Table 2.10 South African preliminary classification: Stability class (Snyman and
Herselman, 2006).
Stability class
1
2
3
Plan/design to
Plan/design to
No stabilisation or
comply with one
comply with one
vector attraction
of the options
of the options
reduction options
listed below on a listed below on a required.
90 percentile
75 percentile
basis.
basis.
Vector attraction reduction options (applicable to stability class 1 and 2
only)
Option 1 Reduce the mass of volatile solids by a minimum of 38 percent
Option 2 Demonstrate vector attraction reduction with additional anaerobic
digestion in a bench-scale unit.
Option 3 Demonstrate vector attraction reduction with additional aerobic
digestion in a bench-scale unit.
Option 4 Meet a specific oxygen uptake rate for aerobically treated sludge
Option 5 Use aerobic processes at a temperature greater than 40oC
(average temperature 45oC) for 14 days or longer (eg. during
sludge composting)
Option 6 Add alkaline material to raise the pH under specific conditions
Option 7 Reduce moisture content of sludge that do not contain unstabilised
solids (from treatment processes other than primary treatment) to
at least 75 percent solids
Option 8 Reduce moisture content of sludge with unstabilized solids to at
least 90 percent solids
Option 9 Inject sludge beneath the soil surface within a specified time,
depending on the level of pathogen treatment
Option 10 Incorporate sludge applied to or placed on the surface of the land
within specified time periods after application to or placement on the
surface of the land
According to the stability classification, sludge with stability class “1” can be
applied without restriction at agronomic rates, class “2” can also be applied at
agronomic rates but additional management systems need to be conducted,
however, class “3” can not be used for agricultural application.
60
According to the pollutant classification, sludge of class “a” can be applied
without restriction at agronomic rates; class “b” can also be applied at agronomic
rates if analyses of the receiving soil prove it can accommodate the load,
however, class “c” can not be used for agricultural purposes (Snyman and
Herselman, 2006).
Table 2.11 Permissible utilisation of sludge in agricultural applications based on
the South African sludge classification system (adapted from Snyman and
Herselman, 2006)
South African
Any additional
restrictions and
classification
an option?
requirements?
Notes
A
Yes
No
B
Qualified yes
Yes
C
May be
Yes
1
Yes
No
2
Qualified yes
Yes
3
No
Not applicable
a
Yes
No
b
c
Qualified yes
No
Yes
Not applicable
Could potentially be used as a
saleable product.
General restrictions/ requirements
apply.
Only permissible if stability class 1 or 2
is achieved. (General restrictions/
requirements apply)
Could potentially be used as a
saleable product.
Additional
management
actions
required to encourage compliance with
class 1.
Stability class 3 may not be used in
agricultural practices.
Could potentially be used as a
saleable product.
If the soil analyses is favourable.
Pollutant class c may not be used in
agricultural practices.
class
Pollutant
Stability class
class
Is agricultural use
Microbiological
sludge
2.2.7 Experiences with sewage sludge on cropping systems
a. Grain crops
A range of studies have shown that maize stover and grain yield increased with
increase in sludge application rate regardless of the soil type (Cunningham et al.
61
1975; Kelling et al. 1977; Binder et al. 2002; Lavado et al. 2006; Bozkurt 2006).
In some cases the response was non linear, with smaller yield increases at high
application rates (Soon et al., 1978b; Binder et al., 2006). At very high application
rates, grain yield has been reduced by salt toxicity (Cunningham et al., 1975).
The optimal sludge application rate for nitrogen depends on the availability of
water. Soon et al. (1978b), found no crop response for sludge rates containing
above 200 kg N ha-1 under dryland conditions, both in a clay loam and loamy
sand soils. In those studies the mean above ground biomass N uptake was 85 kg
ha-1. In other studies conducted by Binder et al. (2002) under irrigated conditions,
however, maize grain yield increased significantly with increase in sludge
application rates which supplied N up until 400 kg ha-1. The mean above ground
biomass N uptake in this irrigated system was just above 300 kg ha-1.
Significant accumulation of nitrate in the top 0.75 m soil layer was reported by
Soon et al., (1978b) following three years of sludge application at rates of 400 kg
N ha-1 under dryland maize production. In other studies conducted by Binder et
al. (2002) under irrigation, however, there was no significant accumulation of N at
this rate of 400 kg N ha-1 mainly due to the higher nitrogen use efficiency of
similar crop under irrigation. The latter study, however, reported a significant
nitrate accumulation when sludge was applied at rates of 950 kg N ha-1 due to
the decline in the nitrogen use efficiency. The accumulation of nitrate in the soil
62
profile could lead to ground water pollution especially in the beginning of the
season if heavy rain falls before the active crop root development.
Since a large fraction of N in sludge is organic pool, there may be significant
carry over effects in the subsequent years after application (Cogger et al., 2001;
Binder et al., 2002). It is therefore important to quantifying the amount of carry
over effects each year before new sludge applications.
The yield of small grain cereals such as wheat (Triticum aestivum) and triticale
(Tritico-secale) increased with sludge application (Oloya and Tagirwa, 1996;
Cooper, 2005; Lavado et al., 2006) up to a certain point. After this, the effects of
high salt dominated (Simeoni et al. 1984).
b. Pastures
It is well documented that sludge application improves dry matter yield of pasture
grasses (Kelling et al., 1977; Soon et al., 1978b; Kiemenec et al., 1987; Michalk
et al., 1996; Sullivan et al., 1997; Zebarth et al., 2000; Cogger et al., 2001),
although responses are dependant on climate and management regimes. For
instance, tall fescue (Festuca arundinacea Schreb.) dry matter yield increased
significantly with increase in sludge application up to rates of 20 Mg ha-1 (905 kg
N ha-1) in an area receiving 1020 mm rainfall and supplemented with irrigation as
required (Cogger et al., 2001) during summer season. In other studies conducted
by Sulivan et al. (1997), prairie grass (Bromus uniloides (Willd)) dry matter yield
63
increased as sludge application rate increased to 26.9 Mg ha-1 (1064 kg N ha-1)
in a 1232 mm rainfall area. Kelling et al. (1977) reported positive rye grass
(Secale cereale L.) response for sludge rates of up to 7.5 Mg ha-1 in a silty loam
soil at Arlington but up to 15 Mg ha-1 in a sandy loam soil at Janesville,
Wisconsin, USA. He argued that this was most probably due to the combination
of factors including the better initial nutrient status of the soil in Arlington.
Although high sludge applications of 20 Mg ha-1 yr-1 (>800 kg N ha-1) significantly
improved tall fescue dry matter yield, this rate was associated with the
accumulation of NO3 across time (Cogger et al., 2001). Surprisingly, similar
findings were reported from earlier studies conducted on Brome grass in a similar
soil type (Soon et al., 1978b). According to this study, nitrate accumulated in the
soil profile to a depth of 0.9 m at higher sludge application rates supplying 800 kg
N ha-1 and above. Other negative environmental impacts of very high sludge
application rates are the accumulation of plant available P above the
concentration required for optimal crop production of 20-50 mg kg-1. This has
been reported in studies conducted by Michalk et al. (1996), where the plant
available P exceeded 300 mg kg-1 on the soil surface. Such high rates could
definitely pose risk to surface water bodies through runoff in solution form (Sims
and Pierzynski, 2000).
In addition to the obvious negative environmental impacts associated with high
sludge loading rates, there are also some concerns related to fodder quality
64
(nitrate concentration in grass). These concerns are extended to the health
issues of workers harvesting forage and animals consuming the forage. Studies
conducted by Soon et al (1978b) showed that high sludge loading rates that
supplied 1600 kg N ha-1 increased the concentration of NO3-N in the grass to
hazardous levels for livestock consumption (>3 mg g-1). Regarding the health
issue of workers and foraging animals, studies conducted by King and Morris
(1972) showed that the forage was essentially coliform-free at harvest time,
indicating that there should be little danger of disease transmission to workers
handling the harvested forage and the animals consuming the forage. This,
however, depends on the microbiological classification of the sludge.
c. Lawn sod production
The application of composted sludge on various turfgrass species planted to
different soil types has improved turf growth and quality. For instance the
incorporation of composted biosolid in a disturbed urban soil facilitated the
establishment rate of two turfgrass species, Kentucky bluegrass (Poa pratensis
L.) and perennial ryegrass (Lolium perenne L.) (Loschinkohl and Boehm, 2001).
Similarly, composted biosolid incorporated into a sandy loam B-horizon improved
kentucky bluegrass turf vegetative cover, colour, and density (Linde and Hepner,
2005). In was also evident from studies conducted by Landschoot and McNitt
(1994) that increasing composted biosolid application rate increased kentucky
bluegrass turf cover dramatically (greater than 80% one month after seeding). In
addition to improving turfgrass quality, the addition of composted sludge
65
increased soil organic matter content, bulk density, and increases water
infiltration rate compared with an unamended control (Landschoot and McNitt,
1994; Cheng et al., 2007).
2.3 Nitrogen modelling
Sewage sludge treatment and handling processes affect the characteristics and
availability of nutrients from sludge. At the same time, crop nutrient uptake is
dependent on crop type, water availability, nutrient availability, climate, soil type,
management practice, and farming intensity. Therefore studies conducted in a
specific crop under specific conditions can only serve to produce guidelines for
that specific situation.
Direct application of such guidelines to different situations can compromise both
the environment and crop yield. A validated mechanistic model can play a
significant role to extrapolate results responsibly, by conducting scenario
simulations for various types of crops grown on various soil types and under
different climatic regimes. This will reduce the requirements of performing costly
long-term site specific field trials and information generated can be used in the
development of updated site specific guidelines for sludge application where
needed.
Modelling the dynamics of nitrogen in the soil system dates back to an early 70s
(Shaffer et al., 2001). The first integrated soil system N models include that of by
66
Dutt et al. (1972) in the USA and Beek and Frissel (1973) in Netherlands. Since
then many N models have been developed for varying objectives with different
levels of complexity and accuracy. Nitrogen models vary from purely research
oriented to management and screening oriented tools (Shaffer et al., 2001).
The development of various N models with differing levels of complexities has
made significant contribution in understanding the nitrogen cycle in the crop-soil
system. Nitrogen models have been developed to simulate N cycle in the soilcrop system under irrigated conditions, dryland conditions or both. At the same
time, some N models are crop specific while others are generic and could
simulate crop growth and N cycle under crop rotation.
This project required an N model that could simulate various irrigation
management practices i.e. irrigating the profile to field capacity, leaving room for
rain, or irrigating with a leaching fraction, all of which are critical for long-term
scenarios. The reason for this is that N management, salt management and
irrigation are inextricably linked. There is also a need to explore trade-offs
between nutrient management, optimal yields and pollution such as through
leaching. In addition there was no generic model available with the option for
perennial grass management practices such as cutting hay or forage at a
selected growth stage or day interval.
67
For the reasons above, it was decided to add a nitrogen module to the SWB
model - a crop growth / irrigation scheduling model that has already been
validated for various crops and grasses under irrigation and dryland conditions.
The N simulation approaches and algorithms were obtained primarily from
CropSyst (Cropping Systems Simulation Model) (Stöckle et al., 2003), since
CropSyst and SWB have similar backgrounds and approaches, having both
grown out of work done by Prof. Gaylon Campbell from Washington State
University.
Generally, mechanistic models provide a better understanding of the nitrogen
cycle by mechanistically describing the major processes involved in the N cycle
such as volatilization, mineralization, immobilization, nitrification, denitrification,
crop uptake, and leaching. The main difference between models is thus the
difference in the degree of sophistication of these major processes (Frissel and
Van Veen, 1982). Some of the models are used for short term seasonal soil N
monitoring while others for predicting long-term N dynamics of decades and
longer.
Shaffer et al. (2001) have detailed the differences in the major nitrogen
transformation processes between various models. They also elaborate about
similarities between models regarding the major nitrogen transformation
processes: mineralization, immobilization, nitrification, and volatilization. Most
nitrogen transformation processes in the soil profile including mineralization,
68
nitrification, and denitrification are quantified using first order kinetics (Ma and
Shaffer, 2001). The rate constants are then modified based on the soil
temperature, water content, and other factors.
2.3.1 Mineralization
The mineralization of organic substances and residues is greatly influenced by
the C:N ratio of the material which influences the net mineralization or
immobilization rate (Vinten and Smith, 1993). Carbon to nitrogen ratio (C:N ratio)
is an approximation of the important parameter energy:nitrogen (E:N) ratio. Low
C:N ratio organic residues have faster mineralization rates and result in net
mineralization. According to Haynes (1986a), the lowest C:N ratio is 8:1 and is
found in microbial biomass. The C:N ratio of clover, beans and lucerne is in the
range of 13:1 to 23:1 where as cereal straw and other mature plant stalks have
60:1 to 80:1 C:N ratios. The C:N ratios of some plant materials with N free lignin
and residual substances from many peat soils with high C:N ratios are poor
sources of energy for most micro organisms (Jansson and Persson, 1982).
Various organic materials do have variable decomposition rates depending on
their resistance to microbial activity. Consequently many models simulate this
difference by classifying the organic matter into pools. This classification, as
explained by Ma and Shaffer (2001), is based on physical (mobile or immobile
carbon and nitrogen), chemical (organic with different C:N ratios and
69
compositions or inorganic) and biological characteristics (capable of transforming
to other forms by various types of microbes at different rates).
Most multiple pool N models include soil microbial biomass pool but only the
RZWQM model simulates the growth and death of microbial populations (Ma and
Shaffer, 2001). Models differ in the approach they follow in simulating organic
matter decomposition. For instance models such as NTRM use regression
equation. On the other hand, models such as RZWQM, CropSyst, NLEAP,
CERES, EPIC, GLEAMS, LEACHM, CENTURY, SOILN, ANIMO, DAISY,
SUNDIAL, and CANDY use first order kinetics. Other models including NCSOIL
use Monod kinetics (Ma and Shaffer, 2001).
Decomposition rates for most of the models such as NLEAP, CROPSYST,
RZWQM, CERES, EPIC, GLEAMS, CENTURY, NCSOIL, LEACHM, SOILN,
ANIMO, DAISY, SUNDIAL, and CANDY are adjusted after the soil temperature
and soil water content. Some of these models consider additional factors which
do affect organic decomposition rates. These factors include soil pH, soil ionic
strength, soil microbial population, C:N ratio, bulk density, soil texture, and lignin
content. Details of the mineralization functions used by various models and the
additional factors used by each model could be found from Ma and Shaffer
(2001) and McGechan and Wu (2001).
70
The N model in SWB which is adopted from Cropsyst partitions organic residue
into three pools (fast, slow and lignified). The model simulates carbon
decomposition from each pool using Eq. 1. Net mineralization is then computed
using Eq. 2. The model, however, estimates soil organic matter decomposition
separately using Eq. 3.
Carbon decomposition = Cres * CF _* (1 - exp(-k * TF)) * WF
[1]
Where
Cres – is carbon mass in residue or manure
CF – is residue or manure contact fraction (from literature)
k–
is residue or manure decomposition constant
TF – is soil temperature function (computed using Eq. 4)
WF – is soil water function (computed using Eqs. 5).
Net N mineralization = (1 / CN ratiodecomp – Ctrans/CNratiopool) * Cdecomposition
[2]
Where
CN ratiodecomp – is CN ratio of the decomposing pool
Ctrans – is carbon fraction transferred from the decomposing pool to another pool.
CNratiopool – is the CN ration of the receiving pool.
Carbon decomposition = C mass * min (TF, WF) * Tillagefactor * k
71
[3]
Where
C mass - is pool carbon mass
min (TF, WF) - is the minimum of the two (soil temperature and water functions)
Tillagefactor - is tillage decomposition rate adjustment factor (computed by the
model (Stöckle et al., 2003) )
k - is carbon decomposition constant.
Temperature and water functions are estimated by the model using Eqs. 4 and 5.
TF = ((T – Tmin) ^ (Q) * (Tmax - T)) / ((Topt – Tmin) ^ (Q) * (Tmax – Topt))
[4]
Where
T - is real time soil temperature
Tmin – minimum temperature below which there is no microbial activity (-5 oC)
Tmax – maximum temperature above which there is no microbial activity (50 oC)
Topt – optimum temperature for microbial activity (35 oC)
Q – is constant (2) estimated using ((Tmin - Topt) / (Topt - Tmax))
Water function (WF) is computed differently for different soil water filled porosity
ranges. The model partitions the water filled porosity (WFP) into three groups
and computes the WF for each group separately as presented in Eqs. 5 - 7.
72
If 0.1 ≤ WFP < 0.5 WF = ((WFP – WFPmin) / (WFPlow – WFPmin))
If 0.5 ≤ WFP ≤ 0.7 WF = 1
If 0.7 ≤ WFP ≤ 1
WF = WFsat+ (1- WFsat) * ((1 - WFP) / (1-WFPhigh)) ^ 2
(5)
(6)
(7)
Where
WFP – is the water filled porosity
WFPmin – is the least amount of water filled porosity (0.1)
WFPlow – is the low water filled porosity (0.5)
WFPhigh - is the highest level of water filled porosity (0.7)
WFsat – is moisture function value at saturation (0.6)
2.3.2 Immobilization
Immobilization is the net incorporation of mineral nitrogen, mainly NH4+, into
organic forms (microbial tissue) during the decomposition process (Vinten and
Smith, 1993). It is a process which works in opposite direction to mineralization.
The decomposition of residues with high C:N ratio results in a negative net
residue nitrogen mineralization or immobilization. This, however, changes in due
time with the growth and stabilization of the microbial population resulting in a
decline of the C:N ratio and the release of mineralizable nitrogen. Most models
set a C:N ratio beyond which immobilization takes place. This ratio is set to 24 by
CropSyst model.
73
A periodic decline in the amount of plant available nitrogen is observed after the
incorporation and decomposition of cereal straw and other mature plant stalks
due to immobilization (Powlson et al., 1987). Nevertheless, mineralization is
improved with time after the continuous incorporation of organic residues.
Nitrogen immobilization is simulated using residue C:N ratio (NTRM and NLEAP),
decomposition rates (CERES and EPIC), or both decomposition rate and C:N
ratios (SOILN, ANIMO, DAISY, SUNDIAL, CropSyst, and CANDY). During
immobilization, some models such as NCSOIL, NLEAP, and CROPSYST take up
inorganic N from the NH4 and NO3 pools, where as other models such as EPIC
immobilize only NO3 (Ma and Shaffer, 2001).
Cropsyst model estimates immobilization only when the net mineralization is less
than zero using Eq. 8.
Nitrogen immobilization = -Nitrogen net mineralization
[8]
2.3.3 Nitrification
Nitrification is the oxidation of ammonium ion or ammonia to nitrate or nitrite ion
(Delwiche, 1981). Howard-Williams and Downes (1993) classified the processes
responsible for nitrification into 2: namely autotrophic and heterotrophic
nitrification. Autotrophic nitrification is the process of transforming ammonium or
ammonia to nitrite by a group of bacteria called Nitrosomonas sp. and
74
Nitrosolobus sp. followed by the transformation of nitrite to nitrate by a separate
group of bacteria called Nitrobacter sp. Whereas, heterotrophic nitrification is the
oxidation of reduced organic nitrogen compounds to oxidised nitrogen species.
The nitrification process is described by three types of kinetics: Monod, firstorder, and zero order models (Hansen et al., 1995). Models such as RZWQM,
EPIC, LEACHM, SOILN, SUNDIAL and CROPSYST follow first-order kinetics to
simulate nitrification. However, models such as GLEAMS, NLEAP and NCSOIL
use zero-order kinetics. On the other hand, NTRM uses regression equation to
estimate the rate of nitrification (Shaffer, 1985) whereas CERES, DAISY, and
CANDY models follow the Michaelis-Menten equation (Godwin and Singh, 1998;
Godwin and Jones, 1991).
Factors which influence the rate of nitrification include soil water, soil
temperature, soil pH. Models such as NLEAP, RZWQM, CERES, EPIC,
GLEAMS, NCSWAP, LEACHM, SOILN, and SUNDIAL consider the impact of
soil water and temperature on the rate of nitrification (Shaffer et al., 1991; MA et
al., 2001; Godwin and Singh, 1998; Williams, 1995; Knisel, 1993; Molina et al.,
1983; Hutson, 2000; Eckersten et al., 1998; Bradbury et al., 1993; Stöckle and
Nelson, 2000). RZWQM, CERES, EPIC, SOILN, and Cropsyst consider one
additional factor, soil pH. Models such as RZWQM consider more factors which
could impact the rate of nitrification such as oxygen concentration, and
population of autotrophs (Ma et al., 2001).
75
Cropsyst simulates nitrification only in the presence of ammonium in the soil
profile. In cases where there is ammonium, the model further checks for the
nitrate to ammonium ratio of the soil. If the soil nitrate to ammonium ratio is less
than 8, the model estimates nitrification as zero. However, if the nitrate to
ammonium ratio exceeds 8, the model uses Eq. 9 to estimate the amount of
nitrification.
Nnitrified = (NH4 – NO3 / 8) * (1 - exp(-0.2 * pHF * TF))) * MF
(9)
Where
Nnitrified – is the mass of nitrified N
NH4
– is the mass of ammonium N
NO3 – is the mass of nitrate N
TF – is the soil temperature function (Eq. 4)
MF – is the soil water function (Eq. 5 – 6 and Eq. 10)
pHF – is the soil pH function (estimated using Eq. 11)
If 0.7 ≤ WFP ≤ 1
WF = 1 * ((1 - WFP) / (0.3)) ^ 2
pHF = (pH - pHmin) / (pHmax - pHmin)
where
pHF - is pH function
76
(10)
(11)
pH – is the measured soil pH
pHmax – is the maximum temperature for nitrification (6.5)
pHmin – is the minimum temperature for nitrification (3.5)
2.3.4 Denitrification
According to Ma and Shaffer (2001), the simulation of denitrification process by
models is empirical due to both the unknown nature of the process and the
spatial and temporal variability of anaerobic conditions in the soil. Similar to
mineralization and nitrification, models describe the denitrification process as
Michaelis-Menten, zero order, or first order kinetics (Hansen et al., 1995).
Denitrification is simulated as zero order by SOILN and first order by models
such as NLEAP, RZWQM, CERES, EPIC, GLEAMS (Hansen et al., 1995), and
Cropsyst. LEACHM model simulates denitrification using Michaelis-Menten
kinetics (Hutson, 2000). Models such as NTRM, however, does not simulate
denitrification (Ma and Shaffer, 2001).
The effects of both soil temperature and soil water on the rate of denitrification
are considered by models such as NLEAP, RZWQM, CERES, EPIC, GLEAMS,
LEACHM, SOILN, CANDY, and Cropsyst (Ma and Shaffer, 2001; Stöckle and
Nelson, 2000). Some of the models consider additional factors such as water
extractable soil carbon content (CERES), percentage of soil organic carbon
77
content (EPIC and GLEAMS), and soil pH, soil carbon substrate, and population
of denitrifiers (RZWQM) (Ma and Shaffer, 2001).
When estimating denitrification, Cropsyst first estimates respiration response
function (RRF) (algorithm too long to present) based on the soil temperature
function (TF) Eq. 4. The model then presents forward the following two
preconditions which should be fulfilled, if denitrification is to take place. These
preconditions are: primarily there must be nitrate_N in the soil profile, secondly
the denitrification soil water function (DWF) (Eqs. 12-14) must exceed zero.
Finally, the denitrified N mass is computed using Eq. 15.
If WFP < WFPNR
DWF = 0
(12)
If WFPNR < WFP < WFPinfl
DWF = 0.9 - 0.9 * (WFPinfl - WFP) / (WFPinfl - WFPNR)
(13)
If WFPinfl < WFP
DWF = 1 - 0.1 * (1 - WFP) / (1 - WFPinfl)
(14)
Where
WFP – water filled porosity
WFPNR – water filled porosity without response to denitrification (0.6)
DWF – denitrification soil water function
WFPinf – water filled porosity where inflection is observed in denitrification (0.8)
Ndenitrified = k * soilmass * minimum (DWF, RRF, (NO3-dry/(NO3-dry+c))
Where
78
(15)
Ndenitrified – denitrified N mass
k – potential denitrification constant (0.000032)
soilmass – dry soil mass of the profile (layer)
RRF – respiratory response function
NO3-dry – nitrate concentration of the dry soil
c – denitrification half rate (0.00006 kg N / kg Soil)
2.3.5 Ammonia volatilization
The sources of NH3 for volatilization include inorganic fertilizers applied to the soil
in the form of NH4+ compounds and the decomposition of organic nitrogenous
sources. The rate of ammonia volatilization is affected by factors such as soil pH,
soil temperature, soil CEC, soil water content, wind speed, fertilizer application
method (surface, incorporation), and pH of fertilizer (Nelson, 1982; Freney et al.,
1983). Ammonia volatilization increases with increase in soil and fertilizer pH,
wind speed, and soil temperature (until 45oC) (Nelson, 1982).
Ammonia volatilization is simulated by most models as first order kinetics (Ma
and Shaffer, 2001). Some models such as NTRM and NCSOIL, however, do not
simulate ammonia volatilization (Ma and Shaffer, 2001). Different models
consider different factors to adjust ammonia volatilization rate. For instance,
NLEAP, EPIC, GLEAMS, Cropsyst, and RZWQM use soil temperature. Some of
these models such as NLEAP (Shaffer et al., 1991) consider additional factors
79
such as residue cover, soil cation exchange capacity, and fertilizer application
method (surface application or incorporation). RZQWM and Cropsyst consider
additional factors such as wind speed and ammonium pressure gradient between
soil and air.
Ammonia volatilization could take place from the inorganic fertilizer, biosolids,
plant residue, or soil ammonium (McGechan and Wu, 2001). Most models (Ma
and Shaffer, 2001) including Cropsyst simulate volatilization from the top soil
layer. Similar to most N models, Cropsyst does not estimate ammonia
volatilization from soil ammonium, because generally soils and crops are not
considered as the major source of NH3 (Bussink and Oenema, 1998). Ammonia
loss from crops is estimated to be below 2 kg NH3-N ha-1 yr-1 (Holtan-Hartwig and
Bøckman, 1994).
Cropsyst simulates ammonia volatilization from inorganic fertilizer (liquid or solid)
and organic fertilizers (liquid or solid). Cropsyst primarily estimates ammonia N
mass available for volatilization (NH4-AV) as a function of soil-CEC and
fertilization application method (broadcast or incorporated) before estimating
ammonia volatilization. The model estimates ammonia volatilization only if the
NH4-AV is greater than zero. If the NH4-AV is greater than zero, Cropsyst
estimates ammonia volatilization using Eq. 16.
NH3-volatilized = ((KH * NH3-conc)/patm) * TTC
80
(16)
Where
NH3-volatilized – is volatilized ammonia mass
KH – is Henrys constant (estimated according to the air temperature)
NH3-conc – is ammonium concentration
patm – is atmospheric pressure
TTC – is turbulent transfer coefficient
2.3.6 Crop nitrogen uptake
Nitrogen is mostly taken up by plants in the form of NO3 during a convective flow
of soil water to plants in response to transpiration from the canopy (Olson and
Kurtz, 1982). This is mainly because of the low attraction of NO3- to the soil
colloid compared with NH4+ which has a higher force of attraction to the soil
colloids. This low force of attraction between the NO3- and soil colloids makes it
easier for the NO3- to be carried by mass flow to the plant roots. In addition, the
rapid nitrification of NH4+ in most soil conditions play a significant role in
minimizing its availability for crop uptake (Haynes, 1986b). Nevertheless, NH4+
could be the major source of N to crops under soil conditions which are not
favourable for nitrification such as anaerobic and acidic soil conditions (Haynes,
1986b).
Crop N uptake simulation in most models is driven by plant N demand (Ma and
Shaffer, 2001). These authors stressed that crop N uptake depends on the
81
availability of NH4+ and NO3-, transpiration rate, and diffusion of soil N to root
surface. Most models estimate crop N uptake as passive uptake (estimated from
the rate of transpiration). Models such as RZWQM, however, consider active N
uptake using Michaelis-Menten kinetics if N supply through passive uptake is not
sufficient to satisfy the crop demand (Hanson, 2000). The N model added to
SWB was adopted to simulate crop N uptake as a passive and active processes.
Crop nitrogen uptake is estimated as the minimum value between crop demand
and soil supply by the following models CERES, EPIC, GLEAMS, and CropSyst
(Godwin and Jones, 1991; Williams, 1995; Knisel et al., 1993; Stöckle and
Nelson, 2000).
2.4 Motivation for this study
Previous studies conducted in South Africa were laboratory incubation trials and
single year field scale studies on specific crops (oats and maize). The majority of
studies relating to beneficial sludge use on agricultural lands have been
conducted in the northern hemisphere, under different soil-climate combinations
than those experienced in South Africa. Since processes which determine the
availability of nitrogen from sewage sludge (mineralization, immobilization,
volatilization, nitrification, and denitrification)
are influenced by climatic
(temperature and water availability) and edaphic factors (soil pH, CEC, texture),
locally based trials are essential.
82
This study was initiated in 2005 while the 1997 South African sewage sludge
guideline was in place (Water Research Commission, 1997). This sludge guideline
had an upper limit of 8 Mg ha-1 yr-1 and was recently increased to an upper limit of
10 Mg ha-1 yr-1, based on short-term laboratory incubation studies (Snyman and
Van der Waals, 2004). According to the current guideline, farming systems which
can responsibly exceed this upper limit norm can be approved upon successful
motivation.
Sewage sludge treatment and handling processes affect the characteristics and
availability of nutrients from sludge. Demand for nutrients by crops is dependent
on crop type, water availability, climate, soil type, management practice, and
farming intensity. The vast number of combinations among the above factors
could never be adequately covered by field trials. The strategy in this project was
therefore to use four widely differing farming systems in order to develop and test
a mechanistic decision support model.
This study investigates four major cropping systems as presented below.
1. Dryland maize (Zea mays L.)
Maize is the staple food for the majority of the South African population and
accounts for 51% of the cultivated land (FAO, 2005). About 8.0 million tons are
produced each year, of which half is white maize for human consumption and
half yellow maize for animal feed (Du Plessis, 2003; Vermeulen, 2005). The
83
rationale for selecting maize as a test crop was because it is the most commonly
grown crop around the study area.
According to the Fertilizer Society of South Africa (2000), the mean N removal by
maize accounts to 15 kg N per ton of grain only or 27 kg N per ton of grain plus
stover. This is similar to the estimation made by O’Gara (2007) who reported that
the total N removal by a whole maize plant is 1.6 times the amount taken off in
the grain. The mean export of P by one ton maize grain was estimated as 3 kg
(Fertilizer Society of South Africa, 2000).
Rainfall is the limiting factor for the productivity of dryland maize, making it
difficult to predict the amount of sludge required. There is likely to be insufficient
N for maximum yields in wet seasons and an excess of N in soil following dry
years. The latter situation would predispose the system to nitrate leaching in the
beginning of the new season, before a deep root system had developed.
2. Irrigated maize-oats rotation (Zea mays L. and Avena sativa L.)
Access to irrigation makes it easier to predict crop yields and hence the
appropriate sludge application rates. Irrigation also allows for double cropping,
and hence the potential to remove more nutrients from the soil. At the same time
the continually wet conditions are more conducive to leaching, should high levels
of N build up in the soil.
84
A substantial proportion of the maize crop is irrigated: 36% in the Free State,
20% in Mpumalanga, and 28% in North West is irrigated (FAO, 2005). Oats can
be grown between two maize crops. Oats are promoted as a healthy grain for
human beings (Peterson, 1992) and a high quality fodder for animals (Schrickel
et al., 1992). According to the Fertilizer Society of South Africa (2000), the mean
N removal by oats grain alone accounts to 33 kg per ton of grain. The mean
export of P by one ton of oats grain was estimated as 4 kg (Fertilizer Society of
South Africa, 2000).
3. Perennial dryland pasture (Eragrostis curvula (Schrad.) Nees)
Dryland pastures occupy large areas, need high amounts of fertilizer for optimal
growth and are common around built up areas (Muchovej and Rechcigl, 1998).
Perennial grasses have the potential to reduce nitrate leaching compared to
annual crops due to their established root system. They are considered a good
choice for repeated sludge applications because of their efficient nitrogen
utilization under intensive management practices and because a number of
harvests can be made in a year (Cogger et al., 2001).
Weeping love grass is a summer growing perennial, which is native to South
Africa (Skerman and Riveros, 2008) and the most widely cultivated grass in
South Africa (Dickinson et al., 2004). Weeping love grass is used for animal feed,
either grazed or as hay. In the drier farming regions it is recommended for leys
with Alfalfa (Duke, 1983). Besides animal feed, weeping love grass is used to
85
stabilize terraces and roads and control wind and water erosion (Duke, 1983;
Cook et al., 2005).
Generally biomass production is poor under conditions of low rainfall and soil
fertility. However, under favourable environmental conditions and adequate N
and water supply, biomass yield could increase to 20 - 30 Mg ha-1 (Cook et al.,
2005). Nitrogen is the key element to a high quality fodder production. Nitrogen
application rate depends on the quantity and quality of fodder needed, rainfall of
the region, and soil production potential. As a rule of thumb weeping love grass
requires 20 kg of N and 2 kg of P for every ton of dry matter produced (Dickinson
et al., 2004).
4. Lawn sod production (Pennisetum clandestinum)
The nutrient content of sludge produced by municipal water treatment works
often exceeds the requirements of nearby crops. Transporting sludge further
afield is not always economically viable. In such situations, excess sludge could
be exported through turfgrass sod production. Turfgrass sod production provides
the opportunity to export large volumes of sludge from limited areas because a
layer of soil is removed with the harvested turf. The sludge removed along with
the sod has the added advantage of being a slow-release nutrient source for the
grass at the final establishment location (Muse et al., 1991). In addition, sludge
application has the potential to minimize soil loss from the turfgrass production
86
site, by minimizing the thickness of soil exported with the sod (Charbonneau,
2003).
Kikuyu is a warm season perennial grass, which was introduced to South Africa
80 years ago. It is an important pasture grass in the high rainfall areas and under
irrigation (Dickinson, 2004), due to its good nutrition values (Butler and Bailey
1973; Marais et al., 1992). It has the capability to regenerate easily following
repeated mowing and thus is used as lawn. Kikuyu is a low maintenance grass
and is able to recover quickly from moderate to severe wear injury (American
lawns, 2008). The rationale for the selection of kikuyu as a test crop for sod
production was because it is the main turfgrass species in South Africa. In
addition, it has the ability to grow under acidic soils, establishes easily in soils
rich with organic matter, is salt tolerant (Dickinson, 2004), and has the ability to
take up large amounts of N (FAO, 2008) and P (Hanna et al. 2004).
According to Williams (quoted by Dickinson, 2004), kikuyu requires about 23 kg
of N and 2.9 kg of P per ton of dry matter produced. Generally, kikuyu is less
sensitive to P deficiency except for extremely deficient soils (Miles, 1997;
Mathews et al. 2001). In this study, however, the main purpose of sod production
was to export sludge with sod as a growing medium.
87
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