Applied Soil Ecology 25 (2004) 1–18
Review
Transport of bacteria from manure and protection
of water resources
Adrian Unc a,∗ , Michael J. Goss b,1
a
Centre for Research on Environmental Microbiology, University of Ottawa, Ottawa, Ont., Canada K1H 8M5
b Land and Water Stewardship, University of Guelph, Richards Building, Guelph, Ont., Canada N1G 2W1
Accepted 13 August 2003
Abstract
Survival and transport of pathogens from manure in the environment depend on a number of complex phenomena. An
important question is how the properties of such a complex environment as the soil–manure medium impact the persistence of
bacteria within the vadose zone. First, manure can change the partitioning of precipitation water between infiltration (enhanced
by solid manure) and surface runoff (stimulated by liquid manure). Components of manure, such as straw and coarse organic
matter, can strain and filter micro-organisms from the transporting water. After infiltrating the soil, the retention of bacteria
depends on the physical configuration of soil, the soil chemistry, and the properties of the microbial cells. Transport of bacteria
in soils obeys the general laws pertinent to macropore flow and the interaction between particles and surfaces of variable
charge. Detailed characterisation of the variable properties within the structured soil profile is a difficult task. Application of
manure can result in significant changes in the physical and electrochemical properties of the soils and microbial cells. Such
changes can affect the interaction between bacterial cells and soils in several ways: increase filtration, modify the kinetics of
the physico-chemical interactions between charged surfaces, and alter the competition for retention sites between suspended
soluble and particulate compounds. Survival of faecal bacteria is affected by the physical and chemical conditions existing
prior to manure application as well as by conditions imposed by mixing soil and manure. Competitive interaction with native
soil bacteria, in the soil–manure mixtures, is an important aspect governing survival of introduced organisms.
© 2003 Elsevier B.V. All rights reserved.
Keywords: Soil; Manure; Bacterial persistence; Surface chemistry
1. Introduction
In the Walkerton (Ontario, Canada) tragedy of May
2000, contamination of the municipal water system
with Escherichia coli O157:H7 and Campylobacter
jejuni resulted in 2300 people requiring medical attention, 7 of whom died. This is a community that
∗ Corresponding author. Tel.: +1-613-562-5800x8568;
fax: +1-613-562-5452.
E-mail addresses: [email protected] (A. Unc),
[email protected] (M.J. Goss).
1 Tel.: +1-519-824-4120x2491; fax: +1-519-824-5730.
has a population of only about 5000 people. The total
cost attributable to the Walkerton water crisis was estimated at US$ 64.5 million (Livernois, 2002). Investigation into the causes of the microbial contamination
of the municipal well water, the common source for
those affected, indicated that the most likely cause was
transport of manure bacteria to the aquifer by infiltrating water, although direct entry of surface runoff into
the well could not be ruled out (O’Connor, 2002). This
tragic event showed that even under current best manure management practices, contamination of water
resources with micro-organisms can occur and have a
0929-1393/$ – see front matter © 2003 Elsevier B.V. All rights reserved.
doi:10.1016/j.apsoil.2003.08.007
2
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
catastrophic impact on public health if other barriers
to the spread of disease in piped water supplies are
inadequately enforced.
Reports have indicated greater occurrence of contamination of ground water with nitrate in areas where
animal manure is applied regularly. However, the overall impact of agricultural land use practices on regional
ground water quality is not well understood (Goss
et al., 1998). The Ontario Farm Groundwater Quality
Survey found that the proportion of wells contaminated with faecal bacteria was significantly greater on
farms where manure was spread than where only mineral fertiliser was used (Goss et al., 1998), pointing to
manure use on farms as a significant source of well
water contamination for both nitrate and bacteria. The
bacterial groups present in greatest number in manure
are faecal coliforms and streptococci, with Salmonella
spp. also being found occasionally (Table 1). E. coli
O157:H7 has been reported in a large variety of farm
and wild animal species, but the main concern is associated with ruminants. The incidence of detection
was greater for farm animals with the greatest incidence in cattle. However, there are commonly about
1010 bacteria (g−1 dry manure), whereas, even when
present pathogenic forms rarely exceed 105 colony
forming units (CFU) g−1 . With such a predominance
of non-pathogenic organisms, their presence in an environmental water sample can be used as an indicator
that the source has been impacted by faecal material,
and so is likely contaminated with pathogens.
The concentration of a given micro-organisms in
manure applied to soil is an important parameter determining the potential for contamination of water resources. The greater the concentration, the more likely
it is that some will be transported (Goss et al., 2002).
The type and number of micro-organisms in manure
can vary with the animal species, age of animals, the
type of bedding used, the method of storage (liquid or
solid), and the storage period (Lachica, 1990; Nodar
et al., 1992). Given the many possibilities for handling
manure on farms, the opportunity to modify the source
strength of the pathogens as well as the form of the material that may gain access to the wider environment,
guidelines for farmers are important for limiting the
negative impact of agricultural activities on the water
resources. However, appropriate recommendations for
manure management to protect water resources from
pathogens cannot be formulated without a detailed
understanding of the factors affecting the survival and
transport of micro-organisms from manure between
source locations and surface or groundwater bodies.
Studies on the transport of faecal bacteria from manure through soil have concentrated on their potential
to travel through soil and reach an aquifer or enter
drainage tiles eventually moving to surface water
courses (e.g. Culley and Phillips, 1982; MacLean
et al., 1983; Patni et al., 1985; Shreshta et al., 1997;
Conboy, 1998; Joy et al., 1998), or the evaluation of
management procedures that influence the contaminant potential of manure micro-organisms, including
research on the changes in bacterial numbers during
manure storage and handling (Sutton, 1983; Lachica,
1990; Nodar et al., 1992; McMurry et al., 1998), the
survival of faecal bacteria in the environment (e.g.
Fenlon, 2000), the accidental contaminant potential
due to point sources of contamination, and the impact
of the soil management on the contaminant potential
of manure bacteria (e.g. Cook et al., 1997; Stoddard
et al., 1998). On the other hand, detailed investigations of the impact of cell properties on transport
have been performed mostly under conditions of reduced ionic strength and low colloidal concentrations
(Marshall, 1985; Stenström, 1989; Gannon et al.,
1991a,b; Huysman and Verstraete, 1993).
Although the contaminant potential of manure bacteria is recognised and the properties that are important
for the transport of these micro-organisms through soil
have been identified (Gerba and Bitton, 1984; Stotzky,
1985; Abu-Ashour et al., 1994; Mawdsley et al., 1995),
there is a lack of specific information on how manure
influences the expression of these properties. Furthermore, there is little information on how the variable
expression of bacterial cell surface properties affects
retention within soil and manured soil. These are important questions for quantifying the risks associated
with faecal pathogens since they originate in manure
and in many cases enter the wider environment when
applied to agricultural fields with the manure. The
long-term effects of addition of manure on the increases in soil organic matter content and therefore on
the soil structural stability are well known (e.g. Tester,
1990). While these effects are also important for the
transport of manure bacteria through soil there is no
information available on the short-term impact of different manure types on the soil properties relevant to
bacterial transport.
Table 1
Examples of bacterial numbers in some animal manure (CFU ml−1 or g−1 fresh manure)
Manure type
Salmonella spp.
×
to 1.3 ×
× 103
× 104 to 1.1 × 106
coli)
9.3 × 103
7.2 × 104 to 4.5 × 105
(streptococci-D)
0
0–1.5 × 103
(S. infantis)
Liquid cattle manure
2.4 × 103
4.5 × 102 to 1.5 × 106
(E. coli)
9.3 × 103
4.5 × 102 to 9.5 × 105
(streptococci-D)
0
0
Dairy slurry
6.3 × 104 to 1.0 × 107
(enterobacteria) 4.7 ×
107
2.7 × 107
Solid beef manure
2.4 × 105
1.9 × 106 to 6.8 × 106
1.5 × 107
Solid dairy manure
2.0 × 105 to 1.0 × 107
(enterobacteria)
Fresh cow manure
Up to 1.0 × 109
Sheep
Horse
Poultry
4.3
2.4
9.5
(E.
103
Protozoa and others
105
6.0 × 106
9.4 × 104
1.3 × 106 to 1.4 × 108
Source
Unc (1999)
Weigel (1995)
Rüprich (1994)
Weigel (1995)
Rüprich (1994)
Östling and Lindgreen
(1991), and Crane et al.
(1983)
0
Weigel (1995)
Unc (1999)
Östling and Lindgreen
(1991)
Up to 1.0 × 109
6.6 × 105
6.3 × 106
6.2 × 105 to 1.9.7 × 108
From 25 to 1.8 × 104 in healthy
animals to 1 × 109 in sick animals
(Cryptosporidium parvum)
Mawdsley et al. (1995)
Clinton et al. (1979)
Scott et al. (1994), and
Smith (1992)
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
Faecal streptococci
Liquid swine manure
Faecal coliforms
Crane et al. (1983)
Crane et al. (1983)
Crane et al. (1983)
3
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A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
The purpose of this review is to indicate recent developments in understanding microbial transport, especially that through the vadose zone, and to highlight
the research gaps that limit the advice available on
manure management to ensure the protection of water
resources.
2. Survival of micro-organisms from manure in
soil
Survival rate of micro-organisms introduced into the
soil is an important factor influencing their potential to
contaminate water resources. The survival conditions
for the enteric bacteria once voided from the animal
organisms are considered unfavourable. Nevertheless,
some can survive for extended periods, at least up
to sixty days (Fenlon et al., 2000), or in what are
considered least hospitable environments such as on
fabrics and plastics (Neely, 2000; Robine et al., 2000).
Antibiotic resistant strains of E. coli and Streptococcus
faecalis were found to persist in high numbers over a
period of at least 32 days in saturated soil conditions
(Hagedorn et al., 1978). Recent research shows that
E. coli and Enteroccocus spp. from pig manure may
survive in soil for even longer periods—40–68 days
after application (Shreshta et al., 1997; Cools, 2001).
The survival rate of bacteria in soil depends on the
source, microbial species and on the manure application method (Rüprich, 1994). During the storage of liquid manure the population of viable organisms initially
declines rapidly only to regain numbers later, in one
example up to five-fold the initial value after 14 weeks
(Nodar et al., 1992). There are gradients of temperature within a completed solid manure pile, which are
the result of different rates and types of organic matter
digestion (from aerobic at the periphery, to more anaerobic toward the centre of the pile). Micro-organisms
have different rates of survival in these zones. The
ones near the periphery have more chances to survive
and form sources of contamination (Sutton, 1983).
The amount of pathogenic bacteria is likely to be
greater in swine and poultry manure than in cattle
manure. Due to the greater mobility of bacteria in
the liquid phase compared to the solid phase, liquid
manure tends to be more uniformly contaminated than
does solid manure. Pathogens may be found in manure
even if the animals present no symptoms (Goss et al.,
2002). While manure samples from individual animals
may not be contaminated, a supply of manure from
storage can contain pathogens that have been shed by
a restricted number of animals (Busato et al., 1999;
Strauch, 1988), thus the greater the number of animals
on a farm the greater the chances for pathogens to be
found in manure.
Methods for the land application of manure depend on the manure type (liquid or solid), soil type,
and crop type (Goss et al., 2002). When manure is
injected into soil, micro-organisms are less likely to
be destroyed by the ultraviolet solar radiation than
if the manure remains on the soil surface. On the
other hand, incorporation by injection increases the
possibility for micro-organisms to be adsorbed onto
soil particles (Patni et al., 1985). The normal tillage
used in farming operations can influence the conditions that microbes from manure experience after
application. For example, biological activity near the
soil surface is greater in no-till than in conventional
tillage, which results in better conditions for survival
of the indigenous soil micro-organisms (Levanon
et al., 1994). Addition of manure results in increased
availability of carbon substrates and mineral nutrients
that in turn boost the soil biological activity. Significant mineral nutrients in manure include ammonium
ions, phosphate, potassium, sodium, magnesium and
calcium; metals such as zinc and copper. The level of
available nutrients in the soil modifies survival rate of
bacteria (Rattray et al., 1992). Tenuta and Lazarovits
(2002) found that ammonia, nitrous acid and fatty
acid toxicity following application of nitrogen rich
organic amendments can kill microsclerotia of plant
pathogens present in soil. Thus, the high level of
aqueous ammonia in liquid manure may be toxic to
microbial populations. Survival of faecal coliforms
is greatly extended in organic soil compared with
that in mineral soils. This might have to do with
the higher water-holding capacity level of these soils
(Gerba and Bitton, 1984). However, laboratory tests
performed by Cuthbert et al. (1950), showed that E.
coli and faecal streptococci survive several weeks in
limestone (pH 5.8–7.8) while dying in a few days in
a peat soil (pH 2.9–4.5), indicating the importance
of pH as well as water content. Bacteria are considered to be more likely to survive a longer period in
soils with high water-holding capacity (Gerba and
Bitton, 1984). Low matric potentials—large negative
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
values—are associated with dry conditions and reduced viability of bacterial cells in soil (Sjogren,
1994). In the soil, cool temperatures favour faecal
bacteria survival. Stress induced by starvation, low
temperatures, and low levels of available water can
reduce the microbial metabolisms, phenomenon observed as viable but not culturable bacteria (Colwell,
1993; Chmielewsky and Frank, 1995; Bogosian et al.,
1996). This means that although bacteria are viable,
and potentially pathogenic, they cannot be readily
grown on identification substrata. However, even partial changes in the environment, such as increased
temperature, can lead to reactivation and accelerated
multiplication of such cells (Jiang and Chai, 1996).
The increase in nutrients engendered by the addition of manure may also result in an increase of the
predatory population. Competition between soil organisms has been found to be a major factor in the
reduction of the bacterial populations introduced in
soils (Acea et al., 1988; Recorbet et al., 1992; Soda
et al., 1998). Murry and Hinckley (1992) found that
the number of Salmonella enteritidis is more limited in
the presence of earthworms (Eisenia foetida)—8% reduction versus only 2% reduction without earthworms.
Normal soil bacterial flora was also reduced in the
presence of earthworms. Other soil micro-organisms
such as protozoa, nematodes and Bdellovibrio—a soil
bacterium—prey on soil bacteria and implicitly on the
ones introduced with manure (Goss et al., 1996). Aggregation of cells triggered by various stress factors,
and nutrient gradients can increase bacterial resistance
to unfavourable conditions in the environment (Bossier
and Verstraete, 1996). The survival of faecal bacteria
can extend over long periods after manure application
and once bacteria reach the ground water the survival
period can be extended to several months (Goss et al.,
1996). Sjogren and Gibson (1981) indicated that faecal bacteria survive even in very dilute mediums such
as lake waters.
3. Manure components that can affect the
microbial contaminant potential of manure
In addition to bacteria, which are the focus of this
review, manure contains mineral nutrients and organic
compounds, including endocrine disrupting substances and antibiotics, that can influence humans as
5
well as other organisms, and the complex breakdown
products resulting from digestion of plant material and
the by-products of animal metabolism. The carbon
content and the forms of carbon vary between manure
sources with the proportion of each carbon compound
type varying during storage (Eneji et al., 2001). In
solid manure there is considerable addition of bedding
material, but most manure contains fibrous material.
The manure of non-ruminants will tend to contain
more cellulose compounds, while the manure of ruminants will have a higher bacterial content because of
their production in the substrate-rich rumen (MacLean
et al., 1983). The carbon content of manure can be
correlated with its dry matter content. Hence, liquid
manures contain a smaller amount of carbon per unit
fresh weight than do solid manures. Although, liquid manures tend to contain fewer bacteria per unit
fresh weight, there are no significant differences in
the number of bacteria per unit dry matter, and thus
carbon content (Unc, 1999). The higher level of ammonium combined with a smaller carbon content in
the liquid manure results in an environment where N
is abundant. Pötter et al. (2001) found the C:N ratio
of liquid manure to be about 1:11. This still results in
an environment where it is generally considered that
C, rather than N, is the limiting factor for bacterial
growth. Nevertheless, much of the carbon in manure
is readily available to micro-organisms, so that when
manure is released into watercourses decomposition
results in a large demand for oxygen. Swine manure
has a very high biological oxygen demand (BOD),
ranging between 70,000 and 200,000 mg l−1 (ASAE,
1998). Such observations cast doubt on whether carbon is limiting in manure or whether it is the build-up
of toxic components that limits microbial activity.
An important feature of the carbon compounds
in manure is that they are a source of both soluble
and insoluble colloidal material (Giusquiani et al.,
1998; Gregorich et al., 1998). Elemental analysis,
acidic functional group determination, E4/E6 ratio,2
gel-filtration chromatography and magnetic resonance
spectroscopy analysis of acid soluble and acid insoluble dissolved organic carbon from pig slurry, organic
wastes and humic and fulvic acids from soils showed
similarities between sources (Giusquiani et al., 1998).
2 E4/E6 ratio: ratio of light absorption at wavelengths 465 nm/
665 nm.
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A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
Table 2
Estimated ionic strength and measured pH of manure solution (after Unc, 1999; Griffin and Jurinak, 1973)
Manure type
Manure–watera
contact time (min)
EC readings, average
(S.D.) (mS cm−1 )
Ionic strength
(M)b
pH average
(S.D.)
Solid beef
0
20
40
60
1.88
2.49
2.44
2.45
(0.33)
(0.14)
(0.24)
(0.18)
0.024
0.032
0.032
0.032
8.63
8.67
8.67
8.63
Raw manure
9.22 (0.29)
0.119
8.66 (0.09)
Liquid swine
a
b
(0.14)
(0.14)
(0.09)
(0.14)
g g−1 ;
Ration of 1:1
mixture mechanically shaken at 150 rpm.
Based on the relationship I (ionic strength) = 0.013 EC (electrical conductivity).
Carboxylic groups present on the surface of dissolved
organic matter from pig slurry conferred a complexing capacity similar to humic acids from a clay-loam
soil (Businelli et al., 1999).
The electrical conductivity (EC) of leachate from
manure indicates the presence of free charges. The
charges are mostly due to the presence of K+ , Cl− ,
Na+ , and NH4 + . Fewer ions enter the soil after application of solid manure than after liquid manure.
The latter has an equivalent ionic concentration of
about 0.1 M (Table 2). Leachates from piles of fresh
dairy manure have an EC value similar to that of liquid manure. Feedlot runoff may have an EC of about
4–6 mS cm−1 (Sweeten, 1998). The EC of manures increases with time during the first 200 days of storage
after which it declines slightly (Eneji et al., 2001).
As indicated in Section 2, the organic and inorganic
components of manure have an impact on the survival
of bacteria, and hence their potential for transport.
The presence of charged ions in solution can modify
the properties of both soil and microbial surfaces (see
Section 4), altering the potential for the adsorption of
microbes onto the soil. Fibrous organic matter will
encourage straining and filtration of microbes from
infiltrating liquid. Other organic carbon components
can alter the electrochemical properties of microbial
surfaces and thereby influence the likelihood that the
microbes will remain in the pore water (Section 4).
4. Transport of water and micro-organisms
from manure
Although bacteria from manure are known to enter
water resources (Goss et al., 2002) and numerous
studies indicate the potential for microbial contamination from manure sources (Joy et al., 1998; Stoddard
et al., 1998; Unc, 1999), there has been little progress
in the way of direct analysis of the physical and chemical interactions between the manure type and bacteria
that may affect the potential for contaminating water
resources. Transport and retention of particulates, such
as bacterial cells, in the soil is recognised as being
a very complex physical and chemical phenomenon
dependent on the interaction of the various complex
properties of soil, cells and suspending solutions. Four
major factors have been identified that influence the
movement of bacteria through the soil (Fontes et al.,
1991; Gannon et al., 1991a): (1) flow characteristics, which depend on of the grain size of the porous
medium, and on the soil structure that combined control the active porosity. We include the partitioning of
the water at the soil surface into runoff and infiltration
as a major aspect of this factor in bacterial movement.
(2) Filtration effects due to soil micropores, clogging
in macropores necks, and filtration pads formed by
solid components from applied manure (solid manure
mostly) as a function of the size of microbial cell. (3)
Straining within organic materials pads formed on soil
surface, which is the sum of the filtration and the electrochemical retention to organic surfaces within the
organic pads on soil surface. (4) Retention of bacterial
cells on soil mineral and organic particles by adsorption and adhesion, with the ionic strength of the soil
solution playing an important role. Retention is the resultant of complex interactions between components
of bacterial cells, soils and soil suspending solution.
4.1. Flow characteristics and water partitioning at
soil surface: general considerations
Transport of bacteria is mostly passive, being determined by the presence of rapid water fluxes, although
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
cell motility can play a significant role in movement
in suspension over short distances. Therefore, the path
followed by water, infiltration or surface runoff determines the direction of transport of bacteria from manure. Partitioning of rainfall or irrigation waters into
infiltration and runoff is a function of the soil hydraulic
properties, soil cover, slope, and the characteristics of
the precipitation. Accordingly, this partitioning of the
precipitation water is the main factor that determines
the resource that is at risk of contamination (surface
water or groundwater) and the rate at which the bacteria will move from the source.
There is evidence that bacterial transport in the
vadose zone may occur very rapidly in any field that
receives water at a sufficient rate to fill the soil pores
(McMurry et al., 1998; Unc and Goss, 2003). Water
can move downwards through the soil profile faster
than expected from estimates of the hydraulic conductivity of the soil matrix due to the preferential
(bypass) flow paths that avoid the soil matrix (Jury
and Flühler, 1992, Unc and Goss, 2003). These flow
paths are commonly equated with soil macropores.
Experimental results indicate that bypass flow is the
rule rather than the exception in most structured soils
(Simpson and Cunningham, 1982; Singh and Kanwar,
1991; Flury et al., 1994). This type of flow may
give contaminants a more direct and rapid path to
groundwater (Goss et al., 2002; Unc and Goss, 2003).
There is little agreement between researchers as to
how macropores are defined. Thus, the minimum
size of soil pores defined as macropores can vary
from >30 ␮m to >3000 ␮m (Beven and Germann,
1981). Under certain conditions, flow-bypassing of
the soil matrix may take place within capillary-sized
pores (Beven and Germann, 1982) or within vertical
zones of loose, porous, fine structured soil through
which rapid saturated flow occurs, phenomenon
known as fingering (Simpson and Cunningham, 1982;
Kung, 1990). Nevertheless, particulate materials move
through soil, carried by infiltrated waters, through
pores of a size that allows their passage. Only the
pores that are several times larger than the dimensions of the particulate allow transport over significant distances. Consequently particulates, such as
micro-organisms, travel only through the larger pores
where the water flux reaches peak velocities. As a
result particulates may be transported rapidly, in suspended state, over a distance determined by the extent
7
to which macropores are continuous (Jacobsen et al.,
1997) and thus the extent of preferential flow has a
direct influence on the depth to which particulates can
be moved. In general, experimental work has shown
no clear effect of the initial soil water content, or
water potential gradient on the flow pattern through
structured soils and the maximum penetration depth of
the water. In some cases, greater preferential flow has
been observed in wet soils (Flury et al., 1994). However, even if the mechanisms that create preferential
flow are relatively well understood, the development
of preferential flow cannot be predicted exactly from
the commonly measured physical characteristics of
soil.
Soil erosion–control practices have been developed
to favour infiltration over runoff, and consequently
infiltration tends to have increased importance in properly managed fields. The infiltration rate of water following a rain is often limited by the formation of a
soil crust that has a hydraulic conductivity of several
orders of magnitude less than that of the soil profile
(Morin et al., 1981). Seal formation on a bare soil depends on the kinetic properties of the raindrop in the
context of the physico-chemical properties of the soil.
The permeability of the seal rather than that of the underlying layer determines the equilibrium infiltration
rate for the soil (Shainberg et al., 1997). The decrease
in flux through the crust can create conditions for unsaturated flow in the soil profile underneath (Morin
et al., 1981). The associated reduction in soil surface
hydraulic conductivity can increase the runoff potential as the rainfall intensity will then be greater than
the infiltration rate.
4.1.1. Potential impact of manure on flow
characteristics and water partitioning at soil surface
The presence of ions with a greater hydration shell
at the surface of clay particles favours chemical dispersion of clay particles. As the exchangeable sodium
potential at the soil surface increases relative to Ca2+
the raindrop impact energy required to cause seal formation is reduced (Shainberg and Singer, 1988; Keren,
1990). Manure application, especially liquid manure,
results in increased ionic concentration at the soil surface. Long-term application of manure high in salts
can increase the amount of Na+ that is exchanged with
Ca2+ in soil favouring enhanced chemical clay dispersion. On the other hand, if few Na+ ions are within
8
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
the soil aggregates at the moment of manure application, the Na+ in manure probably has a less immediate effect. On the contrary, the main effect in such a
situation would be aggregation as the electric double
layers are compressed at the increased ionic concentration of the manure solution.
A change in pH is another factor that should be considered when manure liquids are applied as the clay
is more likely to be dispersed by water as soil departs from its electrical point of zero charge (Gillman,
1974). Application of liquid manure with little dry
matter content wets the soil surface. If rain occurs before the soil dries, even if the kinetic energy of raindrops is reduced, soil aggregates are more likely to
disintegrate and the water dispersible clay can seal the
smaller soil pores.
A cover of manure containing significant amounts
of bedding material can reduce the effect of the kinetic
energy of rain on soil surfaces and limit formation of
soil surface seals by decreasing the water potential at
the soil surface and by preventing ponded conditions.
By limiting pore occlusion at the soil surface, there is
a greater likelihood that colloidal particles and bacteria will move below the soil surface. However, it is
not known how migration of dispersed clay particles
from the bare soil surface in the spaces between the
manure aggregates participates in the overall seal formation. Unc (1999) showed that immediate infiltration rates were less reduced after application of solid
manure than after application of liquid manure. The
formation of a seal following application of liquid
manure, soil saturation, and subsequent ponding of
water, were given as possible causes. The increased
water-absorptive capacity of the bedding material
(Barry, 1999) in solid manure could also have been a
factor, in that it would reduce the hydraulic potential at
the soil surface and thereby alter the water partitioning.
Application of manure may well induce a sealing
effect at the soil surface, independently of the soil
seal formation. Barrington and Jutras (1983) noted
that when swine and dairy slurry was applied to soil,
two zones of filtration were established, one within
the solid mat, which accumulated at the soil surface
and the second at the soil–manure interface. Infiltration rates were reduced to a greater extent by the application of dairy slurry than pig slurry (Barrington and
Jutras, 1983). However, the authors did not describe
the structure of the two filtration zones. The first zone,
within the manure mat could have been caused by the
disintegration of the manure aggregates and sedimentation of the manure colloids. However, it is less certain how the filtration zone at the boundary between
the manure and soil surface was created, and therefore
if it was a function of the manure composition, soil
characteristics, or both.
4.2. Straining and filtration within the vadose zone
Micro-organisms in liquid manure are already in a
dilute solution when the manure is applied to the soil.
Those in solid manure are more associated with material surfaces and have to be dislodged by the impact
of precipitation or the flow of water. Once in suspension, micro-organisms can be removed from the water
stream by sieving within the manure material, filtration in smaller soil pores or pore necks, discontinuities
in the macroporosity, or electrochemical retention on
the surfaces of the soil particles. Sieving and filtration occurs due to the larger size of bacteria relative
to pore diameters. In contrast to sieving and filtration,
retention at the surface of soil particles depends on
the interaction between the charged surface properties
of both bacteria and soil particles, and both can be
altered by the composition of the suspending liquid.
4.3. Charge mediated bacterial retention
Most soil mineral and organic surfaces have a net
negative surface charge. The intrinsic charges at the
soil particle surfaces are a function of the chemical
composition of the soil minerals and the presence
of organic material. Clay minerals carry a net negative charge due to the substitution of divalent cations
in their alumino-silicate structure, and thus cationic
exchange is the most important mechanism for the
adsorption of organic compounds onto clays to form
clay–organic complexes. Organic cations, such as
protonated amines or quaternary ammonium cations,
replace the inorganic cations on the clays. Such organic molecules have both hydrophobic and polar
groups and can act as cationic surfactants conferring
hydrophobic properties to the negatively charged clay
surfaces.
Changes in the environmental pH can change the
ionisation of surface groups and therefore change the
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
electrical potential at the surface. If the ionic strength
of the solution is great then changes in pH will have a
negligible effect (Jones, 1975). Increases in the electrolyte concentration result in the neutralisation of a
greater number of charged surface loci by counter
ions in the solution (Bennett and Hulbert, 1986). This
leads to smaller electrostatic repulsion forces as the
electrical double layer thickness is reduced (Jones,
1975). Presence of polyvalent ions (2–3–4) also
favours aggregation (Jones, 1975; Hunter, 1981) although Wasserman and Felmy (1998), noted that this
is true only at dilute concentrations of the suspending
solution. As the ionic strength increases above 0.01 M
the type of ions present in the solution has little effect on cell surface charge density. Hence, at large
electrolyte concentration, and high thermal energy,
the two surfaces can be brought sufficiently close
for the short range London–van der Waals forces to
take effect. Adsorption at this level occurs at what
it is known as the primary minimum for the sum of
attraction and repulsion forces. Although individually these attraction forces are weak, multiplicity of
the connection points on the two charged surfaces
can lead to an overall strong bonding. However,
presence of the larger hydrated ions generally limits
this occurrence in natural soil environments (Huang,
1995).
4.3.1. Surface charge properties of soil particles
The mineral phase of soils becomes strongly
hydrated in the presence of water. Hydrophobic
particles cannot compete with water and they become desorbed from the mineral surfaces and therefore interact mainly with soil organic matter (Boyd
et al., 1991). Adsorption initiated by cationic exchange can be subsequently enhanced by hydrogen
bonding, van der Waals, and hydrophobic bonding
(Hayes and Himes, 1986). Water content and the pH
of the system can also affect adsorption of organic
acids onto mineral surfaces. A low pH favours adsorption of humic acids into clay interlayers. The resultant
clay–organic complexes can have a strong affinity for
suspended organic matter that is negatively charged,
but still has hydrophobic properties (Mortlandt,
1986; Schnitzer, 1986). For these reasons, the retentive capacity of the soils for hydrophobic compounds is directly correlated with the organic matter
content.
9
While most of the work on the retention of organic
compounds on mineral surfaces has been conducted
on crystalline clay minerals, the presence of iron oxides also has a significant impact on the retention of
organic molecules in certain soil environments. Colloidal iron oxides have very active surfaces and often
coat crystalline minerals, such as silica oxides, in the
soil environment. Consequently, surfaces coated with
oxide colloids are very active and interact readily
with organic colloids (Huang, 1995). Soils, such as
sandy soils, with a significant presence of iron oxide
coatings have a large retention capacity for suspended
organic compounds, despite their small organic matter
content.
4.3.2. Surface charge properties of bacterial cells
The charge properties of bacterial surfaces depend
on the constituents of their cell wall and the presence
or absence of surface appendages. There are two basic organisational schemes for the bacterial cell wall,
which correspond with the Gram staining reaction.
Phospholipids are the chief lipids in many bacterial
cell membranes (Van Bruggen, 1971). Phospholipids
have a hydrophobic end, usually located within the
cell membrane and a negatively charged terminal that
can be located towards the surface of the cell wall.
The cell wall also contains structural, metabolic and
transport proteins as well as other components. The
presence of free amine radicals on the proteins results
in net positive charges. Lipids and proteins interact
through polar and hydrophobic forces leading to the
formation of lipoproteins in the cell wall.
The relatively small amount of protein and the
greater amount of phospholipids results in the bacterial wall mainly expressing hydrophobic and negative
charges. Consequently, bacteria generally possess a
net negative surface charge at most pH values found
in soils (Stotzky, 1985). At first sight, this should
suggest a repellent effect between bacteria and soil
surfaces. However, because of the charge difference
between these surfaces a potential exists between
them and the bulk aqueous phase. To counterbalance
the surface charge, ions of opposite charge are loosely
attracted to the surface to form a diffuse double layer
of ions. Thus, the charge density of a cell surface is a
function of the intrinsic charge density within the cell
wall and the induced charge density as a function of
the ionic make-up of the solution. However, the net
10
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
charge of the cell surface and its associated double
layer will be zero.
If the negatively charged cells come in contact with
a positively charged surface the attachment tends toward 100% and few bacteria remain in the solution
phase. As the concentration of the suspending solution
increases, the percentage of cells that are attached to
soil surfaces can decrease due to competition between
anion groups on the cell surface and anions in solution (Makin and Beveridge, 1996). On the other hand,
interactions that occur at the interface between cell
surfaces and other surfaces present in soil (Marshall,
1985), can be influenced by the hydrophobic interactions between these surfaces and water. Whenever a
charged surface is introduced in suspension, the bipolar water molecules reorient as a function of the surface charge signs and form an ordered layer on the
surface. If two charged surfaces are suspended in water, they can attach to each other only by displacing
this structured layer of water. This can only happen
if the attraction forces overcome the water attachment
forces.
Attachment of a cell to a surface can be influenced
by the roughness of the cell surface (Dickson and
Siragusa, 1994), or the presence of fimbriae (Stenström
and Kjelleberg, 1985; Stenström, 1989) both of which
can override the effect of surface charge or hydrophobicity. The presence of localised positively charged
groups at the tip of hydrophobic fimbriae can be more
essential in the adhesion process than the cell surface
with an overall negative charge (Stenström, 1989). On
the other hand, motile bacteria that possess fimbriae
have increased detachment rates, positively correlated
with the soil temperature, enhancing the potential for
transport through soil (McCaulou et al., 1995). It is
worth noting that under laboratory conditions, the
addition of sub-lethal amounts of antibiotics, as used
to increase feed efficiency in animals (Hays, 1985),
can cause dramatic changes in the cell surface properties, and can even result in the obliteration of surface
structures and appendices (Loubeyre et al., 1993).
Presence of lipopolysaccharides polymer (mucilage), produced by active bacterial cells under specific conditions (e.g. starvation), may enhance their
adhesion to mineral surfaces provided that the bacteria are brought in contact through other mechanisms
and the contact period is long enough (Stotzky, 1985).
On the other hand, initial attachment is impeded by
the presence of lipopolysaccharides capsules if both
the polymer and the surface are hydrated (Wrangstadh
et al., 1986). Whereas, the adsorption of bacterial
cells to soil particles that results from surface electrical charges is reversible under the range of concentrations found in soil solution, the attachment due
to polysaccharide mucilage tends to be irreversible
(Stotzky, 1985).
4.3.3. Impact of manure on charge mediated retention
Charged groups on soil and bacteria can act as
donors or acceptors in the hydrogen bonding processes
and therefore are strongly affected by pH changes
(Bengtsson et al., 1993), and by the ionic strength of
the solution. Adding electrolyte, such as manure liquids, to the suspending solution causes a compression
of the diffuse parts of the electrical double layer and,
in addition, ions are adsorbed in the Stern layer of both
microbial cells and soil particle surfaces. This allows
bacteria to approach soil surfaces at closer range. If
the double layer is compressed sufficiently the range
of the repulsive interaction of the double layer is reduced enough for the van der Waals forces to predominate. This is when flocculation at the surface of the
soil particles may occur.
The pH of manure solution ranges between 8.0 and
8.5 (Table 2). Hence, the application of manure on
neutral to slightly acidic soils could increase the pH
temporarily. An alkaline pH can result in immobilisation of certain surface associated cations, thereby increasing the chances for bacteria to be removed from
the adsorption sites and released into the soil solution
(Stotzky, 1985). In contrast, volatilisation of ammonia
from the liquid phase of manure can lead to temporary slight acidification of soils when excess H+ ions
(left from NH4 + ) remain in the solution. This is most
likely to occur when manure is applied to warm, dry,
alkaline soils. If the ionic strength of the solution is
relatively high, then changes in pH will have a negligible effect (Jones, 1975; Jewet et al., 1995).
The EC of soil solution is generally around
0.3–1.2 mS cm−1 (0.003–0.02 M), but is greater for
saline soils. The ionic strength of manure and that
of manure–water mixtures (Table 2) are significantly
greater than these values. It seems reasonable to assume that the types of ions in the soil solution are
less relevant in the bacteria–solid surfaces interaction,
when manure liquids enter the soil.
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
4.4. Hydrophobic mediated bacterial retention
Hydrophobic interactions are identified by numerous studies in microbial cell surface phenomena
(Doyle and Rosenberg, 1990). The effect of cell hydrophobicity on adhesion can be described as being
less a function of the cell surface interaction and
more of a function of the bacterial cell expulsion
from the bulk of water due to the strong tension at the
cell–water interface (water surface tension). Therefore, strong hydrophobicity of bacterial cell surface
can result in large rates of bacterial adsorption at
the solid–liquid interface (Lindqvist and Bengtsson,
1995). Hydrophobicity was positively correlated with
bacteria adhesion to non-aqueous phases including
mineral particles (Marshall, 1980, 1985; McAneney
et al., 1982; Kjelleberg, 1985; Stenström, 1989),
while hydrophilic strains are more likely to be found
in the bulk water. Hence, hydrophobic bacteria will
tend to attach themselves to the liquid–solid interface at greater rates than would hydrophilic bacteria.
However, attachment is reversible, with a time scale
for detachment being on the order of days or weeks.
Slower attachment and detachment rates were observed for hydrophilic compared with hydrophobic
strains, suggesting that the former would move further before being removed by attachment to soil, but
once attached, they would become detached again at
a slower rate (McCaulou et al., 1994). Consequently,
hydrophobic strains have been found to move slower
than hydrophilic ones (Huysman and Verstraete,
1993).
Surface roughness and presence of fimbriae as well
as increased cell surface hydrophobicity favour cell
adhesion to an air–water interface (Dahlbäck et al.,
1981; Hermansson et al., 1982). Under field saturated
conditions there are no air–water interfaces in the soil,
barring the presence of trapped pockets of air. It might
be assumed that presence of the charges on cell surfaces increase the chances of polar interactions with
water molecules and therefore decrease the overall hydrophobicity. However, there is not always a direct
link between surface charge and hydrophobicity, and
a cell with small or no net surface charge can still
be covered by a large number of positively and negatively charged sites. Presence of hydrophilic molecular groups on the cell surface can neutralise the effect
of nearby hydrophobic molecular groups (Doyle and
11
Rosenberg, 1990). Some relatively hydrophobic cells
also have high negative electrokinetic potentials (van
Loosdrecht et al., 1987). This combination may sound
contradictory but the charged groups only occupy a
minor fraction of the total surface area (under 10% of
the surface), and the surface potential results only in
part from charged groups on the outer surface, the rest
originate from groups situated in deeper layers of the
cell walls.
While most bacterial cells, both Gram-negative
and Gram-positive, have a net negative charge, their
hydrophobic character can vary within the same
species or even strain (Stenström, 1989). Strongly
hydrophobic cells can become irreversibly attached at
the primary minimum of the electrical double layer.
However, for more hydrophilic (charged) cells the
electrokinetic potential becomes more important and
bacteria can attach at the secondary minimum of
the electrical double layer (van Loosdrecht et al.,
1987). Lindqvist and Bengtsson (1991) showed that
the surface charge generally plays a minor role in the
association of bacterial cells with mineral surfaces
compared to the hydrophobic characteristics, even in
hydrophilic strains. However, surface charges may
enhance the effect of hydrophobicity on adsorption
especially at high electrolyte concentrations when the
electric double layer is compressed. Consequently, it
can be concluded that the surface charge as estimated
by electrophoretic mobility (zeta potential) may be
an indicator of the degree of sorption for hydrophilic
bacteria, but has little relevance for hydrophobic
strains (van Loosdrecht et al., 1987).
In practice, it can be expected that a combination of
high hydrophobicity with low surface potential can be
detrimental to bacteria survival, as such bacteria cannot move once the substrate is depleted. Other experimental evidence suggests that bacterial dispersal is
a dynamic non-equilibrium process, possibly shaped
by two sub-populations, one strongly (maybe irreversible) adsorbed to solid surfaces and one with very
slow adsorption kinetics (Lindqvist and Bengtsson,
1991).
4.4.1. Impact of manure on hydrophobic mediated
bacterial retention
The relative amounts of different lipopolysaccharides present on the surface of cells are under
environmental control (Makin and Beveridge, 1996).
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A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
This indicates that the cell surface hydrophobicity
is influenced by the growth medium. Cell surface
hydrophobicity was found to be about eight times
greater in groundwater compared to eutrophic lake
water (Lindqvist and Bengtsson, 1991). The presence of a rich growth substrate, such as manure, may
lead to reduction of cell hydrophobicity; therefore,
increasing the effect of the surface charges on adsorption in the soil (Stotzky, 1985). Under the high
ionic strength of manure solution, this can potentially
lead to cell adsorption and cell aggregation. Increased
sorption, rather than dispersal of bacterial cells, is
associated with poor growth conditions, confirming
that cultivation of bacteria on oligotrophic substrates
results in increased cell hydrophobicity (Lindqvist
and Bengtsson, 1991).
Adding charged salts and soluble organic compounds to water lowers its surface tension. For example surface tension of human urine is around
42–63 mJ m−2 , compared to 72.8 mJ m−2 for pure
water (Reid et al., 1991). This should reduce the hydrophobic forces and the potential for retention of
hydrophobic particles at phase interfaces. Lindqvist
and Bengtsson (1995), found that the importance of
hydrophobicity on the bacterial cell adsorption at surfaces within an aquifer was found to be greater when
the groundwater solution had a small C content. Similar qualitative results were obtained by using ionised
and uncharged bacterial sized microspheres, although
quantitatively the results differed. Differences in the
level of net surface charge and hydrophobicity of the
microspheres compared to bacterial cells, and also to
potential increased bacterial motility, could have been
responsible for the contrast in quantitative rather than
qualitative results.
4.5. Other cell characteristics affecting transport
through the vadose zone
Gannon et al. (1991b) observed that bacterial transport through saturated sterilised soil was a function
of cell size, with other surface characteristics having no significant impact. There was no information
given on the amount of soluble organic carbon in the
soil as a result of the sterilisation procedure. Similar
conclusions were drawn after testing the transport of
carboxylated artificial microspheres through repacked
sandy soil columns (Harvey et al., 1993, 1995). Size
selective transport is consistent with macropore flow
being important in bacterial transport (Gannon et al.,
1991a) in both undisturbed and disturbed soils. The
presence of active macropores is nevertheless a function of the soil structure in undisturbed soils and grain
size in disturbed soils.
The shape of the particles may influence the behaviour and structure of the electrical double layer.
Jones (1975), and Hunter (1981), provided analyses of
the impact of the double layer on the interaction between surfaces of spherical and cylindrical particles.
Generally, however, the local radius of curvature of
bacteria cells is large compared to the double layer
thickness. Thus, it may be assumed that the actual
shape of the bacterial cell has little influence on the
double layer behaviour, and thus the model for plane
surfaces can be used (Jones, 1975; Wasserman and
Felmy, 1998).
4.6. Impact of organic carbon compounds on the
retention of bacterial cells in the vadose zone
In natural soil environments there are soluble and
insoluble C compounds that interact with mineral and
particulate surfaces. Faster organic matter mineralisation can result in more water-soluble carbon in
sandy soils compared with finer textured soils (Bernal
et al., 1992). The strongly adsorbed organic material
on the clay particles favours retention of non-ionic organic compounds. Dissolved humic substances have
been found to associate predominantly by hydrophobic
forces, as the apolar components of humic substances
largely control their aggregation and reactivity in the
environment (Piccolo et al., 1999). Dissolved soil organic acids act similarly to a cationic surfactant (Smith
et al., 1991). Long carbon chain organic cationic surfactants displace the ionic cations bound at the soil surfaces. Cationic surfactant molecules attach themselves
or as hemi-micelles to the cationic exchange sites in
soil (Yeskie and Harwell, 1988). Thus, adsorption of
cationic surfactants tends to increase the hydrophobic
character of the soils and hence favour retention of
suspended hydrophobic molecules (Mortlandt, 1986;
Lee et al., 1989; Smith et al., 1991).
Retention of organic cations in soils is dependent on the soil cation exchange capacity (CEC).
Hence, the use of cationic surfactants for increasing
hydrophobic retention is expected to be more effective
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
in soils with greater clay content and hence greater
cation exchange capacity (Mortlandt, 1986; Burris
and Antworth, 1992). Most of the tests using organic
cationic surfactants have been performed on aquifer
materials with low organic matter content (Brown and
Burris, 1996).
4.6.1. Impact of manure carbon compounds on the
retention of bacterial cells
Increased amounts of organic matter in the unsaturated soil profile, as well as the greater organic and
bacterial concentration following manure application,
can enhance the number of possible interactions between a surfactant and the supplementary polar and
non-polar surfaces present, hence increasing the complexity of the system. Dissolved soil organic matter
tends to interact with the free ions and hydrophobic
pollutants increasing their solubility, and hence favouring their transport (Smith et al., 1991). As the solubility of the hydrophobic molecules is increased by the
presence of organic material in solution, manure application would be expected to favour more hydrophobic
bacteria to stay in solution. As the carbon compounds
present in soil can interact with the soil surfaces in the
same way as bacterial cells, the presence of a large
amount of manure carbon compounds may increase
the competition for the attachment sites in the soil
(Stotzky, 1985).
Bacterial cells have very active surfaces where
organic molecules from the solution can attach
(Bengtsson et al., 1993). Polar or apolar organic
molecules can interact with both the charged and the
hydrophobic sites on bacterial cells similar to the interaction of the cells and organic material with soil
surfaces. This leads to the flocculation of carbon compounds on the bacteria surfaces. Hence, the properties
of the new particle, with bacteria at the centre, may
have electrochemical properties that are entirely different from the electrochemical characteristics of the
cell. The bacteria may metabolise some of the carbon
compounds with the result that the electrochemical characteristics of the surfaces of bacteria–carbon
compound complexes become fluid having the potential to vary independently from external factors.
The attachment behaviour of those bacteria capable
of degrading dissolved organic carbon is not very
well understood. Cell flocculation in the presence of
large amounts of dissolved organic carbon could be
13
under some genetic control (Bossier and Verstraete,
1996).
While the electrochemical behaviour of the soil organic components has been extensively researched, the
same cannot be said for the manure organic components applied to soil. However, manure organic matter
was found to have similarities with the soil organic
matter compounds (Businelli et al., 1999, Giusquiani
et al., 1998).
Manure can be a significant source for colloidal
matter of faecal origin. Greater concentrations of colloids favour their transport. Some of the colloids are
attached to the soil particles in a single layer. Thus, the
colloids left in solution are then less likely to become
attached given their similar surface charge with the
attached colloids (Kretzschmar et al., 1995). Alternatively, attachment of natural organic matter on the
surface of colloids enhances colloid mobility by decreasing the attachment efficiency of the colloids to
the matrix surface. Greater colloid stability due to adsorbed natural organic matter decreases soil retention
capacity thereby increasing transport (Kretzschmar
et al., 1995). At low colloid concentrations, soil
retention capacity is maintained longer (Seta and
Karathanasis, 1997). Given the large concentration of
colloidal material in the manure solution, one cannot
exclude the possibility that bacteria attach to smaller
colloidal particles from the manure solution. This
may retard adsorption to larger mineral particles and
therefore enhance transport.
5. Conclusions
Survival and transport of manure bacteria in the vadose zone have been reviewed in terms of the impact
that manure components may have on the physical and
chemical properties of soil and on the surface characteristics of the bacteria themselves. The impact of
manure components on the contaminant potential of
manure bacteria can be manifold (Table 3). Manure
changes the soil pH and increases the amount of salts
and soluble and insoluble organic compounds. Bacteria grown in manure can exhibit surface properties
entirely different from the strains grown in the laboratory. The presence of faecal and bedding particles
at the soil surface has an impact on the likelihood
that water from precipitation will enter the soil. The
14
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
Table 3
Potential impact of manure properties on the estimated contaminant risk to surface and groundwater resources
Manure properties
Potential impact
Estimated contaminant
risk to water resources
Microbial population
Source for contamination
Variable
Dry matter content
Filtration and straining of bacteria within manure and at manure soil
interfaces
Protection from UV rays
Reduced hydraulic gradient at surface (vs. liquid manures) limits soil
dispersion caused by the kinetic energy of rain; more pore available for
transport
Decreased infiltration due to surface sealing
Residual absorbent capacity of bedding material
At low dry matter content (liquid manure) flowing manure solution; flow
rate and path dependent on hydraulic gradient at soil surface
Diminishes
Ionic concentration and
ionic species
Favours interaction of like charged surfaces (soil and bacterial cells);
increases cell retention
Competition for retention sites; decreases cell retention
Increased soil structural stability; improved macropore stability and continuity
If high Na+ content; clay dispersion; decrease aggregate stability and thus
macropore integrity
Available mineral nutrients from manure can modify the physiological
response of soil organisms
Increases
Increases infiltration
risk
Increases runoff risk
Diminishes
Variable
Diminishes
Increases
Increases
Diminishes, Variable
Variable
pH
Initial manure pH is slightly alkaline; ammonia volatilisation can acidify
slightly the soil environment
Variable (likely not
significant)
Soluble and colloidal,
charged and hydrophobic
carbon compounds
Competition for retention sites
Increases
Interaction with bacterial cell charged and hydrophobic surface loci
Surfactant-like behaviour may favour flocculation and retention
Variable
Diminishes
Available carbon
Increases the physiological response of soils. Accelerates soil metabolic
activities, increasing predation and competitive pressures
Diminishes, Variable
High ammonia and fatty
acids concentration
Toxic effects within manure
Diminishes
Toxic effect in soil after application can modify the interaction between
manure and soil microbes
Variable
available carbon and nutrients in manure can affect
the survival of manure pathogens in soil in part because the same components can also impact the soil
microbial population thereby influencing the interaction between the two groups of micro-organisms. Manure impacts on all aspects of transport and survival
of manure pathogens in the soil environment need to
be fully quantified before an evaluation of the importance of each type of impact and their interactions can
be appropriately done.
Consideration needs to be given to the impact of manure characteristics on the transport of both precipitation (rainfall and irrigation water) and bacteria, shortly
after manure application, as well as over longer periods
of time. Thus, future research is required to evaluate
the effect of manure composition on soil properties
controlling water transport, the flow regime within the
soil, and the potential impact of manure components
on the water partitioning at the soil surface. There is
a need to bridge the gap between the practical information on the contaminant potential of manure and
the detailed assessments that are relevant to the interaction of bacterial cells with soil surfaces. More understanding is required on how soluble and colloidal
organics in manure interact with the surfaces of bacteria and affect their retention to soil surfaces. Given
A. Unc, M.J. Goss / Applied Soil Ecology 25 (2004) 1–18
that many of the studies on surface properties have
been performed in laboratories, there is a need to evaluate the level at which surface properties derived from
such studies can be used to estimate bacterial retention
and transport in field soils.
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