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Bell & Howell Information and Learning
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Timberley Michelle Roane
Dissertation Submitted to the Faculty of the
In Partial Fulfillment of the Requirements
For the Degree of
In the Graduate College
UMI Number: 9934848
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As members of the Final Examination Committee, we certify that we have
read the dissertation prepared by
Timhprlpy Mirhpllp Rnanp
Bioaugmentation with Metal-Rp..qi st-anr Mirrnnrgam'in >-hp
Remediation of Metal and Drgam'r CnntaTninarpri
and recommend that it be accepted as fulfilling the dissertation
requirement for the Degree of
Doctor of Philosophy
Ian/ L^ fepp^
Charles P. Gerba^
Leland S. Pierson III
Wayne L. Nicholson
Final approval and acceptance of this dissertation is contingent upon
the candidate's submission of the final copy of the dissertation to the
Graduate College.
I hereby certify that I have read this dissertation prepared under my
direction and recommend that it be accepted as fulfilling the dissertation
Dissertatioli Director
Ian L. Pepper
This dissertation has been submitted in partial fulfillment of requirements for an
advanced degree at The University of Arizona and is deposited in the University Library
to be made available to borrowers under rules of the Library.
Brief quotations fi:om this dissertation are allowable without special permission,
provided that accurate acknowledgement of source is made. Requests for permission for
extended quotation fi:om or reproduction of this manuscript in whole or in part may be
granted by the head of the major department of the Dean of the Graduate College when in
his or her judgement the proposed use of the material is in the interest of scholarship. In
all other instances, however, permission must be obtained from the author.
I have truly enjoyed my Ph.D. experience and while this dissertation is the
culmination of long hours, many laughs and some tears, this accomplishment is not solely
mine. I owe a lot to my fellow labmates (Mark Burr, Scott Dowd, Miriam Eaton, Karen
Josephson, Eileen Jutras, Debbie Newby, Kelly Reynolds and Christine Stauber) for their
support and humor (remember the fire in the hood?). Many thanks to Dave Bentley at
The University of Arizona Imaging Facility for, well, not using his right to refuse
service... Special thanks to Beth Marlowe (for whom the gene exists) who is and always
will be a best fiiend and a respected colleague. Thanks Emily Petrovich for making me
think, making me laugh and making me cry. I hope we will meet again someday.
Of course, thanks to my committee members Charles Gerba, Raina Maier, Wayne
Nicholson, Ian Pepper, Sandy Pierson and Norval Sinclair for their encouragement and
suggestions. I would like to especially acknowledge the support of Ian and Raina.
Because of their different viewpoints, the end product was always that much better.
Thank you Ian for giving me the opportunity and the creative freedom to conduct
research, teach, write book chapters and grant proposals, to make mistakes and to correct
them, to grow and become my own scientist. And Ian, "How am I doing?"
This dissertation is dedicated to my parents, James and Lavata Roane, for without
their endless patience, support and guidance this dissertation would not have been
possible, and to Livy Williams HI, for without his encouragement and understanding, I
would be less of a scientist and less of a person.
"I am only one.
But still I am one.
I cannot do everything,
But still I can do something;
And because I cannot do everything
I will not refiise to do the something that I can do."
Edward Everett Hale
Problem Definition
Microbial Remediation of Metals
Metals in the environment
Physical and chemical remediation of metal-contaminated soils
Metal speciation and bioavailability
Metal toxicity to microorganisms and microbial resistance
Origin of microbially based metal remediation
Microbial interactions with metals
Innovative approaches to microbial remediation
Soil and sediments
Aquatic systems
Nuclear waste
Concluding remarks
Dissertation Format
Remediation of a Co-contaminated System
Cadmium-Resistance in Microorganisms
Degradation of 2,4-D
Materials and Metiiods
Field site characterization
Community cadmium resistance
Characterization of cadmium-resistant isolates
Results and Discussion
Culturable Microorganisms in Metal-Contaminated and
Types of Cadmium-Resistant Isolates from Contaminated and
Effect of Cadmium on Resistant Isolates from Contaminated and....71
Figure Legends
Materials and Methods
Bacterial strains and culture conditions
Plasmid profiles
Detection of cad A and cadC
Production of extracellular polymers
Production of microbial surfactants
Transmission electron microscopy
Cadmium mass balance
Results and Discussion
ATP-dependent efflux
Extracellular binding
Cellular accumulation of cadmium
Effect on cadmium solubility
Figure Legend
Materials and Methods
Isolation of cadmiimi-resistant bacteria
Cadmium mayimnm resistance level
Degradation studies
Results and Discussion
Solution studies
Laboratory soil microcosms
Intermediate field scale
Figure Legend
Table 1-1. Typical background levels of heavy metals
Table 1-2. Tj^ical metal concentrations in contaminated
Table 1. Characterization of contaminated and uncontaminated soil
Table 2. Changes in soluble cadmium with community growth
Table 3. Isolate identification and cadmium-resistance
Table 4. Antibiotic profiles
Table 5. Changes in soluble cadmium with Pseudomonas HI
Table 1. Cadmium-resistant isolates and mechanisms of resistance
Table 2. Influence of microbial growth on cadmium solubility
Table 1. Cadmiimi-resistant soil bacterial isolates
Table 2. Description of field scale bioreactors
Figure 1-1. Rhamnolipid and siderphore
Figure 1-2. Remediation strategies for soils and sediments
Figure 2-1. Microbial metal resistance mechanisms
Figure 2-2. Model of microbial cadmium resistance
Figure 2-3. 2,4-D degradation pathway
Figure 1. Total community response to cadmiiun
Figure 2. Response of cadmium-resistant populations to cadmium
Figure 3. Agarose gel of ERIC PCR
Figure 4. Response of Pseudomonas HI to increasing cadmium
Figure 1. TEM micrograph of Pseudomonas II a with EPS layer
Figure 2. TEM micrograph showing cellular cadmium accumulation ....106
Figure 3. X-ray analysis of cellular accumulation
Figure 1. Dual bioaugmentation in solution studies
Figure 2. Dual bioaugmentation in soil microcosms
Figure 3. Dual bioaugmentation in soil microcosms
Figure 4. Dual bioaugmentation in intermediate field scale
Current thinking is that co-contaminated sites (i.e., sites with both organic and
metallic pollutants) are difficult to bioremediate because the metal toxicity is such that
organic degradation is inhibited. The objective of this research was to evaluate the
potential of bioaugmentation with metal-detoxifying microbial populations as a viable
remediative approach for such sites. Divided into three sections, this research found that
metal-detoxifying microorganisms could facilitate the remediation of co-contaminated
The objective of the first study was to compare the microbial community response
to cadmium exposure between metal-contaminated and uncontaminated soils. This study
found that while cadmium adversely affected the numbers of culturable microorganisms
in all soils, cadmium-resistant isolates were found in each soil, regardless of prior metal
However, the metal-contaminated soil microbial communities were more
resistant than the uncontaminated soil community. In one metal-stressed soil, resistance
increased with increasing cadmium stress. A cadmivmi-resistant Pseudomonas spp. was
foimd to increase in nimibers with increasing cadmium, suggesting a different mechaoism
of cadmium resistance at high cadmium concentrations.
The second study evaluated the diversity of cadmixrai-resistance/detoxification
mechanisms in six cadmium-resistant isolates found in the iBrst study. Genetic and
microscopic analyses foimd several different approaches to cadmium resistance. Two
mechanisms known to confer resistance were observed, including exopolymer and
biosurfactant production.
Two other isolates demonstrated intracellular cadmium
accumulation via as yet unknown mechanisms. The mechanism of resistance for one
isolate could not be identified. Four out of the six isolates detoxified cadmium as part of
their resistance.
Since metal detoxification is necessary to allow for organic degradation, these
four isolates were included in 2,4-D degradation studies under co-contaminated
conditions. The last study examined the use of cadmium-detoxifying microorganisms to
enhance organic degradation under co-contaminated conditions. In pure culture and
laboratory soil microcosms with cadmitim and 2,4-dichlorophenoxyacetic acid (2,4-D) as
model contaminants, four cadmium-detoxifying isolates supported the degradation of 2,4D by the cadmium-sensitive 2,4-D degrader Alcaligenes eutrophus JMP134 in the
presence of toxic levels of cadmium. In a pilot field study, a cadmium-detoxifying
Pseudomonas isolate enhanced 2,4-D degradation by A.eutrophus JMP134 in the
presence of cadmium.
Problem Definitinn
Thirty-seven percent of all contaminated sites in the United States are cocontaminated with both metal and organic pollutants. These sites are generally not
considered for bioremediation because the toxicity of the contaminating metal(s) is too
great. Traditional approaches to remediation, including excavation and incineration, are
often too expensive and too destructive and are not effective when large areas of
contamination are involved. New approaches to the treatment of increasingly complex
industrial wastes and akeady existing sites need to be addressed. One of these new
approaches is microbially-facilitated remediation.
Few studies address the ability of microorganisms to treat co-contamination.
Remediation with microorganisms has several advantages, such as in situ treatment, ease
of application, practicability for large areas, potential for surface and subsurface
treatment, effectiveness in soils, sediments and waters, and potential for the transfer of
genetic elements to indigenous microbial populations. Much research has gone into the
development of so-called "SuperMicroorganisms" with both metal-resistant and
degradative genetic components. However, these engineered organisms have not been
successful under field
conditions, possibly due to the inability to compete with
indigenous microbial populations and/or the substantial energy requirements needed to
sustain both degradation and metal-resistance cannot be met. Much is known about the
ability of various microbial groups to degrade xenobiotic compoimds, and about the
ability of microorganisms to resist and in many cases detoxify, metals. This research
addressed the question as to whether separate populations of metal-detoxifying and
organic-degrading microorganisms could work together to remediate co-contaminated
Microbial Remediation of Metals
Metals in the Environment
Metal pollution is a widespread problem, in fact, in industrially developed
countries it is normal to find elevated levels of metal ions in the environment. In
addition, it has been estimated that approximately 37% of sites in the U.S. contaminated
with organic pollutants, such as pesticides, are additionally polluted with metals
(Kovalick, 1991). Despite this, biological treatment or bioremediation of contaminated
sites has largely focused on the removal of organic compounds, and only recently has
attention turned to the treatment of metal-contaminated wastes (Brierley, 1990; Summers,
1992). Due to their toxic nature, the presence of metals in organic contaminated sites
often complicates and limits the bioremediation process. Such metals include the highly
toxic cations of mercury and lead, but many other metals are also of concem including
arsenic, beryllium, boron, cadmium, chromium, copper, nickel, manganese, selenium,
silver, tin and zinc.
Metals are ubiquitous in nature and even those metals generally considered as
pollutants are found in trace concentrations in the environment (Table 1-1). For the most
part, metal pollution problems arise when human activity either disrupts normal
biogeochemical cycles or concentrates metals (Table 1-2). Examples of such activities
include mining and ore refinement, nuclear processing, and industrial manufacture of a
variety of products including batteries, metal alloys, electrical components,
preservatives, and insecticides (Gadd, 1986; Hughes and Poole, 1989; Suzuki, Fukagawa
and Takama, 1992).
Metals in these products or metal wastes from
processes can exist as individual metals or more commonly as metal mixtures.
Unfortunately, past waste disposal practices associated with
and manufacturing
activities have been such that air, soil, and water contamination was common, and as a
result there are many metal-contaminated sites that pose serious health risks.
A case in point is the Bunker Hill Mining Company located near Kellogg, Idaho. In
1972, the Environmental Protection Agency (EPA) ordered the Company to limit dumping
of its smelter wastes. At this time, typical smelter releases were up to approximately 27,215
kg lead per 1.6 sq. km on surrounding areas within a six month time period. As a result,
there was little or no vegetation within 1.6 km of the smelter and, according to a 1974 study
conducted by the Idaho Department of Health and Welfare, approximately 40%
Table 1-1. Typical background levels of heavy metals in noncontaminated soil and
aquatic environments.
gold (Ag)
aluminum (Al)
arsenic (As)
barium (Ba)
cadmiimi (Cd)
cobalt (Co)
chromixim (Cr)
cesium (Cs)
copper (Cu)
mercxiry (Hg)
manganese (Mn)
nickel (Ni)
lead (Pb)
tin (Sn)
zinc (Zn)
Source: Goldman & Home, 1983; Leppard, 1981; Sigg, 1985.
''Source: Lindsay, 1979. "^ND = no data reported.
Srace = levels usually below detection.
Table 1-2. Typical metal concentrations in contaminated environments.
River Godavari, India
paper plant
233.41 pg/L
157.91 pg/L
Sudhakar et al. 1991
Otago Harbor, New Zealand
tannery effluent
7000 pg/g
Johnson etal. 1981
California coast sediment
1317 pg/g
Mearns & Young 1977
Tennessee Valley (wastestreani)
leaking coal slurry
6900 pg/L
9300 ng/L
Douglas 1992
Clark Fork River sediment,
nonpoint sources
100 pg/g
Moore et al. 1988
Ninemile Creek sediment, N.Y.
chemical effluents
5.45 pg/g
Barkay & Olson 1986
Ley Creek sediment, N.Y.
chemical effluents
4.25 pg/g
Barkay & Olson 1986
Quebec, Canada
sewage sludge
879.50 pg/g
193.5 pg/g
762.5 pg/g
Couillard & Zhu 1992
Silver Valley, Idaho
mining activity
197.64 pg/g
1.28x10"' pg/g
2.2x10'' i^g/g
Roane & Kellogg 1994
Tucson, Arizona (soil)
aircraft smelting
55 pg/g
Roane & Pepper 1999
of the children tested living in the Kellogg area had abnormally high blood lead levels
(Mink, Williams and Wallace, 1971; Keely et al., 1976; von Lindem, 1981).
Elevated metal concentrations in the environment have wide-ranging impacts on
animal, plant, and microbial species. For example, human exposure to a variety of metals
can cause symptoms such as hypophosphatemia, heart disease and liver damage, cancer,
neurological disorders, central nervous system damage, encephalopathy and paresthesia
(Carson, Ellis and McCann, 1986; Hammond and Foulkes, 1986). Plant exposure to metals
is the cause of most morphological and mutational changes observed in plants (Brooks,
1983). These include shortening of roots, leaf scorch, chlorosis, nutrient deficiency and
increased vulnerability to insect attack (Caimey et al, 1979; CarHsle et al, 1986).
Likewise, microbial growth is often slowed or inhibited completely in the presence of
excessive amoxmts of metals (Duxbury, 1981; Baath, 1989).
Some plant and microbial species have developed unique and sometimes high
tolerance for metals. Plant species of Agrostis, Minuartia, and Silene are known for their
tolerance of heavy metals (Sieghardt, 1990; Verkleij et a/., 1991). In a study by Sieghardt
(1990), leaves of Minuartia vema were able to accumulate, on average, 2900 mg zinc/kg.
Silene vulgaris, on the other hand, accumulated up to 1000 mg lead/kg in root tissue.
Similarly, microorganisms have developed a variety of strategies to deal with high metal
concentrations in the enviromnent These include binding of metals to the cell surface or
cell wall, translocation of the metal into the cell, and metal transformations including
precipitation and volatilization (Hughes and Poole, 1989; Francis, 1990).
It is the
observation of these resistance phenomena in plants and microorganisms that has helped
lead to the development of biologically based metal-remediation strategies.
Perhaps the most difficult inherent problem in metal remediation is that metals,
although they may be released in the breakdown of a metal-containing compound, are not
degradable in the same sense as carbon-based molecules. The metal atom is not the major
building block for new cellular components, and while a significant amount of carbon is
released to the atmosphere as CO2, the metal atom is often not volatilized. Both
incorporation into cell mass and volatilization facilitate carbon removal firom environmental
In contrast, metals, unless removed completely from
a system through
intervention, will persist indefinitely.
Physical and Chemical Remediation of Metal-Contaminated Sites
The following sections review the traditional physico-chemical approaches that
have been used to remediate metal-contaminated sites and wastestreams.
It is the
complications and cost involved in use of such traditional remediation technologies that has
led to interest in and development of alternative remediation approaches.
Soils. Traditional metal remediation technologies have typically involved physical
removal by excavation and transport of contaminated soils to hazardous waste landfills (see
Fig. 1-2). This is thought by some to be the best method for metal reclamation (Gieger,
Federer and Sticher, 1993). However, the increasing cost of excavation and transport and
shrinking available landfill space make alternative options attractive. Two alternative
strategies for remediatioii of metal-contaminated sites are immobilization and metal
removal by soil washing, or pump and treat technologies. Both of these strategies use or
depend on pH to influence metal solubility.
Metal solubility in soil generally decreases with increasing pH.
Thus, pH can be used to immobilize metals, effectively making them less toxic and
preventing their movement into noncontaminated areas. As pH and cationic exchange
capacity is increased, the electrostatic attraction between metals and soil constituents,
including soil particles and organic matter, is enhanced (Sposito, 1989). As a result basic
soils are characterized by precipitated calcic and phosphoric metal-containing minerals.
Similarly, in some situations, liming can be used to increase soil pH causing precipitation of
soluble contaminating metals. Addition of organic matter can be used to aid this process by
acting as a sorbent for free metal in the soil matrix.
Metal Removal. Metal removal can be achieved by in situ or ex situ soil washing
which is used primarily for surface soil contamination, or by pump and treat technology
which is used at sites with deep soil or aquifer contamination. Removal by soil washing or
pump and treat is often difficult and time-constmiing because soils and sediments sorb
metals so strongly because of interactions with colloids such as humics and negatively
charged clays. Therefore, approaches have been developed to enhance removal during the
washing/flushing process (Fig. 1-2).
Soil washing with acidic solxitions is one way to facilitate metal removal. Under
acidic conditions sorbed metals are released as the increase in hydrogen ions causes
competition for available phosphate and results in formation of phosphoric acid. The
released metals have increased solubility and thus, are more easily removed from the
system during treatment.
For example, Tuin and Tels (1991) describe the use of
concentrated acid or phosphate solutions to wash metal-contaminated soil. The metals in
wash effluents are then extracted by complexation with added resins. A second approach to
recovery of metals is to use a flocculant to separate the metals from soil particles following
an acid wash. The metals are then concentrated and recovered by sodixmi hydroxide
As an alternative to acid washing, soils can also be flushed with chelating agents.
Examples of effective chelating agents include ethylenediaminetetraacetic acid (EDTA) and
nitrilotriacetic acid (NTA) both of which readily bind and solubilize metals. Using this
approach, Peters and Shem (1992) have recently reported on the removal of lead from a
contaminated soil. In this study, O.IM EDTA was able to remove 60% of the lead in a soil
containing 10,000 mg lead/kg.
Metal reclamation of sediments uses many of the same approaches as
for soils, except that sediment access is often more difficult Once removed from the
bottom of a lake or river, sediments can be treated and replaced, or landfiUed in a hazardous
waste containment site. The actual removal of sediments involves dredging. This can pose
serious problems since dredging includes the excavation of sediments from
anaerobic conditions to more atmospheric oxidizing conditions. This can result in increased
solubilization of metals, along with increased bioavailability and potential toxicity, and the
risk of contaniinant spreading (Moore, Fickiin and Johns, 1988; Jorgensen, 1989; Moore,
1994). There are ongoing discussions as to whether it is more detrimental to remove
sediments, whether for treatment or removal, or to simply leave them in place.
Aquatic Systems. Metal removal from surface water, groundwater or wastewater
streams is more straightforward than from soils.
Typically, removal is achieved by
concentration of the metal within the wastestream using flocculation, complexation, and/or
precipitation. For example, the use of lime or caxistic soda will cause the precipitation and
flocculation of metals as metal hydroxides. Alternatively, ion exchange, reverse osmosis,
and electrochemical recovery of metals can be used for metal removal (Chalkley et ai,
1989; Moore, 1994). Uiifortunately, these techniques are expensive, time-consuming and
sometimes inelBfective depending on the metal contaminant present
Metal Speciation and Bioavailability
An important aspect of metal-microbe interactions but one that is rarely addressed is
metal speciation and metal bioavailability. It is the metal species present and their relative
bioavailability rather than total metal concentration in the environment that determines the
overall physiological and toxic eflFects on biological systems (Bemhard, Brinckman and
Sadler, 1986; Hughes and Poole, 1989; Morrison, Batley and Florence, 1989).
Metal Speciation. The speciation of a metal in any environment is a result of the
combined effects of pH, redox potential and to a somewhat lesser extent, ionic strength
(Sposito, 1989; AUoway, 1990; Brierley, 1990; Moore, 1994). At high pH, metals are
predominantly found as insoluble metal mineral phosphates and carbonates while at low pH
they are more commonly found as free ionic species or as soluble organometals.
Besides pH, redox potential also influences speciation. While the redox potential of
an environment is determined primarily by environmental factors, microorganisms and their
metabolic acti\ities play essential roles in establishing redox potential as well. Redox
potential is established by oxidation-reduction reactions in the environment, reactions that
are particularly slow in soils. Reduced (anaerobic) conditions (negative E^) found in
saturated soils often result in metal precipitation due, in part, to the presence of carbonates,
ferrous (Fe^") ions and microbial reduction of sulfate to sulfides by sulfate reducing bacteria
such as Desulfovibrio. Under these conditions metals may combine with such sulfides (S^")
to form nontoxic, insoluble sulfide deposits. Under oxidizing conditions (positive Eh),
metals are more likely to exist in their free ionic form and exhibit increased water solubility.
In addition, pH may decrease slightly or dramatically under oxidizing conditions, the
classical example being acid mine drainage, where suLfiir is oxidized by Thiobacillus
thiooxidans to sulfuric acid resulting in pH values < 2.
The susceptibility of metals to changes in pH and redox demonstrates how
important it is to clearly define the chemical parameters of metal-contaminated systems.
System pH, redox potential, and ionic strength will strongly influence the success of
remediation. As an aid to prediction of metal speciation in the environment, there are
several geochemical equilibrium speciation modeling programs which may be used in
laboratory and natural settings, such as MINTEQA2 (EPA, Wash., D.C.) and PHREEQE
(National Water Research Unit, Ontario, Canada). These models predict metal distribution
between the dissolved, adsorbed and solid phases under a variety of environmental
Bioavailability. In contrast to metal speciation, metal bioavailability is determined
by the solubUity of metal species present and the sorption of metal species by solid surfaces
including soil minerals, organic matter, and colloidal materials (Babich and Stotzky, 1980,
1985; Bemhard et al, 1986; Morrison et al., 1989; AUoway, 1990). For the purposes of
this chapter, bioavailability is defined as metals in solution that are not boxind to solid phase
particles. Organic matter is a significant source of metal complexation, especially in soils
(Stevenson, 1982; AUov^^ay, 1990). Living organisms, organic debris and htimus sorb
metals reducing metal solubility and bioavailability. Organic matter consists of himiic and
non-humic material. Non-humic substances include amino acids, carbohydrates, organic
acids, fats and waxes which have not been chemically altered from their original biological
form (Alioway, 1990). Humics comprise high molecular weight compounds altered from
their original structure. Anionic functional groups on such compounds, including carboxyl,
carbonyl, phenohc, hydroxyl and ester groups, bind cationic metals sequestering metal
activity. Some organic complexing agents form soluble complexes with metals while
others form insoluble complexes. For instance, amino acids, simple aliphatic acids and
other microbially-produced agents can form soluble chelates. The formation of these
soluble structures is of concem since such structures are prone to transport and may result in
contaminant spreading. Insoluble complexes form when metals bind to high molecular
weight humic materials. In this case, toxic metal concentrations may be reduced to
nontoxic levels.
Metal bioavailability is generally increased with decreasing pH. This is due to the
presence of phosphoric, sulfuric and carbonic acids, which increasingly solubilize organicand particulate-bound metals. Particulate-boiind metals are considered those boimd to
secondary minerals, for example, clays, iron and aluminum oxides, carbonates, sulfidic and
phosphoric minerals.
Due to the heterogeneous nature of soils and sediments, wide
fluctuations in pH can exist in a given environment For instance, metals may be more
soluble in surface layers where plant exudates, microbial activity, moisture and leaching
lower pH.
The toxicity of a metal to a biological system is dependent upon metal
bioavailability. One problem with metal remediation is that metals in the environment are
subject to chemically and biologically mediated oxidation-reduction reactions which can
alter both metal speciation and bioavailability. Generally, the most dangerous inorganic
form of a metal is the metal cation (McLean and Beveridge, 1990), a metal species that has
greater solubility with decreasing pH. As the solubiHty of the metal increases, metal
toxicity increases due to enhanced mobility and bioavailability to biological systems. In
general, metals are more soluble when oxidized than reduced and are thus more likely to
exist as free ionic species. For example, trivalent chromium, Cr(T[I), is highly insoluble and
poses litde threat to the health of an environment. Environmental oxidation of Cr(III),
however, results in the release of the highly toxic and soluble chromate ion, Cr(VI).
Metal Toxicity to Microorganisms and Microbial Resistance Mechardsms
Metals are essential components of microbial cells, for example, sodium and
potassium regulate gradients across the cell membrane, and copper, iron and manganese are
required for activity of key metalloen2ymes in photosynthesis and electron transport.
However, metals can also be extremely toxic to microorganisms, impacting microbial
growth, morphology and biochemical activities as a result of specific interactions with
cellular components (Foster, 1983; Gadd, 1986; Beveridge and Doyle, 1989; Gadd, 1992;
Freedman, 1995). Perhaps the most toxic metals are the nonessential metals such as
cadmium, lead, and mercury.
Mechanisms of metal toxicity differ. Toxicity may occur as a result of binding of
the metal to hgands containing reactive sulfliydryl, carboxyl or phosphate groups such as
proteins or nucleic acids. The larger metal ions such as the mercury or cadmium cations
readily bind sulfliydryl groups, while smaller more highly electropositive metal ions such as
tin react with carboxyl and phosphate groups. Other interactions which cause inhibition of
cell growth include metal catalyzed decomposition of essential metabohtes and analog
replacement of
structurally important cell components.
A good example of analog
replacement is arsenic which is bactericidal because it acts as an analog of phosphate
disrupting nucleic acid stmcture and enzyme action. Photosynthetic and nitrogen fixing
organisms are particularly sensitive to metals, and high concentrations of lead and nickel
retard cell division. Another sensitive cell component is the cell membrane which is
susceptible to disruption by copper and zinc (Baath, 1989).
As a consequence of, or perhaps in spite of, metal toxicity, some microorganisms
have developed various resistance mechanisms. The strategies are either to prevent entry of
the metal into the cell or to actively pump the metal out of the cell. This can be
accomplished by either sequestration, active transport or chemical transformation through
metal oxidation or reductioiL
Sequestration. Sequestration involves metal complexation with microbial products
such as extracellular polymeric substances (EPS) and metallothionein-like proteins.
Sequestration may also involve the binding to electronegative components in cellular
membranes. Regardless of the agent, the goal is to reduce or eliminate metal toxicity via
complexation. This may be accomplished by binding the metal extracellularly to prevent
metal entry into the cell, as is the case with exopolymers and, to some extent, cell surface
binding (Rudd, Sterritt and Lester, 1984a,b; Stupakova, Demina and Dubinina, 1988;
Beveridge and Doyle, 1989; Ehrlich, 1990; Gadd, 1992; Gr^peli et al, 1992; Shuttleworth
and Unz, 1993). Alternatively, metals may enter a cell and be concentrated there as a result
of passive difEiision or by active transp>orL Psendomonas aeruginosa has been shown to
accumulate uranium by both passive difilision and, in some instances, by a metabolismdependent translocation process (Strandberg, Shumate and Pairott, 1981; Hughes and
Poole, 1989).
Cadmium uptake in Stcq)hylococcus aureus occxars via the manganese
transport system (Silver, Misra and Laddaga, 1989). Following metal uptake, some cells
sequester metals intracellularly utilizing low molecular weight, cysteine-rich proteins called
metallothioneins. Such proteins, initially discovered in fimgi, rapidly bind metals as they
enter the cell effectively reducing their toxicity (Robinson and Tuovinen, 1984; Gadd,
1990b). The complexed metal may then either be transported back out of the cell or stored
as intracellular granules.
Which mechanism predominates, intracellular or extracellular sequestration, is
dependent on the organism involved. For example, uranium was found to accumulate
extracellularly as needle-like fibrils in a layer approximately 0.2 (im thick on the surface of
a yeast, Saccharomyces cerevisiae, but formed dense intracellular deposits in Pseudomonas
et aL, 1981).
Active Transport As already mentioned, active transport of metals out of the cell is
one mechanism of microbial resistance to metal toxicity. Highly specific efflux systems
can rapidly pump out toxic metal ions that have entered the cell. Such efflux pumps may
derive their energy firom membrane potential or from ATP, and in fact ATP-dependent
efflux pimips have been identified that are specific for arsenic, cadmium and chromium
(Horitsu et al, 1986; Laddaga et al., 1987; Hughes and Poole, 1989; Rani and Mahadevan,
1989; Nies, 1992; Silver, 1992). For cadmium, a plasmid gene, cadA, encodes a cadmium
specific ATPase that transports cadmium out of the cell. This gene has been found in
several genera including
Staphylococcus aureus, Pseudomonas putida, and Bacillus
Similarly, Escherichia coli xitilizes a plasmid-encoded arsenic ATPase for
arsenate efflux (Silver and Misra, 1988).
Qxidation-ReductiorL Microbially-facilitated oxidation-reduction (redox) reactions
are a third mechanism for microbial metal resistance. Microorganisms may either oxidize
to mobilize or reduce to immobilize a metal and both reactions are used to prevent metal
entry into the cell. For example, the reduction of mercuric ion, Hg(II), to methyl mercury
volatilizes the mercury which can then difiiise away firom the organism (Robinson and
Tuovinen, 1984; Beveridge, 1989; Beveridge and Doyle, 1989; Silver et ai, 1989; Gadd,
Other toxic metals imdergo microbial reduction.
Microbial hydrogen sulfide
production is widespread and confers resistance through the precipitation of insoluble metal
sulfides (Gadd and Griffiths, 1978; Ehrhch, 1990). Enzymatically-produced phosphate, for
example, by Citrobacter spp., has been shown to precipitate and detoxify lead and copper
(Aiking et ai, 1984; Brierley, 1990). The ubiquitous nature of microorganisms and their
abihty to tolerate a variety of anthropogenic and environmental stresses, including heavy
metals, have made them primary candidates for exploitation in the remediation of metalcontaminated systems. As previously mentioned, some plants, especially of the genera
Silene and Agrostis, are particularly metal-resistant, and, it should be noted that plants may
be used in the bioremediation of metal-contaminated systems. However, the role of plants
in metal reclamation will not be discussed in this chapter. Several excellent reviews on
plant-facilitated removal of metals are available (Mitsch and Jorgensen, 1989; Otte, 1991;
Verkleij etal, 1991).
Origin of Microbially Based Metal Remediation
Microbial systems have long been used to efiBciently. treat domestic waste, and to
recover precious metals fi-om ores. It is therefore logical to extend the application of
microbial systems for use in remediation of metal-contaminated soils and wastestrearos.
Perhaps the inception of microbially-based metal remediation methods was in the 18th
century in Rio Tinto, Spain, during the extraction of copper from copper ore. Although
unknown at the time, copper solubilization during the extraction process was a result of
microbial activity. The use of microorganisms ia the recovery of precious metals, such as
copper and gold, was later coined biohydrometallurgy. An efficient and cost-effective
process, by 1989, more than 30% of U.S. copper production utilized biohydrometallurgy,
the key component of which was the microorganism Thiobacillus ferrooxidans (Debus,
Bioleaching generally involves the oxidation of iron or sulfur-containing minerals.
Three major organisms with unique capabilities are involved in the bioleaching process.
Thiobacillus ferroxidans is thought to be the dominant organism in acidic environments
where it oxidizes iron sulfides such as bomite (CujFeSJ and pyrite (FeSj). In the oxidation
of ferrous iron (Fe^"^ to ferric iron (Fe^"^ by Thiobacillus ferrooxidans, the produced ferric
ion is able to oxidize sulfide (S^') to sulfate (S04^') resulting in the production of sulfuric
acid. The sulfuric acid lowers the pH of the environment and facilitates additional leaching.
Thiobacillus thiooxidans, on the other hand, is unable to oxidize iron but can readily
oxidize other minerals where leaching is dependent on elemental sulfur oxidation such as
zinc stilfides. Leptospirillum ferrooxidans is also commonly associated with bioleaching.
This organism oxidizes pyrite under acidic conditions, however, Leptospirillum
ferrooxidans carmot oxidize sulfur limiting its ability to solubilize certain mineral sulfides.
such as chalcopyrite (CujS). The discovery of these microorganisms and others has led to
increased interest in metal-microbe relations (Lodi, Del Borghi and Ferraiolo, 1989; Ehrlich
andBrierley, 1990).
Microbial Interactions with Metals
Francis (1990) has summarized the numerous possible microbially-mediated
reactions resulting in the mobilization or immobilization of metals and found that major
interactions include oxidation-reduction processes; biosorption and immobilization by cell
biomass and exudates; and mobilization by microbial metabolites. A profoimd issue iu
metal remediation is that through microbial action, metals can readily be re-mobilized
creating toxicity issues in sites where metals are not completely removed.
Oxidation-Reduction Reactions.
As briefly discussed, microorganisms can
decrease metal toxicity by the oxidation or reduction of metals. Some microorganisms
actively reduce metals to decrease bioavailability while others may oxidize metals to
facilitate their removal from the environment. Early laboratory studies by Konetzka
(1977) found Pseudomonas spp. capable of the aerobic oxidation of arsenite, As(0H)3, to
the less toxic arsenate, As(0)(0H)3. Selenite, Se'"', and selenate, Se^, reduction under
aerobic conditions to elemental selenium, Se°, has been observed in certain fimgi as a
mechanism for tolerance (Konetzka, 1977; Gharieb, Wilkinson and Gadd, 1995). Tomei et
al. (1995) found Desulfovibrio desulfiiricans anaerobically reduced selenate and selenite to
elemental selenium as a means of selenium detoxification. Bacterial reduction of chromates
(CrVI) to less toxic and less soluble Cr(III) under aerobic conditions by a chromiumresistant Pseudomonas fluorescem, isolated jfrom chromium-contaminated sediments from
the Hudson River, N.Y., has been reported by Bopp, Chakrabarty and Ehrlich (1983).
Similarly, Ishibashi, Cervantes and Silver (1990) observed chromium reduction in
Pseudomonas putida using a soluble chromate reductase.
In some cases, metals are solubilized microbiologically through oxidation to
facilitate metal removal from a contaminated system. With some metals, toxicity increases
with oxidation since bioavailability is also increased. The greater mobility of metals in
these circumstances and the likelihood of removal from a microorganism's habitat acts as a
mechanism of metal tolerance. As previously mentioned, Thiobacillus spp. and Leptothrix
spp. readily solubili2e a variety of metals via oxidation, including manganese, uranium and
copper (Ehrlich and Brierley, 1990). Bacillus megaterium has been shown to oxidize
elemental selenium to selenite increasing selenium mobilization (Sarathchandra and
Watldnson, 1981). Francis and Dodge (1990) have reported that a Nj-fixing Clostridium
sp. is capable of anaerobic solubilization of cadmiirai, chromium, lead and zinc.
What is apparent in the examination of microbial metal redox reactions is that while
one microorganism can immobilize a metal, another is capable of solubilizing the metal.
This can lead to inadequate metal detoxification and complexation in systems where metal
removal is not performed.
Complexation. Bacteria, algae, fimgi and yeasts have all been found to complex or
sorb metals (Gadd, 1990b), however, bacteria have been most extensively studied.
Complexation of metals by microorganisms occurs in two ways: 1) the metals may be
involved in nonspecific binding to cell wall surfaces, the slime layer, or the extracellular
matrix, or 2) they may be taken up intracellularly. Studies have shown both types of metal
complexation are used to reduce metal toxicity and mobility (Bitton and Freihefer, 1978;
Scott and Palmer, 1988; Gadd, 1990b; Marques et al, 1990; Roane, 1994; Roane and
KeUogg, 1994).
There are numerous studies of metal complexation by whole cells indicating that
complexation depends both on the bacterium, the metal, and pH. It has been found that at
low pH, cationic metal complexation is reduced, however, binding of anionic metals such as
chromate (Cr04^") and selenate (Se04^") is increased. Bacteria seem to show selective
affinities for different metals. This was demonstrated by a study of the complexation of
four metals by four different bacterial genera that showed that the affinity series for
bacterial removal of the metals studied decreased in the order Ag> La> Cu>Cd (MtiUen et
al., 1989). Of the bacteria studied, Pseiidomonas aeruginosa was the most efficient at
metal complexation while Bacillus cereus was least. Complexation capacity is reported to
be species specific as well as genus specific. A study of metal complexation between
different Pseudomonas sp. showed up to one order of magnitude difference in complexation
of several metals includiag lead, iron, and zinc (Corpe, 1975). The efficiency of metal
complexation by microorganisms has resulted in the development and use of several
bioremediation processes (Brierley, 1990).
The presence of soil complicates metal removal because soils sorb metals strongly
and can also affect microbial-metal complexation.
Walker et al. (1989) showed that
purified preparations of cell walls fi:om Bacillus subtilis and Escherichia coli (423 to 973
mmol metal/g cell wall) were more effective than either of two clays, kaolinite (0.46 to 37
mmol metal/g clay) or smectite (1 to 197 mmol metal/g clay), in the binding of seven
different metals. However, in the presence of cell wall/clay mixtures, binding was reduced.
In summary, there are several parameters that affect metal complexation. These include
specific surface properties of the organism, cell metabolism, metal type, and the
physicochemical parameters of the environment.
Methylation of metals generally results in the volatilization and
increased toxicity of a metal. The addition of methyl or alkyl groups on metals increases
their lipophilicity and their permeability across biological membranes (Hughes and Poole,
1989). Such conversions firom inorganic to organic metallic
are of considerable
interest fi"om an ecological standpoint since toxicity and mobilization of the metal is
modified. Methylation of the mercuric ion, Hg^% results in formation of monomethyl,
CHsHg", or dimethylmercur>', (CH3)2Hg, which can be 10 to 100 times more toxic than
inorganic mercury. As seen with mercury, however, although volatilization increases metal
toxicity, microorganisms use volatilization to facilitate metal diffusion away fi'om their
sim:oundings, thus, avoiding the toxic effects. In addition to mercury, other elements
including tin, lead, arsenic and seleniiun are among the metals commonly methylated by
microorganisms (Reamer and Zoller, 1980; Thayer and Brinckman, 1982; Barkay and
Olson, 1986; Gilmonr, Tuttle and Means, 1987; Barkay et al, 1992). The volatilization of
metals can be a significant source of metal loss in estuarine and freshwater sediments, as
well as heavily contaminated soils and sewage (Robinson and Tuovinen, 1984; Compeau
and Bartha, 1985; Lester, 1987; Barkay, Liebert and Gillman, 1989). Smdies conducted by
Reamer and Zoller (1980) and Frankenberger and Karlson (1995) have observed selenium
methylation in soils, sediments and sewage sludge. Saouter, Turner and Barkay (1994)
found that microbially-induced mercury volatilization can be a significant mechanism of
removal from a mercury-contaminated pond. Both studies found biomethylation was part
of the detoxification process for microorganisms.
Biosurfactants and Siderophores. Although their role in nature is still not clear,
there are extracellular microbial products, such as biosurfactants, bioemulsifiers and
siderophores, that complex or chelate metals quite efficiently.
Examples of these
compounds are shown in Fig. 1-1.
There are several reviews that discuss biosurfactants and their properties and
^plications (Zajic and Sefifens, 1984; Rosenberg, 1986; Lang and Wagner, 1987; Miller,
1995b). It has been suggested that biosurfactants may play a role in the adhesion and
desorption of microbial cells to surfaces (Rosenberg, 1986). In addition, biosurfactants
have been shown to enhance the bioavailability of hydrocarbons with low water solubilities,
thereby stimulating growth and biodegradation of such hydrocarbons
and Miller,
1994; 1995). Recent work has suggested that in addition to these roles, biosiarfactants
complex metals very efBciently. A rhamnolipid biosurfactant (see Fig. 1-1A) has been
shown to complex metals such as cadmitim, lead, and zinc with a complexation capacity
comparable to those reported for exopolysaccharides (Miller, 1995a). The stability constant
(log K = -2.47) was higher than those reported for cadmium-sediment (-6.08 to -5.03) and
cadmium-humic acid systems (-6.02 to -4.92) indicating that the cadmium-rhamnoUpid
complex was much stronger than the cadmium-sediment or cadmium-humic acid complex
(Tan era/., 1994).
Similarly, Zosim, Gutnik and Rosenberg (1983) have reported uranium binding by
emulsan (up to 240 mg uranium (UOj^Vnig emulsan), a bioemulsifier produced by
Acinetobacter calcoaceticiis RAG-l. The intriguing aspect of this system is that in a
hexadecane-water solution, the emulsan preferentially binds to the hexadecane-water
interface which effectively concentrates the complexed uranium for easy recovery.
Siderophores are chelating agents often produced under iron-limiting conditions.
They contain reactive groups such as dicarboxylic acids, polyhydroxy acids and phenolic
compounds. As shown in Fig. 1-lB, siderophores complex metals and they have been
reported to facilitate movement of metals in soil (Duff, Webley and Scott, 1963; Bolter,
Butz and Arseneau, 1975; Cole, 1979). In addition to iron, siderophores can complex
gallium, chromium, nickel, uranium, and thorium (Hausinger, 1987; Macaskie and Dean,
Figure 1-1. (a) Rhamno lipid from P. aeruginosa ATCC 9027 showing cadmium binding,
(b) Structure of the iron-siderophore complex of enterobactin.
\Ietai Removal
Metal Immobilization
metal soiufailizadoa
e.g.. acid leaching,
metal precipitation
e.g.. phosphate, carbonate,
suifidic minerals
replacement of reclamed soil]
monitor status of soiuble metal
metal-containina effluent
Coataminated soil
Metal Removal
metal oxidadon & bacterial leaching
e.g.. acid production, cheladon.
Metal Immobilizadon
bacterial binding
e.g.. EPS, phosphate productionmetal reducrion
monitor soluble metal status
metal-containing effluent
Figure 1-2. Remediation strategies for soils and sediments.
However, limited progress has been made regarding the role of siderophores in metal
recovery and/or removal.
Innovative Approaches to Microbial Remediation of Metal-Contaminated Environments
Current approaches to metal bioremediation are based upon the complexation,
oxidation-reduction, and methylation reactions just discussed. Until recently, interest was
focused on technologies that could be applied to achieve in situ immobilization of metals.
the last few years, the focus has begun to shift toward actual metal
removal because it is difficult to guarantee that metals will remain immobilized indejSnitely.
Numerous bioremediation strategies have been suggested to have potential
applicability for removal of metals from contaminated environments. Unfortunately, the
number of field-based studies have been few for several reasons. Often the institutions that
perform the basic research and feasibiKty studies are not equipped to gear up to field
application making this a difiScult and expensive transition. Second, it is sometimes
difficult to change entrenched attitudes which in the past have been favorable toward
excavation and removal of contamination and unfavorable toward bioremediation
processes. Finally, bioremediation has not always been a predictable technology, however
as the field record of success grows, it is to be expected that acceptance of bioremediation
as a viable technology wUl also grow.
Soils and Sediments
Microbial Leaching. Microbial leaching of metals from metal-containing soils is an
extension of the practice of bioleaching of metals from ores (Lodi et al, 1989; Fig. 1-2). In
bioleaching, copper, lead, and zinc have successfiilly been recovered through metal
solubilization by Thiobacillus ferrooxidans and Thiobacillus thiooxidans. This process has
also been used to leach uranium from nuclear waste contaminated soils (Hutchins et al,
1986; McCready and Gould, 1990; Rossi and Ehrlich, 1990; Macaskie, 1991), and to
remove copper in the bioremediation of copper tailings (Gokcay and Onerci, 1994). In a
recent study by Phillips, Landa and Lovley (1995), uranium was removed from
contaminated soil using bicarbonate. The uranium was then precipitated out of the soil
effluent using Desulfovibrio desuljuricans to reduce U(VI) to U(IV) with removal
efficiencies ranging from 20-100%.
While used extensively in mining and metal recovery, microbial leaching of metals
has, unfortunately, received little attention as a microbial-based technology for metal
remediation. One potential application is the treatment of sewage sludge earmarked for
disposal in soil.
Sludge-amended soils have increased plant-available nutrients and
improved productivity, but, unfortunately, sludge addition to soils also increases metal
content (Reimers, Akers and White, 1989; Sauerbeck, 1991). Concem over the metallic
and organic contaminants in sludge has limited its application despite the benefits of
nutrients and organic matter. This concern, has, in turn, stimulated interest in microbial
removal of heavy metals from sewage sludge-amended soils. Thiobacillus ferrooxidans
and Thiobacillus thiooxidans have been used to leach metals from contaminated sludge
before soil application. This process was successful in making the sludge suitable for
agricultural application (Couillard and Zhu, 1992).
The use of bacterial surfactants for metal
remediation of soils is gaining attention. These molecules (see Fig. 1-1A) are water soluble
and have low molecular weights (»500- to 1500). Above a critical micelle concentration
(CMC) biosurfactant monomers aggregate to form structures such as micelles, vesicles and
bilayers. Because of their small size, biosurfactant monomers as well as smaller aggregate
structures, as in micelles and small vesicles, may freely move through soil pores. It was
estimated in one study of a rhamnolipid biosurfactant and cadmium that the biosurfactantmetal complexes ranged up to 50 nm in diameter (Champion et al, 1995). Particles of that
size should not be removed by physical straining in most soils (Gerba, Yates and Yates,
Bacterially-produced surfactants consist of a variety of molecular structures which
may show metal specificity and, thus, may be optimized for a particular metal (Miller,
1995a). Herman, Artiola and Miller (1995) have demonstrated removal of cadmium (56%)
and lead (42%) from soil by a rhamnolipid biosurfactant produced by Pseudomonas
aeruginosa. These studies indicate the potential for metal-containing soils to be flushed
with a surfactant-containing solution thereby causing metal desorption and removal. The
use of biosurfactants is a promising treatment of soil systems because it offers the
possibility of using an in situ treatment,
the need for excavation and external
washing (see Fig. 1-2).
bioemulsifiers, such
as emulsan
calcoaceticus, have been shown to aid in removal of metals. Potential for remediation of
soils using bacterial exopolymers is indicated by a study which showed that purified
exopolymers from 13 bacterial isolates removed cadmium and lead from an aquifer sand
with efficiencies ranging from 12 to 91% (Chen et aL, 1995). Although such molecules
have much larger molecular weights (—10®) than biosurfactants, this study showed that
sorption by the aquifer sand was low suggesting that in a porous medium with a sufficiently
large mean pore size use of exopolymers may be feasible.
Volatil1 ration. Although methylation is not a desirable reaction for most metals
particularly arsenic or mercury, methylation and volatilization have been proposed as
techniques for remediation of selenium contaminated soils and sediments (Huysmans and
Frankenberger, 1990; Karlson and Frankenberger, 1990; Frankenberger and Karlson, 1995).
Seleniiim-contaminated soils from San Joaquin, CA, were remediated using selenium
volatilization stimulated by the addition of pectin in the form of orange peel. In their study,
Karlson and Frankenberger (1989) found that the addition of pectin enhanced the rate of
selenium alkylation, ranging from 11.3 to 51.4% of added selenium, and that this was a
feasible approach to treat soils contaminated with selenium.
Aquatic Systems
Metal reclamation from
acid mine drainage, and contaminated surface and
groundwater and wastewaters has been extensively studied.
Technologies for metal
removal from solution are based on the microbial-metal interactions discussed earlier: the
binding of metal ions to microbial cell surfaces; the intracellular uptake of metals; the
volatilization of metals; and the precipitation of metals via complexation with microbiallyproduced ligands.
Acid Mine
Drainage. Acid
drainage and mining effluent waters containing
high amounts of zinc, copper, iron, manganese, as well as lead, cadmium and arsenic, are
commonly remediated using wetlands. Wetland treatment of acid mine drainage has proved
to be cost-effective and less labor intensive than traditional chemical treatments. The
Teimessee Valley Authority reports an 80% success rate in meeting discharge standards
using wetiands alone to remediate acid drainage (Douglas, 1992). Wildeman et cd. (1994)
reported the following metal reductions in contaminated water in response to wetland
treatment: zinc, 150 to 0.2 mg/L; copper, 55 to <0.05 mg/L; iron, 700 to 1 mg/L; and
manganese, 80 to 1 mg/L.
Wetland remediation involves a combination of interactions including microbial
adsorption of metals, metal bioaccumulation, bacterial oxidation of metals, and sulfate
reduction (Fermessy and Mitsch, 1989; Kleinmann and Hedin, 1989). Sulfate reduction
produces sulfides which in turn precipitates metals and reduces aqueous metal
concentrations. The high organic matter content in wetland sediments provides the ideal
environment for sulfate-reducing populations and for the precipitation of metal complexes.
Some metal precipitation may also occur in response to the formation of carbonate minerals
(BCleinmarm and Hedin, 1989). In addition to the aforementioned microbial activities,
plants, including cattails, grasses, and mosses, serve as biofilters for metals (Brierley,
Brierley and Davidson, 1989).
While proving to be an effective method of treatment for a number of metals,
wetlands may need additional inputs of organic matter which can be added in the form of
mushroom compost, as an example. Wetland technology is limited by the problems of
buildup and disposal of contaminated biomass, plant and microbial, and by the possible
toxic effects of the metal-containing influent upon wetland plant and microbial
communities. In response, some researchers are investigating the possibility of treating acid
mine drainage biologically without wetlands.
Surface and Groundwater. Microbial biofilms are a common treatment technology
for metal-contaminated surface waters and groundwater. Immobilized biolBlms, viable and
nonviable, essentially trap metals as the contaminated water is pumped through (Macaskie
and Dean, 1989; Summers, 1992). A variety of microorganisms have been used to create
such biofilms (Brierley et al, 1989; Gadd, 1992). For example, live immobilized biofilms
of Citrobacter spp. have been used to remove uranium firom contaminated wastewater in
both laboratory and field
studies (Macaskie, 1991).
Biofilms of Streptomyces
viridochromogens have similarly been used in uranium removal. Arthrobacter sp. biomass
and exopolymers have been implied in the capture of cadmium, chromium, lead, copper.
and zinc from fluid wastes on the laboratory scale (Gxappelli et al., 1992). Exopolymers of
Bacillus spp. have been used to remove cadmium, chromium, copper, mercury, nickel,
uranium and zinc from wastewater (Brierley et al, 1989). Fungi and yeasts have also been
shown to be effective in metal removal from aquatic systems (Gadd, 1990a). In response to
silver concentrations ranging from 0 to 2 mM and copper concentrations from 0 to 3 mM,
strains of Saccharomyces cerevisiae and Candida sp. were able to accumulate from 0.03 to
0.19 mmol silver/g dry weight and 0.05 to 0.18 mmol copper/g dry weight, respectively
(Simmons, Tobin and Singleton, 1995).
In response to the success of microbially-mediated clean-up of metal-contaminated
waters, some commercial bioremediation products such as BIOCLAIM, AlgaSORB, and
BIO-FEX are available. More detailed descriptions are provided by Brierley (1990).
BIOCLAIM and BIO-FEX use immobilized bacterial preparations while AlgaSORB utilizes
a nonviable algae matrix for metal removal. In addition to these products, there are several
proposed proprietary processes including the use of immobilized Rhizopus arrhizus
biomass for uranium recovery (Tzezos, McCready and Bell, 1989). A froth flotation
method for enhanced contact between biomass and contaminated water has been proposed
by Smith, Yang and Wharton (1988).
Similar to the microbial biofilm preparations described above, free-floating, viable
microbial mats are also successfiil in removal of metals from solution (Bender and Phillips,
1994; Vatcharapijam, Graves and Bender, 1994).
Consisting primarily of algae,
cyanobacteria and bacteria, microbial mats perform a number of activities which promote
metal complexation and subsequent removal. The mat contains oxidizing and reducing
zones that aid in the immobilization and precipitation of metals. There is an increased pH
which promotes metal precipitation as metal hydroxides, and the mat releases negatively
charged extracellular substances to promote flocculation.
Microbial mat formation may also stimulate metal removal through sulfate
reduction. Bames, Scheeren and Buisman (1994) have developed a process that specijBcally
uses sulfate-reducing bacteria to treat metal-contaminated groundwater. In this process, as
groundwater is pumped through the water treatment plant, sulfide produced by sulfatereducing bacteria precipitates the metals in the water. Metal concentrations in the treated
water was reportedly reduced to \is/L quantities and was suitable for release into the
Marine Water. Few studies have evaluated the potential for use of microorganisms
in the remediation of sea water, however, the problems encountered are similar to those of
other aquatic systems. Stupakova et al. (1988) have proposed the use of the marine, bacteria
Deleya venustus and Moraxella sp. for copper uptake from sea water. Additionally, Corpe
(1975) performed metal-binding studies with copper using exopolymer from film-producing
marine bacteria and fotmd that insoluble copper precipitates formed effectively decreasing
copper toxicity.
Wastewater. Metal removal from domestic waste is a better studied system. The
efi5ciencies of metal removal vary both with respect to waste and metal type (Lester, 1987).
However, removal of metals from wastewater is generally an efficient process. Cheng,
Patterson and Minear (1975) studied heavy metal uptake by activated sludge and found that
in an activated sludge effluent containing 2,100-25,200 iig copper/L, 89% of the copper was
removed during treatment. It was also observed that 98% of the lead was removed from
effluent containing 2,100-25,500 ^g lead/L. In a similar study, in effluent with 9,000 ng
zinc/L, 89% of the zinc was removed during sewage treatment (Barth et al, 1965).
Metal removal in wastewater treatment can also be done efficiently on a smaller
scale as reported by Whitlock (1990). The Homestake Goldmine in Lead, South Dakota,
utilizes a biological treatment plant to detoxify 4 million gallons of wastewater that are
discharged daily to Whitewood Creek. The wastewater matrix contains
levels of cyanide, ammonia and metals. The plant consists of a traia of five rotating
biological contactors (RBCs) which sequentially treat the wastewater. The plant had been
in successfiil operation for 9 years at the time of the report with the following results:
treatment capacity and resistance to upset improved with time, copper and iron were
removed at 95 to 98% efiflciency while nickel, chromium and zinc removal were
inconsistent, ammonia and cyanide were removed to acceptable levels, and finally,
Whitewood Creek was successfiiUy reestablished as a trout fishery.
The removal of metals from wastewater streams and sewage rehes primarily on their
immobilization and complexation of metals by extracellular compounds, e.g., EPS (Brown
and Lester, 1982a,b; Rudd et al, 1984a;). The genus Zoogloea, an important organism in
sewage treatment, readily forms an anionic slime matrix. Toxic metals complex with the
matrix and precipitate out of solution. Klebsiella aerogenes is another common sewage
bacterium that binds metal ions with extracellular polymers. Complexed metals are then
removed from the wastewater via sedimentation during the treatment process.
Other proposed mechanisms of metal removal from sewage include physical capture
by microbial floes,
cellular accumulation and volatilization by such organisms as
Klebsiella, Pseudomonas, Zoogloea, and Penicillium spp. (Brown and Lester, 1979). In
laboratory and pilot scale studies, up to 98% metal removal by these mechanisms combined
has been documented (Lester, 1987).
Nuclear Waste
More recent concem about metal release into the environment comes from the
nuclear industry. Nuclear waste can contain a combination of non-radioactive metals, such
as lead and copper, and radioactive metals including uranium (^®U), thorium (^°Th) and
radium (^®R). Nuclear waste disposal, perhaps better termed as storage, has the potential to
introduce both radioactive and non-radioactive metals into siirrounding aquifers, soils and
surface waters as a result of poor construction or placement of storage facilities. The risk of
environmental contamination via nuclear waste is not yet well understood and has received
relatively litde attention (Macaskie, 1991; Haggin, 1992).
Radioactive metal wastes from the nuclear industry are of increasing concem as the
amotmt of waste to be disposed of increases. Current treatment of nuclear wastewater
involves the addition of lime which is effective in precipitating most metals out of solution
with the exception of radium (Tse2»s and Keller, 1983). Barium chloride (BaClj) is used to
precipitate radium from sulfur rich effluents as barium-radium sulfate. Other treatment
methods include incineration for some solid wastes, filtration, adsorption and crystallization
for liquid wastes (Godbee and Kibbey, 1981).
In the removal of nuclides from contaminated systems, biological adsorbents are
superior to conventional adsorbents, such as zeolite and activated carbon. For example,
Penicilliian chrysogenum was found to adsorb up to 5x10'* nCi radixmi/g biomass from an
initial radium concentration of 1000 pCi/L compared to adsorption of only 3600 nCi/g by
activated carbon under the same conditions (Tsezos and Keller, 1983; Tsezos, Baird and
Shemilt, 1987). There are a variety of biological products including tannins, melanins,
bacterial chelating agents, microbial polysaccharides and metallothioneins, whole microbial
cells, including yeasts and fimgi, and chitin that have been investigated for application in
the bioremediation of nuclear wastes. An extensive review of biological treatment of
nuclear wastes is provided by Macaskie (1991).
Due to the danger uraniimi poses to hioman health and to its ubiquity in nuclear
byproducts, uranium is the most extensively studied radionuclide. Several microbiological
systems have been proposed for the removal of uranium. Phosphate-producing Citrobacter
spp. have been studied for the precipitation of uranium and lanthanum metals (Plummer and
Macaskie, 1990; Macaskie et al, 1992). Microbially-released products, such as emulsan
from Acinetobacter sp. and an exopolysaccharide from a Pseudomonas sp., have also been
stiggested for use in the recovery of uranium (Zosim, Gutnick and Rosenberg, 1983;
Marques et al, 1990). Tsezos, McCready and Bell (1989) reported the use of immobilized
biomass of Rhizopus arrhizus in uranium recovery from contaminated water. As more
studies are performed, microbial reclamation of these wastes will be better understood.
Concluding Remarks
Bioremediation of metal-contaminated environments is a rapidly advancing field.
The use of microbial biomass, exopolysaccharides, biosurfactants, bioemulsifiers,
siderophores, and microbially-induced oxidation-reduction and methylation reactions,
offer promising altematives to traditional technologies in the treatment of heavy metals.
Future research focused on in situ mining to help reduce contaminated waste and
development of microbial applications in metal removal from industrial and
wastes before release into the environment will provide still more treatment options. To
develop new microbial-mediated remediation technologies, more research on microbial
reduction of metals in environmental systems which increase metal mobility; the effects
of co-contaminants, including the presence of other metals, in metal reclamation; and the
speciation and bioavailability of metals in the environment, an area of utmost importance
in biological remediation, need to be conducted. Finally, more field-based studies need
to be performed since, while laboratory studies are necessary, in situ studies can provide
essential information about environmental interactions. While there is much to be done,
the future promises more iimovative applications in the field of metal bioremediation.
This section titled Microbial Remediation of Metals has been published as:
Roane, T.M., I.L. Pepper and R.M. Miller. 1996. Microbial Remediation of Metals, pp.
312-340. In Bioremediation: Principles and Applications by R.L. Crawford and D.L.
Crawford (eds), Cambridge University Press, United Kingdom.
Dissertation Format
This dissertation consists of a book chapter as represented in the above section
titied "Microbial Remediation of Metals" that is already published. In addition to the
book chapter, this dissertation consists of three manuscripts prepared for publication in
the microbiology journals as stated. While my advisor, Professor Ian L. Pepper, and I
agreed on the initial premise for this research, this research is based on my personal
experience, findings and ideas. And while science cannot be proven, I have worked to
support my hypotheses and theories as completely as possible. I take fiiU responsibility
for the content of these manuscripts. Much of my work has been enriched thanks to
extensive discussions with Professor Pepper. Karen L. Josesphson, a co-author on the
final manuscript, was instrumental in carrying out the mechanics of the field study.
Finally, both Professors Pepper and Raina M. Maier reviewed the manuscripts making
suggestions to improve them.
The development of the theory, methods, discussion and conclusions are
presented in the papers appended to this dissertation. The following is a summary of the
most important findings in these papers.
Remediation of a Co-Contaminated System
In sites contaminated with both metallic and organic pollutants, it is believed that
bioremediation is often unsuccessful as a consequence of metal toxicity. Yet, many
microorganisms have the ability to resist metal toxicity while others can degrade a variety
of organic pollutants. In this research, metal-detoxifying microbial populations and an
organic degrading population were used to remediate a co-contaminated soil. The
contaminants chosen for this research were the metal cadmium (Cd) and the herbicide
2,4-dichlorophenoxyacetic acid (2,4-D). Both represent model contaminants since much
is known about how microorganisms deal with each one.
The overall objective of this research was to evaluate how microorganisms
respond to metal stress and how microorganisms can be used to detoxify metal imder cocontaminated conditions. The specific objectives of this research were to: (1) compare
the microbial commxmity response between cadmium-contaminated and uncontaminated
soils; (2) evaluate the diversity of cadmium-resistance mechanisms in cadmium-resistant
isolates from one soil; and (3) use dual bioaugmentation to enhance 2,4-D degradation
under co-contaminated conditions.
Cfldmium-Resistance in Microorganisms
In response to metal toxicity and prevalence in the environment, microorganisms
have evolved ingenious mechanisms of metal resistance and detoxification. Figure 2-1
summarizes the various mechanisms of metal resistance in bacteria. Some resistance
mechanisms are plasmid-encoded and tend to be ver>' specific for a particvilar metal.
Others are general, conferring resistance to a variety of metals.
In this research, while several cadmium-resistant isolates were found in both
metal-contaminated and uncontaminated soils, the degree of resistance in those
communities from the contaminated soils is addressed in Appendix A. It is generally
thought that the more resistant organisms, that is, organisms resistant to the highest
concentrations of cadmium, will be found in soils with prior metal exposure.
discussed in the Summary section of this chapter, this was not necessarily the case.
Several mechanisms for cadmium resistance are known to exist.
mechanisms include active cadmium efflux systems (Fig. 2-2), intracellular sequestration
with metalloproteins, cell surface binding and extracellular binding in
' EPS sequestration
ATP efnux pumps
Pb^'Pb" Pb'
protein production
-(cys-cys), - Cd
Reduction ^ 1
Outer membrane
or cell wall binding
Cun_^ Cu"Precipitation as
metal salts
Cd" —• CdS
Cd'"—^ CdPO
Figure 2-1. In response to metal toxicity, many microorganisms have developed unique
mechanisms to resist and detoxify harmful metals. These mechanisms of resistance may
be intracellular or extracellular and may be specific to a particular metal or a general
mechanism able to interact with a variety of metals.
Figure 2-2.
A proposed model for cadmium influx and efflux in a bacterial cell.
Cadmium enters via a manganese transport pathway that relies on membrane potential.
The cadmium efflux system excretes cadmium via a Cd^V2Br antiport. Adapted from
Silver (1983).
exopolymeric layers. Generally thought to be plasmid-encoded (with the exception of
cell surface and exopolymeric binding), the frequency of occurrence of these mechanisms
is unclear. The effectiveness of each mechanism of resistance is also unclear. The
diversity in and effectiveness of cadmium-resistance mechanisms in environmental
isolates is addressed in Appendix B.
Degradation of 2.4-D
The genes for 2,4-D degradation are both chromosomal and plasmid-encoded
(Fig. 2-3). The endproduct of 2,4-D degradation is succinic acid, which can be degraded
to CO2 or can enter the tricarboxylic acid (TCA) cycle. One of the most commonly
studied 2,4-D degrading microorganisms is Alcaligenes eutrophus JMP134. A. eutrophus
JMP134 has an 80 Kb plasmid, pJP4. The degradation of 2,4-D is partially encoded on
pJP4 (Fig. 2.1). While there have been several studies where pJP4 has been genetically
transferred to other microorganisms, conferring partial 2,4-D degradation; the prevalence
of pJP4-like plasmids in the environment is not clear.
In general, indigenous microbial populations can adapt to the presence of 2,4-D to
eventually degrade it and some populations have the ability to use 2,4-D as a sole carbon
While 2,4-D is generally not considered toxic to microorganisms, in high
concentrations the intermediate 2,4-dichlorophenol (2,4-DCP) can be toxic. The effect of
metals on the ability of microorganisms to degrade organic pollutants was examined in
2.4-D-ciioxYgenase itfdA)
hydroxylase {tfdB)
2,6 - dichiorophc not
3,5 -dichlorocatecnol
Chlorocatechol 1,2dioxygenase {tfdO
cycloisomerase (tfdD]
Chiorodfene lactone
isomerase (f/dF)
irons-2-chlorodiene kaclone
CIS-2-chlorodiene lactone
Chlorodiene lactone
hydrolase [tfdE)
reductase (chromosomal)
Succinic acid
Figure 2-3. Degradatioa pathway for 2,4-dichlorophenoxyacetic acid eacoded by pJP4
and chromosomal genes.
this research. It is commonly thought that the stress imposed by metals completely
inhibits the ability of organic-degrading microbial populations to perform. This research
used several cadmium-detoxifying microbial populations to detoxify cadmium to
facilitate 2,4-D degradation by the cadmium-sensitive A. eutrophiis JMP134 in cocontaminated systems as presented in Appendix C.
In this research, a new approach to the bioremediation of co-contaminated
systems is introduced together with new information on how soil microbial populations
respond to cadmium stress and mechanisms of microbial cadmium resistance.
Objective 1, the response of soil microbial communities from both metal-contaminated
and imcontaminated soils were compared in order to get a better understanding of how
microbial communities respond and adapt to metal stress. In this study, while cadmiumresistant microorganisms were isolated from
both contaminated and uncontaminated
soUs, the communities with prior metal exposure were resistant to higher levels of
cadmium. An unustial pattern of resistance was observed with one community from a
metal-contaminated soil. The number of resistant organisms increased with increasing
concentrations of cadmium. Typically, the number of surviving organisms decreases
with increasing metal exposure. The increased resistance pattern was not observed in
soils with no established metal exposure. One cadmitmi-resistant isolate was found to
have a similar response to increasing cadmium toxicity. It was hypothesized that the
activation of a cadmium-specific mechanism of resistance was responsible for the
increased resistance.
Of several cadmium-resistant isolates found, six were examined for the potential
for cadmium detoxification. Identified using ribosomal 16S DNA sequencing, the six
isolates represented the common soil genera Arthrobacter, Bacillus and Pseudomonas.
While all isolates were cadmiimi-resistant to some degree, the range of cadmium
resistance was broad.
The highest level of cadmium-resistance was 275 \xg ml*'
cadmium. Cadmium and antibiotic resistances are thought to co-occur as a result of
similar evolutionary origins. However, only one of the cadmium-resistant isolates was
highly resistant to cadmium and a variety of antibiotics.
In Objective 2, the polymerase chain reaction (PGR), DNA sequencing,
transmission electron microscopy and X-ray dispersive spectroscopy, atomic absorption,
in conjunction with biochemical techniques, allowed for the determination of specific
mechanisms of cadmium-resistance in these isolates. The degree of cadmium resistance
and the mechanism of cadmium resistance appeared to be linked in that the two isolates
with intracellular cadmium accumulation were also the most cadmium resistant. Two
other highly resistant
isolates accumulated cadmium in microbially-produced
extracellular polymers. Finally, the two least resistant of the six isolates examined were
one with a cadmium efflux system and another that produced a biosurfactant.
Of the six cadmium-resistant isolates, four had cadmium detoxifying mechanisms
of cadmium resistance. These same isolates were found to detoxify cadmium in pure
culture, laboratory soil microcosms and in a pilot field study where both high levels of
cadmium and 2,4-D were present. In Objective 3, these foiar isolates were found to
detoxify cadmium to sufficiently low levels such that complete degradation of 2,4-D by
the cadmium-sensitive 2,4-D degrader A. eutrophiis JMP134 occurred.
degradation of 2,4-D was observed in pure culture and laboratory soil microcosms. In the
field study, there was an up to 80% reduction in the 2,4-D level.
The research comprised by these three studies demonstrated that while cadmiumresistance is widespread, only those organisms that detoxify cadmium as part of their
resistance have the potential to facilitate the remediation of metal-contaminated systems.
In the effort to remediation both metal- and organic-contaminated sites, the cooperation
between metal detoxifying and organic-degrading microbial populations should be
T.M. Roane and I.L. Pepper
^To be submitted to Microbial Ecology (1999)
Department of Soil, Water and Enviroimiental Science
429 ShantzBldg. #38
The University of Arizona
Tucson, AZ 85721
Running title: Cadmium-resistant microbial communities
The development of modem technology has generated a rapid increase in the
production and consumption of metals and metalloids. Microbially-based remediation
technologies can potentially reduce the risks associated with such environmental metal
In this smdy, we analyzed the soil microbial communities from
uncontaminated and two metal-impacted soils, and found that while cadmium adversely
affected culturable nimibers in all the soils, cadmium-resistant isolates were found from
each of the soils. With exposure to 24 and 48 jxg ml"' soluble cadmium, the metalcontaminated soil communities were more resistant than the uncontaminated soil
community. In addition, in one metal-stressed soil, the resistant population became more
resistant with increased cadmium levels. Ribosomal 16s DNA sequencing identified the
isolates as Arthrobacter, Bacillus or Pseudomonas spp.
Further characterization
demonstrated that two of the isolates were highly resistant to soluble cadmium with
maximum resistance at 275 [ig ml"' cadmiimi. These isolates were also resistant to a variety
of antibiotics namely ampicilh'n, gentamicin, penicillin and streptomycin, but no overall
correlation was found between enhanced antibiotic resistance and cadmium-resistance. One
Pseudomonas isolate HI did become more resistant with increasing cadmium levels,
suggesting a different resistance mechanism at high cadmium concentrations.
Despite decades of effort, metal-contaminated soils represent one of the most
difficult challenges facing bioremediation. In soils, metals readily sorb to soil particles or
organic matter and can precipitate as inorganic salts. In contrast, soluble metal ions are
generally available to interact with biological systems often with dire consequences. While
more attention has been focused recently on the remediation of metals, there is no routine
treatment for toxic metals dispersed in soils and sediments, and those commonly used rely
on physical or chemical approaches (McLead and Beveridge, 1990; Turn and Tels, 1991).
However, the high costs and, in some cases, the physical impracticality associated with
these treatments make alternative remediative options a necessity. One such alternative
strategy for the remediation of metal-contaminated soils is microbially based metal
remediation (Summers, 1992).
Several studies have found that metals influence microorganisms by adversely
affecting their growth, morphology, and biochemical activities, resulting in decreased
biomass and diversity (Baath, 1989;Dean-Ross and Mills, 1989; Heggo and Angle, 1990;
letswaart et al., 1992; Reber, 1992; Roane et al., 1996). Despite these toxic stresses,
numerous microorganisms have evolved metal resistance/detoxification mechanisms,
including volatilization, extracellular precipitation and exclusion, intracellular sequestration
and membrane-associated metal pumps (Hughes and Poole, 1989; Silver 1998). Microbialbased metal remediation relies heavily on the abihty of some microorganisms to resist and
detoxify metals.
To increase the success of microbially-based metal remediation technologies, a
better xmderstanding of microbial population responses to metal stress is necessary. Metal
mobility and variable metal speciation that occur due to environmental factors exacerbate
the response of microbial populations in metal-contaminated systems. Changes in metal
bioavailability can greatly affect toxicity and stress microbial populations. Using aged
metal-contaminated and uncontaminated soils in this study, we examined how some soU
microbial populations respond to metal stress. The present work evaluated the difference in
metal resistance response in microbial populations from different soils exposed to cadmium
(the second most common metal found at Superfund sites (Enger and Smith, 1992). Since
cadmium resistance is known to occur in many bacterial genera, additional objectives were
to isolate and characterize cadmium-resistant isolates with the potential for use in
microbially based metal remediation.
Field Soil Characterization
Field Site
In the late 1940s, more than 3.2 million pounds of aluminvim dross were deposited
in the area now known as the Olive Grove neighborhood, located just east of Tucson, AZ,
bordering the Davis-Monthan Air Force Base. Discarded before the Resource Conservation
and Recovery Act RCRA) was established in 1976, aluminum-dross, a metallic ash by­
product of the meltdown of scrap aluminum, contains potentially toxic levels of cadmium
and lead. Containment efforts are currently underway to reduce human exposure to these
metals. For this study, two dross-impacted soils (OGl and 0G2) were collected from the
Olive Grove neighborhood. An uncontaminated control soil, Brazito, was collected from
the University of Arizona Campbell Avenue Agricultural Station, Tucson, for comparison
Chemical/Physical Characterization
Soil samples were collected from the top 10 cm of the soil surface horizon. The
University of Arizona Soil, Water and Plant Analysis Laboratory performed the following
soil analyses: pH (Page et al., 1982), percent organic carbon (Artiola, 1990), and total
cadmium and lead (U.S. EPA, 1986). Physiochemical parameters known to influence soil
microorganisms were measured (Table 1). Both metal-contaminated soils (OGl and 0G2)
and the non-metal-contaminated soil (Brazito) were similar in terms of soil pH
(approximately 8), clay content (8.6 - 15.7%), soil texture (all were sandy loam) and ranged
from 0.2-0.5% organic carbon. Soil OG2 had substantially higher cadmivim and lead levels
than soil OGl (55 and 5 (j,g g"'; and 1660 and 75 jig g"', respectively) and concentrations in
the uncontaminated Brazito soil were below detection (<0.1 |ig ml"' Pb; <0.01 jig ml"' Cd).
The similarities in soil parameters, other than metal levels, provided the basis for the
microbiological comparisons of these soils.
Microbial Characterization
Total microbial numbers in each study soil were based on direct acridine orange
staining of diluted soil (between 10-100 cells per field) slurries (one g-soil dry weight with
9.5 ml sterile sodium pyrophosphate, Na4P207(H20),o. Thirty fields per slide were examined
using fluorescence microscopy (Hobbie et al., 1977).
Culturable soil bacterial numbers were based on two replicate experiments using
conventional plating techniques, from the soil slurries described above, onto the minimalnutrient medium for heterotrophic organisms, R2A (Difco, Baltimore, MD). Nutrient agar
(Difco, Baltimore, MD) was also used to enimierate culturable bacteria. However, since
similar numbers were obtained with either media, R2A was used to determine heterotrophic
numbers. Plates were incubated for one week at 25°C prior to covmting.
Community Cadmium Resistance
The degree of cadmium toxicity on a community level was assessed by exposing
three dififerent soil microbial communities, two from metal-contaminated OGl and 0G2
soils and one from the uncontaminated soil Brazito, to various cadmiimi concentrations in a
defined mineral medium (MSM) per liter:
0.5 g sodium citrate, CgHjNajOv; 0.1 g
magnesium sulfate, MgS04; 1.0 g ammonium sulfate, (NH4)2S04; 1.0 g glucose, CeHijOg, and
0.1 g sodiijm pyrophosphate, Na4P207(H20),o, buffered to pH 6.0 with potassium phthalate,
KHC8H4O4. Cadmium was supplied as CdClj with bioavailable concentrations being 95%
of the total cadmium added. One-ml of a 1:10 soil slurry (1 g soil-dry weight with 9.5 ml
sterile sodium pyrophosphate, vortexed 2 min.) was used to inoculate 25 ml of the defined
mineral medium amended with cadmixmi.
Bacterial growth was enumerated every 24 hr for 9 days when stationary phase
growth was reached. To determine the actual numbers of cadmium-resistant bacteria at
defined time intervals, the above cultures were plated onto MSM amended vidth the same
level of cadmium as the corresponding liquid culture flask. Microbial-induced cadmium
sequestration was of interest as a possible mechanism of cadmium detoxification. In
replicate experiments, any reduction in cadmium solubility as a result of microbial growth
was determined using a flame atomic absorption spectrophotometer (AA) following a 0.2
|im filtration to remove immobilized cadmium and cell debris. Controls consisted of MSM
amended with cadmium in the absence of inoculation.
Characterization of Cadmium-Resistant Isolates
Enterobacterial repetitive intergenic consensus polymerase chain reaction (ERIC
PCR) vras used to distinguish the isolates firom each other genetically (Versalovic et al.,
1994). ERIC PCR uses primers representing conserved repetitive DNA sequences found in
all bacteria. Amplification products of differently sized DNA segments between the repeat
sequences result ia genetic fingerprints
that allowed differentiation of isolates.
products were separated by electrophoresis in a 2% agarose gel at 100 V cm"' and stained
with ethidium bromide (1 ja,g ml"').
Isolate Identification
In this study, 16S rDNA sequences were used to identify, to the genus level, six
cadmium-resistant isolates from both metal-contaminated and uncontaminated soils (Dowd
et al.. 1999). The six isolates were chosen based on differing colony morphologies.
Cadmium Maximum Resistance Level
The maximum resistance level (MRL) was defined as the highest concentration of
cadmium at which at least lO"* cells ml"' consistently remained viable after 48 hr from an
initial 10® cells ml"' inoculation. The MRL of cadmium reflected the degree of cadmiumresistance. Reduction in soluble cadmium in response to growth of the cadmium-resistant
isolates was determined by filtering liquid cultures through a 0.2 |jin polycarbonate filter
and measuring the remaining cadmium in solution with a flame AA.
Antibiotic sensitivity profiles
Since antibiotic resistance and metal resistance are thought to be linked because of
their presence on plasmids, antibiotic profiles were made for each cadmium-resistant
isolate. Isolates were grown in mineral salts broth, pH 6.0, at 28°C for 24hr. A 0.1 ml
firaction from each culture was plated onto R2A plates. The following antibiotics (Becton
Dickinson, CockeysvUle, Md.) were tested by aseptically placing the appropriate antibiotic
disks (no more than six per 100-mm petri plate) onto the inoculated plate: ampicillin (10
fxg), bacitracin (10 U), carbenicillin (100 |ag), erythromycin (15 |ag), gentamicin (10 |ag),
novobiocin (5 jig), penicillin G (10 U), rifampin (5 |j.g), streptomycin (10 |j.g) and
tetracycline (30 fxg). Plates were incubated at 28°C for 24-48 hr. Zones of inhibition
(clearing) were measured and sensitivity or resistance was determined using a zone
diameter interpretive chart supplied by the manufacturer.
Culturable Microorganisms in Metal-Contaminated and Uncontaminated Soils
Total bacterial numbers in the three soils ranged from 7.2x10^ to 2.7x10'° cells g"'
soil (Table 1). As expected, total numbers were higher in all soils when compared to
culturable numbers; the uncontaminated soil exhibited the greatest culturable recovery
(3.2x10^ CFU g"' soU), and both metal-contaminated soUs showed lower cultural counts
(9.9x10' and 5.7x10^ CFU g"' soU, respectively). The primary difference among the soils
examined in this study was the metal stress reflected in both the total and culturable
numbers of microorganisms. Smdies have shown that chronic metal stress results in
decreased bacterial diversity, biomass and activity (Baath, 1989).
The present study
confirmed this. Perhaps more striking was that the culturable numbers were < 0.1% of the
total number of organisms in the contaminated soils, compared to 44% in the
uncontaminated soil. This indicated a high number of viable but nonculturable organisms
in the metal-contaminated soil. The high number of viable but nonculturable organisms
was likely due to the metal stress imposed in these soils.
Within 24 hr, 24 and 48 ^g ml"' cadmium was more toxic to the community from
the uncontaminated soil than the metal-contaminated soils. The nimiber of culturable cells
recovered in all soUs after 24 hr incubation in broth decreased with increasing cadmium
concentration in the test medixim (Fig. 1). However at the higher cadmium concentrations.
both metal-contaminated soils, OGl and 0G2, exhibited one to two orders of magnitude
more growth than the uncontaminated Brazito soil.
The same trend was repeatedly
observed in three subsequent trials. Each of the communities exhibited cadmium toxicity
upon initial exposure to 12 |ig ml"'cadmium; however, for all three soils, the differences
among the soils incubated at 12 j^g ml"' cadmium were similar in magnitude to the
differences observed with 0 fig ml"' cadmium. Exposure to concentrations greater than 12
|j.g g"' resulted in a 2-order of magnitude decrease in the uncontaminated Brazito soil while
metal-contaminated soils OGl and 0G2 exhibited less than a log decrease, indicating
greater cadmium-resistance inherent to these communities.
Following 72 hr incubation, it was noticed that the OGl population was two to three
orders of magnitude greater at 48 jig g"' than 24 [ig g"' bioavailable cadmium. This
phenomenon was repeatedly observed in subsequent trials (Fig. 2). Again, as expected,
numbers declined upon low level cadmium exposure; however, the numbers recovered
upon exposure to higher concentrations, suggesting a reduction in cadmium toxicity at
higher cadmium levels. We hypothesize that the reduction in cadmixim toxicity was due to
development of a resistant population. There were noted changes in the remaining soluble
cadmium by the end of the 9 day incubation period: the initial solution that contained 12 p,g
ml'^ cadmium had 4.2 [ig ml"' soluble cadmium; 24 jj.g ml"' cadmium had 17.8 p.g ml"'
soluble cadmium; and 48 p,g ml"' cadmium had 45.2 |j.g ml"' soluble cadmiimi (Table 2).
There were no detected changes in soluble cadmium concentrations in iminoculated flasks
amended with cadmium incubated for the same period of time. This phenomenon was not
evident in the vmcontaminated Brazito soil, and was seen to a lesser extent in the other
metal-contaminated soil 0G2 (data not shown).
Types of Cadmium-Resistant Isolates from Contaminated and Uncontaminated Soils
Based on distinct colony morphologies from twenty-five resistant isolates, six were
fiirther characterized. Each isolate was shown to be genetically distinct from each other
using ERIC PGR (Fig. 3), Using I6S rDNA sequencing, the isolates were identified as
Arthrobacter, Bacillus and Pseudomonas spp. (Table 3).
resistance to cadmium.
Each isolate differed in its
Although it is believed that antibiotic resistance and metal
resistance are linked genetically and physiologically (Choudhury and Kumar, 1998; Xu et
al., 1998), such a pattern between antibiotic resistance and cadmium resistance was only
found in one isolate. Pseudomonas HI, which was highly resistant to cadmium, was also
resistant to each of the antibiotics tested (Table 4).
Two of the most resistant isolates were Bacillus H9 and Pseudomonas HI resistant
up to 275 and 225 y.g ml"' soluble cadmium, respectively. Two other bacilli were less
resistant, only up to 5 (xg ml"' cadmium. Pseudomonas, a metaboUcaUy diverse organism,
was expected to be highly resistant and was in fact found to be resistant to 225 |j.g ml"'
soluble cadmium. These six isolates only begin to demonstrate the diversity in organisms
and degrees of cadmium resistance to be found in response to metal exposure.
Effect of Cadmium on Resistant Isolates from Contaminated and Uncontaminated
This study confirms that cadmium adversely affects soil microbial commimities.
resulting in decreased culturable counts, an effect that has been reported by others
(Frostegard et al., 1993; Hattori, 1992). However, in broth amended with cadmium, the
highest numbers of resistant organisms were found in the metal-contaminated soils
indicating a strong response of the soil community to cadmium addition. While cadmiumresistant organisms from
the metal-contaminated soils cultured initially, many of the
isolates did not grow upon transfer either with or without cadmium present indicative of
specific growth requirements not met This may, in part, explain the high number of viable
but nonculturable organisms in these soils. As a result most of the metal-resistant isolates
examined in this study were from the uncontaminated Brazito soil. The nonculturability of
the isolates from the metal-contaminated soils on R2A, mineral salts agar and nutrient agar
may be a reflection of the high number of viable but nonculturable organisms used to
explain the difference between the viable and total direct counts in Table 1. Because of this
observation, interpretation of the degree of resistance in an environmental system may need
to be done on a commimity basis to avoid the nonculturability of individual isolates within
the system. Further research is being conducted to find the conditions to culture these
Based on the interesting pattem of increased resistance with increased cadmivim
levels observed with the microbial commimity from soil OGl (Fig. 2), the cadmium
resistant isolates were screened for individual isolates showing a similar pattem of
resistance. One such isolate was Pseiidomonas HI from the imcontaminated Brazito soil.
"^Iiile one would expect the number of surviving organisms to continue to decrease as
cadmium levels became increasingly toxic, one resistant isolate HI exhibited 6X more
growth at 48, 60 and 150 |ag ml"' cadmium than at 24 (ig ml'^ cadmium within 72 hr.
Figures 4a and 4b represent replicate trials. Dramatic changes in the soluble cadmium
concentrations were evident within a 72 hr incubation period. For example, at 48, 60 and
150 |ag ml"' cadmium, there was up to a 28 (xg ml"' decrease in soluble cadmium (Table 5).
This observation
raised the question
some metal-resistant
microorganisms can use multiple mechanisms of resistance to the same metal. There are
several known pathways for metal resistance in microorganisms. Can a bacterium use one
resistance mechanism at lower metal concentrations and then shift to another under more
stressfiil conditions? Yoshida et al. (1998) found different resistance reactions when
Thiobacillus intermedius 13-1, Escherichia coli JM109 and Agrobacterium radiobacter
IFOI2665bl were observed under low-level, long-term and high-level, short-term exposure
to lead, molybdenum, nickel and zinc. Two other studies by Roane and Kellogg (1996) in a
study of lead-resistance in soil communities from lead-contaminated soils, and by Sandrin
et al. (1999) in a study of the cadmitim-resistance of a naphthalene degrading bacterium
have also observed increasing resistance with increasing metal concentrations.
This study found that the microbial populations from the imcontaminated soil
responded equally well as the populations from the metal-contaminated soils to low levels
of bioavailable cadmium stress with respect to overall numbers of resistant organisms.
Differences in cadmium-resistance among the soils were only observed at higher levels of
cadmium stress. It was observed that some cadmium-resistant populations might be able to
increase resistance imder higher levels of cadmium stress, suggesting a possible change in
resistance mechanism.
Although several metal-resistant microorganisms have been
identified and are
being studied, the action of these microorganisms in soil
communities is not well understood. The potential for the use of metal-resistant populations
in remediation of contaminated sites is limited until we more fiilly understand the
correlation between microbial metal-resistance and the environment.
This work was supported in part by grant number 5 P42 ES04940-07 firom NEEHS,
Superfund Program. We thank the Arizona Department of Environmental Quality for
access to the Olive Grove contaminated soils.
Table 1. Characterization of two metal-contaminated soils, OGl and 0G2, and the uncontaminated control soil Brazito.
Organic Carbon
Total Cd
Total Pb
Culturable Cells
Total Cells
(Hg g"')
(Kg g')
(CFU g-')
(Cells g')
sandy loam
sandy loam
Brazito 8.2
sandy loam
"Not detected (<0.1 |ig ml"' Pb; <0.01 |.ig ml"' Cd). ± represents standard deviation.
Table 2. Changes in Soluble Cadmium Concentrations (jxg ml"') Upon Growth
of the Soil Community from Soil OGl.
Cadmium in Solution
(|ig ml"')
Expected Cadmium
Actual Cadmium®
ToUowing a 9 day incubation at 28°C.
Tables. Isolate identification and cadmiimi-resistance.
I6S Id and %S''
MRL'' of Cd (^g ml"')
Bacillus (99%)
Arthrobacter (98%)
Pseudomonas (100%)
Bacillus (88%)
Pseudomonas (99%)
Pseudomonas (99%)
= % siroilarity with GeneBank sequences.
''Maximum resistance level (MRL) was based on soluble cadmium concentrations.
Table 4. Antibiotic profiles for each of the six cadmium-resistant isolates.
Ampicillin (10 |Lig)
Bacitracin (10 U)
Carbenicillin (100 fig)
Erythromycin (15 |.ig)
Gentamicin (10 |.ig)
Novobiocin (5 (.ig)
Penicillin (10 U)
Rifampin (5 fig)
Streptomycin (10 |.ig)
Tetracycline (30 |ig)
"R = resistant.
•"S = susceptible.
= intermediate.
Table 5. Changes in Soluble Cadmium Concentrations (|ig ml ') Upon Growth of CadmiumResistant Isolate HI.
Cadmium in Solution
(^g ml')
Expected Cadmium
Actual Cadmium"
"Following a 72 hr incubation at 28°C.
Figure 1. Total microbial community response to cadmium concentration from different
soils measured as culturable cell concentrations following a 24 hr exposure. Both OGl
and 0G2 soils are metal-contaminated, while Brazito represents an uncontaminated
Figure 2. Four replicate trials looking at the cadmium-resistant populations in metalcontaminated soil OGl as a fiinction of increasing soluble cadmium. Data at 72 hr
Figure 3. Agarose gel of ERIC PGR fingerprints for each cadmium-resistant isolate.
Fingerprints for all isolates were unique. Lanes: 1, 123-bp DNA ladder as a size standard;
2, Isolate LI; 3, Isolate I la; 4, Isolate HI; 5, Isolate D9; 6, Isolate H9; 7, negative control
(no DNA).
Figure 4. (a) Cadmium-resistant HI showing a similar pattem of resistance with increased
resistance to 60, 150 and 300 |ig ml"' cadmium and less resistance at 48 p.g ml"' cadmium.
•Initial inoculant concentration was 1.0 x 10® CFU ml"', (b) Cadmium-resistant isolate HI
showed higher resistance to 48, 60 and 150 jig ml"' cadmium with at 76, 92, and 114 hr
exposures than to 24 |ig ml"' with the same exposure times. *Initial inoculant concentration
was 1.0 X 10^ CFU ml"'.
S ^l.Oe+3
I .Oe+1
Soluble Cadmium Concentration
(|Lig ml ')
•1 Trial 1
mm Trial 2
•1 Trial 3
•• Trial 4
Z a
| b l.Oe+4
Soluble Cadmium
(lig ml"')
24 Hrs
48 Hrs
76 Hrs
100 Hrs
76 Hrs
92 Hrs
114 Hrs
l.Oe+9 l.Oe+8 l.Oe+7
60 150 300
Soluble Cadmium
Soluble Cadmium
(Mg ml"')
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T.M. Roane and I.L. Pepper
'To be submitted to Environmental Microbiology (1999)
Department of Soil, Water and Environmental Science
429 Shantz Bldg. #38
The University of Arizona
Tucson, AZ 85721
Running title: Mechanisms of bacterial cadmium-resistance
While there is an increasing understanding of microbial metal resistance, few
studies have addressed the diversity of metal resistance mechanisms in environmental
isolates. In this study, the overall mechanisms of cadmium-resistance for six cadmiumresistant soil isolates were explored. Three diEFerent mechanisms of cadmium resistance
were observed. The isolate Pseudomonas LI was resistant up to 5 |j.g ml*' cadmium and
produced a biosurfactant. Pseudomonas Ila and Arthrobacter D9, resistant to 20 and 50
|j.g ml"^ cadmiimi, respectively, produced exopolysaccharides, which xising transmission
electron microscopy (TEM) were shown to accumulate cadmium.
Finally, another
Pseudomonas isolate HI and Bacillus H9 were most resistant at 225 and 275 p.g ml'^
cadmium, respectively.
When viewed with TEM, both HI and H9 exhibited an
intracellular mechanism of cadmiimi sequestration. X-ray analysis confirmed cadmium
An Arthrobacter isolate E92, with an unidentified mechanism, was
resistant up to 5 |i.g ml"' soluble cadmium. Only three of the six isolates had plasmids,
Arthrobacter E92 (8.1 Kb), Pseudomonas HI (18.5 Kb) and Bacillus H9 (10.5 Kb).
None of the isolates were positive for the cadA. nor the cadC genes involved in ATPdependent cadmium efflux. The mechanisms most efficient at cadmium detoxification
were exopol5Tner production and intracellular accumulation which also accounted for the
most cadmium-resistant isolates.
Cadmium is one of the most environmentally significant polluting metals.
Concentrations of cadmium have been found as high as 50,000 mg Kg"' in sediments and
as high as 1000 mg Kg'' in water columns (Forstner 1986). Generally considered more
mobile than other metals because of less affinity for soil components, cadmium
contamination posses a serious long term threat to the environment and to human health
and necessitates a better understanding of how microorganisms influence cadmium
toxicity in the environment.
As a result of a high affinity for sulfhydryl groups, cadmium is highly toxic to
microorganisms, inhibiting respiration, growth and metabolism.
Metals in general,
including cadmium, are thought to inhibit the ability of microorganisms to degrade
organics because of the inhibition of microbial metabolism. Inhibition of allochthonous
and anthropogenic organic degradation only exacerbates the need for more information
on cadmium-resistance and detoxification mechanisms.
Many microorganisms have been foimd to be resistant to cadmium. Cadmiumresistance includes cadmium-specific and nonspecific mechanisms.
The cadmium-
nonspecific mechanisms involve extracellular or intracellular sequestration by binding to
cellular polymers. For example, cadmium binding to the cell wall has been observed
with Proteus mirabilis (Andreoni et al. 1991) and to extracellular slime layers with
Arthrobacter spp. (Kurek et al. 1991). Cadmium can also be precipitated with
microbially-produced sulfides seen with Klebsiella aerogenes (Aiking et al. 1982) or
phosphates as with Citrobacter spp. (Macaskie and Dean 1984).
The intracellular
binding of cadmium with metallothionein-like proteins has been reported for
Pseudomonas putida and Synechococcus sp. (Higham et al. 1984; Gupta et al. 1992).
Cadmium-dependent energy-dependent efflux systems have been intensively studied in
Staphylococcus aureus and Alcaligenes eutrophus. In S. aureus, the cadAC genetic
determinant mediates an ATP-dependent cadmiimi and zinc efflux mechanism (Nucifora
et al. 1989). Another gene in S. aureus, cadB is thought to be an inducible cadmiumbinding protein to sequester cadmium (El Solh and Ehrlich 1982). The czc system of A.
eutrophus encodes for an active cation efflux mechanism driven by cation-proton antiport
(Nies and Silver 1989).
A study by Trajanovska et al. (1997) of Gram-positive and Gram-negative
bacteria from a lead-contaminated soil found cadmium-resistance in isolates but no
apparent homology with the czc genes. Jones et al. (1997) foimd the regulatory genes
involved in eukaryotic cell invasion in Burkholderia pseudomallei were homologous to
czcR gene in A. eutrophus.
Clostridium thermoaceticum when exposed to 1 mM
cadmium precipitates cadmium intracellularly with a metalloprotein (Cunningham and
Lundie 1993), while Listeria monocytogenes uses the cadAC system to resist up to 512
|xM cadmium (Lebrun et al. 1994). Few studies, however, have attempted to study the
diversity in cadmium-resistance mechanisms in enviroimiental isolates.
The prevalence and distribution of these cadmium-resistance mechanisms in the
environment is not clear. The purpose of this study was to evaluate the diversity in
mechanisms in cadmium-resistant soil bacterial isolates to increase our understanding of
microbial responses to cadmium toxicity. To determine whether cadmium-resistance
mechanisms are similar or vary in a soil, an uncontaminated Brazito soil was screened for
cadmium-resistant microbial populations. Five cadmium-resistant populations from the
uncontaminated Brazito soil and one resistant population from metal-contaminated soil
OGl were characterized in terms of cadmixmi-resistance, plasmid profiles, presence of
the cadK and cadC, genes, production of biosurfactant and exopolymer, cellular
accumulation of cadmium, and the ability to reduce the amount of soluble cadmium in
Bacterial Strains and Culture Conditions
Cadmium-resistant bacteria were isolated from growth experiments performed in
a defined mineral medium (MSM) amended with soluble cadmiimi as CdCli in
concentrations from 0-48 ^ig ml'^ as previously described in Roane and Pepper (1999). In
the initial isolation, 1 ml of a 1:10 soil slurry (1 g soil-dry weight with 9.5 ml sterile
sodium pyrophosphate, vortexed for 2 min.) was used to inoculate 25 ml of the MSM
amended with cadmium.
The microbial communities from
both a cadmiimi-
contaminated and an uncontaminated soil were examined for cadmium-resistance. All
subsequent culturing took place in 25 ml MSM amended with cadmium incubated at
28°C on a rotary shaker at 180 rpm.
A total of 20 distinct (based on colony morphology) cadmium-resistant
populations, only 50% of which could be subcultured, were isolated from soil slurries
when plated directly onto MSM agar with 48 |j.g ml"' cadmium. Six isolates (based on
distinct colony morphology and ability to grow to high numbers in 48 hrs) were then
chosen for further study.
These isolates were identified based on their 16S rDNA
sequences using the method of Dowd et al. (1999). GenBank was searched for DNA
sequence similarities with the BLAST program. Repeated several times, the
resistance level (MRL) was defined as the highest concentration of cadmium at which at
least lO'^ cells ml'^ consistently remained culturable after 48 hr firom an initial 10® cells
ml"^ inoculation. The MRL of cadmium reflected the degree of cadmium-resistance.
Cadmium concentrations
were determined using a flame
spectrophotometer following a 0.2 jjm Nuclepore filtration to remove particulates.
Plasmid Profiles
Siace most metal resistance is thought to be plasmid encoded, the six isolates
were examined for the presence of plasmids ranging in size from 2.6 to 350 megadaltons
following incubation in the presence of 3 or 12 mg ml*^ cadmium depending on the
resistance level of the isolate. The alkaline lysis procedure of Kado and Liu (1981) was
used to isolate and purify the plasmid DNA for PGR analysis.
Detection of cadK and cadC,
Both chromosomal and plasmid DNA were examined for the presence of a cadAlike gene and/or the regulatory region of the Gad operon. One set of cadA. primers was
designed from the cad A sequence from S. aureus pI258 using the LASERGENE program
(DNA STAR Inc., Madison,
The cadA primers from S. aureus consisted of a 19-
5'CGGTAGATGATGAAGTA riT'3' with a 553 bp product. A second set of primers for
the detection of the cadC gene developed by Endo and Silver (1995) were used. To
confirm the absence of the cadmium efflux system, a third set of 17-mer primers from the
5'CCTTGTATTCGATTAAC3' (Endo and Silver, 1995) were used.
Touchdown PGR with step annealing temperatures ranging from 54 to 61°C was
used to amplify target sequences in lysed cell extracts and on plasmid DNA.
Amplification occmxed at 60°C for 25 cycles using AmpliTaq Gold Polymerase (PerkinElmer, Foster City, CA) and 2.5 mM MgCb final reaction concentration. All reactions
used the lOX PCR buffer without MgCb provided by Perkin-Elmer with the AmpliTaq
Gold Polymerase. PCR products were nm on a TBE 1.2% agarose gel at 100 V/cm,
stained with ethidium bromide (1 |J.g ml-1) and viewed imder ultraviolet lighL
Production of Extracellular Polymers
Many microorganisms have extracellular polymeric layers that confer metal
resistance. These polymeric layers are anionic in nature and so attract and sequester
cationic metals. The rapid screening method developed by Liu et al. (1998) was used to
of two
exopolysaccharides succinoglycan or galactoglucon (EPS II). The method relies on the
differential staining of polymer-producing versus nonpolymer-producing organisms.
Production of Microbial Siirfactants
Biosurfactants or microbially produced surfactants have been shown to complex
metals (Tan et al., 1994), thereby reducing metal bioavailability. For this study, each
cadmium-resistant isolate was screened for the production of biosurfactant using the
method of Bodour and MiUer-Maier (1998). In this modified drop-collapse technique,
the ability of a drop of cell suspension to remain as a drop is used to assess cell
hydrophobicity and surfactant production. Cells were grown in MSM, pH 6.0, and
incubated for 48 hr at 28°C.
Transmission Electron Microscopv
Transmission electron microscopy (TEM) was used to assess morphological
changes in response to cadmiimi exposure. Bacteria (1.5 ml) grown in MSM pH 6.0
containing 3 fig ml"^ cadmiimi for Arthrobacter sp. E92; 10 (ig ml'^ cadmium for
Pseudomonas sp. Ila; 125 p.g ml"^ cadmium for Pseudomonas sp. HI and Bacillus sp. H9
were pelleted at 14,000 g for 2 min. The cells were rinsed in sterile deionized water and
fixed in 3% glutaraldehyde in O.IM cacodylate buffer, pH 6.0, saturated with oxine
(Sigma Inc.) using microwave fixation
(Gibberson et al., 1997; Lindley 1992) to
minimize cadmium leaching associated with traditional fixation.
Oxine reacts
specifically with metals increasing contrast under the TEM (Yoshizuka et al., 1990).
Cells were postfixed in 2% osmium in O.IM cacodylate buffer, pH 6.0. Following
fixations, cells were dehydrated using a reagent grade ethanolrHiO gradient at each
concentration: 30, 50, 70, 95, 100, 100 and 100% (Hayat 1989). Cells were infiltrated
with 50:50 and then 100% Spurr resin (Ted Pella Inc., Redding, Calif.). Samples were
polymerized overnight at 70®C, thin-sectioned with an RMC MT-7000 Microtome
(Research Manufacturing Corp., Tucson, AZ) and viewed at 60 kV with a Philips 420
transmission electron microscope (Philips Electron Optics Inc., Mahwah, NJ). Cadmium
accumulation was confirmed using the Noran Series Voyager II X3 elemental dispersive
spectroscopy (Noran Instruments, Inc., Middleton, WI).
Cadmium Mass Balance
Any reduction in soluble cadmium levels in broth cultures indicative of cadmium
precipitation/accumulation was measured using atomic absorption. In replicate flasks,
each isolate was grown in 25 ml MSM broth for 48 hr at 28°C at varying cadmium levels
up to the MRL. Precipitated and cell-associated cadmium were collected following
centrifugation at 10,000 x g for 20 min. and acidified with IN HCl to solubilize the
Both supernatant and the acidified cell suspension were examined for
cadmium. In control flasks without inocida, greater than 95% of the total cadmium was
The purpose of this study was to determine the diversity in mechanisms of
cadmixim-resistance in isolates firom soil.
The observed mechanisms of resistance
included extracellular accumulation in the EPS layer for Arthrohacter D9 and
Pseudomonas I la; intracelliilar accumulation by Pseudomonas HI and Bacillus H9; and a
an unknown mechanism for Arthrobacter E92 (Table 1). A sixth isolate Pseudomonas
LI from a metal-contaminated soil was additionally examined and found to produce
It was interesting to note the diversity of mechanisms, especially
considering that each Pseudomonas isolate used a different mechanism of cadmiumresistance. As simmiarized in Table 1, the six isolates were resistant to a wide range of
cadmium, from 5 to 275 |ag ml"^ Only three of the six isolates had plasmids ranging
from 8.1 Kb to 18.5 Kb in size. All of the isolates were grown in the presence of
cadmium to increase the copy number. Arthrobacter E92 had an 8.1 Kb plasmid.
Isolates Pseudomonas HI and Bacillus H9 had 18.5 Kb and 10.5 Kb plasmids,
respectively. Both isolates HI and H9 exhibited intracellular accumulation in response to
cadmium exposure. Of the three isolates with no plasmids, two produced an EPS layer
and one produced surfactant.
ATP-dependent Efflux
A common mechanism in cadmium-resistant bacteria is ATP-dependent cadmixmi
efQux encoded by the Cad operon. The first cadmium-resistance mechanism examined
was the presence/absence of the cad A. gene. Found on the cad operon, the cad A gene
encodes for an ATP-dependent cadmiimi efflux. Using primers designed from the cadA
gene in S. aureus, no cacfA-like genes were detected. When the isolates were examined
for the cadC gene, which regulates the Cad operon, no cadC-]ik& genes were detected.
The absence of the cad genes was surprising since it is believed that the active cadmitim
efflux (partially coded for by the cadA gene) occurs in the majority of cadmium-resistant
bacteria. One possible explanation for the absence of the Cad operon in these isolates
was that these organisms were isolated from a soil with no known previous exposure to
Consequently, these organisms used other and perhaps less specific
mechanisms of cadmium-resistance.
Extracellular Binding
Metal binding to exopolymer is known to reduce metal bioavailability. While
generally associated with adhesion and protection against desiccation, exopoljoners act as
strong ionic attractants and, thus, readily bind metals. Two cadmium-resistant isolates
were found to produce exopolysaccharides (EPS), Arthrobacter D9 and Pseudomonas
I la. The metal-binding ability of EPS for isolate II a was observed with transmission
electron microscopy (Figure la,b) where sequestered cadmium was evident outside of the
cell associated with the EPS layer. EPS is generally not readily visible with TEM, as the
highly hydrated EPS layer collapses during dehydration in sample preparation. However,
the binding of cadmium to the EPS layer increased both the density and the rigidity of the
EPS layer, preserving it during preparation and subsequent viewing with the TEM. Both
isolates D9 and Ila were moderately resistant to cadmium (50 and 20 [ig ml"' cadmium,
There is increasing knowledge that microbially-produced surfactants can be used
in reducing metal toxicity. Biosurfactants, such as rhamnolipid, bind metals in an ionic
interaction. The result is that while the metal remains soluble (surfactants act somewhat
like chelators), the bound metal is less chemically-reactive and so less toxic.
Pseudomonas LI (from a metal-contaminated soil) was the only isolate foxmd to produce
a biosurfactant. Pseudomonas LI was resistant up to 5
ml"' cadmium.
Cellular Accumulation of Cadmium
Transmission electron microscopy yielded valuable information as to where
cadmium was being transformed in the cell. As previously discussed, Pseudomonas I la
produced an EPS layer which, when examined under the TEM, acctimulated cadmium
(Fig. 2a,b). Two other isolates, Pseudomonas HI and Bacillus H9, showed cadmium
accumulation intracellularly. In the presence of 125 lug ml"' cadmium, large diffuse
accumulations were apparent in Pseudomonas HI (Fig. 2b). These accumulations were
absent in the absence of cadmiiun exposure (Fig. 2a). Also evident were smaller granules
in the control cells (cells with no cadmium exposure) which were absent when the cells
were exposed to cadmium. There was also an increase in outer membrane density with
the treated cells, indicative of cadmium binding to lipopolysaccharides (Langley and
Beveridge, 1999). An X-ray analysis of the dark (electron dense) areas within the treated
cells and along the outer membrane showed some cadmium accumulation (Fig. 3). No
cadmium was detected in cells with no cadmium exposure (data not shown). The precise
mechanism of cadmium accumulation merits further investigation.
metallothionein production or polyphosphate precipitation present two possible
explanations wherein cadmium was sequestered intracellularly. We hypothesize that the
disappearance of the small granules in the control cells (not evident in the treated cells)
may be evidence of energy consumption in the form of the breakdown of polyphosphate
granules or the utilization of p-hydroxybutyrate.
When Bacillus H9 was exposed to 125 p.g ml"' cadmium (Fig. 2c,d), there was an
overall increase in cell density, especially in the cell wall, in conjunction with the
appearance of dark accumulations. Again, the disappearance of the small dark granules
in the control cells may indicate energy consiunption. It was interesting to note that the
two most cadmium-resistant isolates were Pseudomonas HI (up to 225 ^ig ml*') and
Bacillus H9 (up to 275 jag ml''), the two exhibiting intracellular cadmium accimiulation.
Effect on Cadmium Solubility
To ascertain which of the cadmium-resistance mechanisms found were most
efficient at cadmimn detoxification, a mass balance experiment was performed to
measure the reduction in soluble cadmium upon growth (Table 2). The amount of
cadmium present in solution decreased with isolates I la, D9, HI and H9 with growth
fi-om 10"* to 10^ CFU ml"'. The most dramatic decreases in soluble cadmiimi were seen
with Pseudomonas HI and Bacillus H9 (also the most resistant), such that there was an
average 36% loss in soluble cadmium with growth. Growth of Arthrobacter D9 and
Pseudomonas IIa resulted in 22% and 11% decreases in soluble cadmium. The isolates
that were least cadmivim resistant, E92 and LI, showed no detectable reduction in the
soluble concentration of cadmium when cell numbers increased from 10"* CFU ml"' to 10^
CFU ml"'. No reduction in soluble cadmium with Pseudomonas LI was expected since
biosurfactantrmetal complexes remain soluble. Based on the controls with no inocula,
>99% of the total cadmium should remain soluble.
The purpose of this study was to evaluate diversity in mechanism of cadmiumresistance in environmental isolates.
In one soil alone, at least three different
mechanisms of cadmium resistance were found. This finding suggested that diversity in
cadmium resistance may be greater than the literature suggests. This study foimd that
cadmium resistance in the environment is diverse both in terms of mechanism of
resistance and degree of resistance.
The mechanisms most efficient at cadmium
detoxification were exopolymer production and intracellular accxmiulation which also
accounted for the most cadmiimi-resistant isolates. More information is needed on the
occurrence of various metal resistance systems and how those systems can best be used to
detoxify metals in the environment.
We thank David Bentley of the University of Arizona Imaging Facility for his
assistance with the transmission electron microscopy, Adria Bodour for her assistance
with screening for biosurfactants, and Scot Dowd for his assistance with primer
development. This work was supported in part by grant number 5 P42 ES04940-07 firom
NIEHS, Superfimd Program and by grant nimiber DE-FG03-97-ER62470 firom the
Department of Energy, Joint Program on Bioremediation.
Table 1. Cadmium-resistant isolates and examined mechanisms of cadmium-resistance.
Arthrobacter sp. E92
Pseudomonas sp. LI
Pseudomonas sp. I la
Arthrobacter sp. D9
Pseudomonas sp. H1
Bacillus sp. H9
(base pairs)
"Cadmium maximum resistance level,
''not detected.
Table 2. The influence of microbial growth on the solubility of cadmium in MSM broth, pH 6.0.
Total Cadmium
Average % Soluble:
Total Cadmium
Average % Soluble:
Total Cadmium
Average % Soluble:
®not determined.
(% soluble)
Pseudomoms H1
Bacillus H9
(% soluble)
(% soluble)
(% soluble)
Arthrobacter D9
Pseudomonas 11a
(% soluble)
(% soluble)
(% soluble)
Arthrobacter E92
Pseudomonas LI
(% soluble)
(% soluble)
Figure 1. TEM micrograph of Pseudomonas II a in the absence of cadmium (a) and when
exposed to 20 p.g ml"' cadmium (b). Note the dark precipitate surrounding some of the
cells in (b), indicative of cadmium accumulation in the EPS layer produced by this
organism. Bar is 1|jm.
Figure 2. TEM micrographs of Pseudomonas HI in the absence of cadmiimi (a) and
when exposed to 125 |j.g ml"' cadmium (b) and Bacillus H9 in the absence of cadmium
(c) and when exposed to 125 ^g ml"' cadmium (d). Both isolates show increased cell
density and the appearance of diffuse dense accumulations, confirmed to be cadmium
with elemental X-ray analysis. The disappearance of the small dark granules present in
the control cells was thought to be consumption of either polyphosphate or PHB granules
in response to the cadmium stress. Bar is l|im.
Figure 3.
Elemental X-ray analysis of the difiuse dense accxmiulations present in
Pseudomonas HI in response to 125 jj.g ml"' cadmium exposure. The copper (Cu) peaks
were from the copper grid, the silicon (Si) peak was from the embedding medium Spurrs
and the osmium (Os) peak was from staining with OSO4. The cadmium (Cd) peaks
confirmed the presence of cadmium.
Figure 1.
Figure 2.
Figure 3.
Energy (keV5
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227: 138-.
T.M. Roane, K.L. Josephson and I.L. Pepper
^To be submitted to Applied and Environmental Microbiology (1999)
Department of Soil, Water and Environmental Science
429 Shantz Bldg. #38
The University of Arizona
Tucson, AZ 85721
Running title: Microbial remediation under co-contaminated conditions
While metals are thought to inhibit the ability of microbial communities to
degrade organic pollutants, several microbial-metal resistance mechanisms are known to
exist. This study examined the potential use of cadmium-resistant microorganisms to
enhance organic degradation imder co-contaminated conditions.
Several cadmiirai-
resistant soil microorganisms were tested to determine whether degradation of 2,4dichlorophenoxyacetic acid (2,4-D) in the presence of cadmium could be enhanced.
Resistant to up to 275 |ag ml"' cadmium, these isolates represent common culturable soil
genera including Arthrobacter, Bacillus and Pseudomonas. While none of the cadmiumresistant isolates could degrade 2,4-D, results of degradation studies conducted in pure
culture and laboratory soil microcosms showed that four cadmium-resistant isolates
supported the degradation of 500 |j.g ml'^ 2,4-D by the cadmium-sensitive 2,4-D degrader
Alcaligenes eutrophus JMP134. Degradation occurred in the presence of up to 24 |ig ml"^
cadmium in pure culture and up to 60 jj.g g"^ cadmium in amended soil microcosms. In a
pilot field study conducted in five-gallon soil bioreactors, the cadmivmi-resistant isolate
Pseudomonas HI enhanced 2,4-D degradation in the presence of 60 pig g"' cadmiimi.
Co-contaminated soils, soils contaminated with both metals and organics, are
considered difficult to remediate because of the mixed nature of the contaminants.
Traditional methods of treatment for such metal-impacted sites include incineration and
excavation (Gieger et al., 1993).
However, the increasing cost of excavation and
transport, and shrinking available landfill space, make altemative remediation options
attractive. One such altemative is bioaugmentation with metal detoxifying and/or organic
degrading microorganisms. Many microorganisms are known to degrade a variety of
organics, and likewise, a number of metal-resistant microorganisms are known to
detoxify metals, including selenium, mercury and cadmium (Stephen et al., 1999; Roane
et al., 1996). In co-contaminated sites, metal toxicity may inhibit the activity of organic
degrading microorganisms.
Consequently, bioremediation efforts focus on reducing
metal toxicity in sites with mixed contaminants. Until recently, bioaugmentation studies
focused on the introduction of a microorganism that is both metal resistant and capable of
organic degradation. Under field conditions, such an approach is often unsuccessfiil.
One reason may be that the energy requirements to maintain concurrent metal-resistance
and organic degradation are too high, such that the introduced organism cannot perform
both activities imder environmental conditions. The approach used in this study was to
co-inoculate with a metal-detoxifying population and an organic degrading population
that worked together to remediate both metal and organic pollutants in co-contaminated
While little is known about the degradative capabilities of microorganisms in the
presence of a metal stress, it has generally been thought that metal toxicity inhibits the
degradation of organics (Said and Lewis, 1991). While the use of microbial metalresistance in the remediation of metal-contaminated sites has been poorly documented,
studies have shown that 2,4-dichlorophenoxyacetic acid (2,4-D) degradation can occur in
the presence of nickel and zinc, showing it is possible to have organic biodegradation
under metal-contaminated conditions (Springael et al., 1993; Mergaey et al., 1985). The
issue of co-contamination is a serious one since approximately 37% of all contaminated
sites in the United States alone contain both metal and organic contaminants (Kovalich,
1991; Riley et al., 1992). Within the last decade, the U.S. Environmental Protection
Agency has made the remediation of co-contaminated sites a priority.
Metals, including cadmium, lead and mercury, are, in most cases, microcidal;
however some bacteria have developed the ability to be resistant to and detoxify these
metals. Metal detoxification strategies, including those for cadmium, may include metal
sequestration and precipitation (Silver and Phung, 1996). Unlike organics, metals cannot
be degraded and, thus, most biological metal remediation approaches rely on the
detoxification and immobilization of the metal to both reduce the biological toxicity and
to retard metal transport.
Exceptions to this include the microbial remediation of
methylated mercury and selenium via volatilization.
Stephen et al. (1999) used metal-resistant bacteria to protect indigenous soil Psubgroup proteobacterium ammonia oxidizers. The objective of the current study was to
determine the efficacy of a dual bioaugmentation strategy to facilitate organic
degradation within co-contaminated sites and then to demonstrate that the strategy works
in solution, soil and at the intermediate field scale. Six different cadmium-resistant
bacterial isolates that did not degrade 2,4-D were tested for the ability to allow 2,4-D
degradation to occur in the presence of toxic levels of cadmium, using cadmium-sensitive
Alcaligenes eutrophus JMP134 as the 2,4-D degrader.
Isolation of Cadmium-Resistant Bacteria
Cadmium-resistant bacteria were isolated from growth experiments performed in
a defined mineral salts medium (MSM) amended with cadmium-as CdCl2-concentrations
from 0-40 jj.g ml'V The MSM contained the following: 0.5 g sodium citrate, CgHsNasOy;
0.1 g magnesium sulfate, MgS04-7H20; 1.0 g ammonium sulfate, (NH4)2S04; 1.0 g
glucose, CsHiiOe, and 0.1 g sodium pyrophosphate, Na4P207(H20)io, buffered to pH 6.0
with potassium phthalate, KHC8H4O4. One-ml of a 1:10 soil slurry (1 g soil-dry weight
with 9.5 ml sterile sodium pyrophosphate, vortexed for 2 min.) was used to inoculate 25
ml of MSM amended with cadmium. Isolates were identified using the 16s ribosomal
DNA sequencing method of Dowd et al., 1999.
Cadmium Maximxmi Resistance Level
The maximum resistance level (MRL) was defined as the highest concentration of
cadmium at which at least lO'^ cells ml'^ remained culturable after 48 hr from an initial
10^ cells ml"^ inoculation. The MRL of cadmium reflected the degree of cadmiumresistance. Cadmium concentrations were determined using a flame atomic absorption
spectrophotometer following a 0.2 (im filtration of the sample.
Degradation Studies
Alcaligenes eutrophus JMP134 contains the 80 kb pJP4 plasmid that encodes for
the degradation of 2,4-D to 2-chloromaleylacetate. In contrast, the degradation of 2chloromaleylacetate to succinc acid is chromosomally encoded (Top et al., 1995; Perkins
et al., 1990; Short et al., 1991). A. eutrophus JMP134 was cadmixim-sensitive showing
no cadmium-resistance at levels greater than 3 {j.g ml"^ cadmium. Since degradation is
often inhibited in the presence of metal(s), the ability of cadmium-resistant isolates to
support 2,4-D degradation by A. eutrophus JMP134 was examined. The degradation of
2,4-D by A. eutrophus JMP134 was monitored in the presence of various cadmium levels
upon inoculation with one of the cadmium-resistant isolates. In the presence of >3 p.g ml*
' cadmiimi, A. eutrophus JMP134 alone could not degrade 500 fxg ml"' 2,4-D, whereas in
the absence of cadmium, A. eutrophus JMP134 degraded 500 (ig ml"' 2,4-D within 48 hr
at 28°C. Since the cadmium-resistant isolates alone were unable to degrade 2,4-D, and A.
eutrophus JMP134 only degraded 2,4-D in the absence of cadmium toxicity, the use of A.
eutrophus JMP134 provided an assay wherein cadmium toxicity could be directly
Pure culture: In replicate pure culture experiments, 25 ml of MSM buffered to pH
6.0 with 2-[morpholino]ethanesiilfomc acid (MES; Sigma Inc.) was amended with 500
^ig ml'" 2,4-D and either 12 or 24 ^ig ml'^ cadmium depending on the MRL of each
individual isolate. The potassium phthalate buffer used in the original medium recipe
was found to interfere with the 2,4-D absorbance readings and so MES was substituted
since it did not absorb at 230 nm. Each culture flask was inoculated with 10'' CPU ml"^
of either cadmium-resistant isolate Arthrobacter sp. D9, Arthrobacter sp. E92,
Pseudomonas sp. HI, Bacillus sp. H9, Fseudomonas sp. Ila or Pseudomonas sp. LI and
allowed to incubate at 28°C for 48 hr at 180 rpm. After 48 hr, the culture flasks were
inoculated with lO'^ CPU ml"' A. eutrophus JMP134. Concentrations of 2,4-D in cxiiture
extracts were measured every 24 hr at 230 nm following centrifugation at 14,000 x g for
2 min. to remove cell debris.
The relationship between 2,4-D concentration and
absorbance at 230 nm was linear with y = 0.03x - 0.07, r^ = 0.991.
Soil microcosms: Once established in pure culture, the ability of successful
isolates to protect A. eutrophus JMP134 from cadmium-toxicity was examined in
artificiaily metal-contaminated soil.
To determine if 2,4-D degradation could be
facilitated in a cadmium-contaminated soil, 100 g of an uncontaminated Brazito soil was
amended with 1% (w/w) glucose, 500 |j.g ml"' 2,4-D and 60 |^g ml"' cadmium (final
concentrations). Glucose was used as a readily metabolizable carbon source to support
the cadmium-resistant populations. Soil microcosms (500 ml wide-mouth polypropylene
jars) were incubated at 28°C and kept at 14% (w/w) soil moisture (75% of field capacity).
Control microcosms consisted of soil amended with glucose, 2,4-D and cadmium without
either inocula and soil amended with glucose, 2,4-D and cadmium inoculated with only a
cadmium-resistant isolate or A. eutrophus JMP134. The uncontaminated soil used was a
Brazito sandy loam with 12% clay, 0.21% organic matter, pH 8.2 and no known previous
metal exposxire. Indigenous microbial numbers in the soil were 3.2x10^ ± 9.1x10®
culturable CFU g'^ dry weight on R2A medium (Difco, Baltimore, MD) and
7.2xl0^db3.2xl0^ total cells g"' dry weight as determined by acridine orange direct
microscopic coimts.
Similar to the pure culture experiments, each soil microcosm was inoculated with
10"* CFU g"' of one of the cadmium-resistant isolates Arthrobacter D9, Pseudomonas HI,
Bacillus H9 and Pseudomonas Ila.
Following a 48 hr incubation, appropriate
microcosms were inoculated with lO"* CFU g'^ A. eutrophus JMP134. Concentrations of
2,4-D in soil extracts were measured daily for a total of 50 days. Soil extracts were used
for 2,4-D determinations. One-to-ten soil slvirries were made using 0.1% w/v sodium
pyrophosphate to neutralize soil particle charge, centrifliged at 14,000 rpm for 10 min
and read spectrophotometrically at 230 nm. Samples were analyzed in duplicate and the
soil microcosm experiment was performed three times. Soil samples without 2,4-D were
used as blanks to substract backgroimd absorbance.
Field bioreactors:
Laboratory soil microcosm studies were repeated at an
intermediate field scale. In the soil microcosms, when soils contaminated with 2,4-D and
cadmium were inoculated with the 2,4-D degrader A. eutrophus JMP134 and the
cadmium-resistant Pseudomonas HI isolate, biodegradation of 2,4-D occurred.
Bioreactors were set-up under field conditions to confirm laboratory microcosm results.
The field study was initiated in June 1998 and concluded in September 1998.
Five-gallon polypropylene bioreactors (1.5 ft x 2.5 ft), located at the University of
Arizona Campbell Avenue Agricultural Station, Tucson, AZ, were placed under a
constructed shaded area so as to preclude direct sunlight since daytime temperatures were
routinely in excess of 37.8°C (100°F). Each reactor contained approximately 27 Kg of
Brazito sandy loam at 14% w/w moisture content amended with 500 jag g"^ 2,4-D and/or
60 (xg ml'^ cadmium.
Cadmium-resistant Pseudomonas isolate HI and the 2,4-D
degrader A. eutrophus JMP134 were used as inoculants at lO'^ CFU g"^ dry weight soil.
Inoculants and soil amendments were thoroughly mixed into the soil prior to the
start of the experiment while providing a 48 hr incubation period between Pseudomonas
HI inoculation and A. eutrophus JMP134 addition. There were 7 treatments (Table 2)
each replicated twice for a total of 14 bioreactors. Soil moisture was maintained at 14%
w/w throughout the experiment. Ambient air temperature ranged from 16.7°C (62°F) to
48.9°C (120°F). Soil cores (18 in. x 1 in.) were collected weekly and analyzed for 2,4-D
concentrations. Background absorbance at 230 nm was monitored in reactors without
2,4-D amendment. Background absorbance was subtracted from each sample 2,4-D
reading. To eliminate cross-contamination, the soil corer was disinfected with 10%
bleach between samples.
The nimiber of culturable 2,4-D degrading microorganisms were enimierated on
an EMB medium containing the following per liter: 500 ^g ml"^ 2,4-D; 50 p-g ml"' yeast
extract; 80 mg eosin B; 13 mg methylene blue; 20 g Noble agar; 500 mis 2X MSB,
adjusted to pH 7.0. The 2X MSB contained the following: 0.224 g magnesium sulfate,
MgS04-7H20; 0.01 g zinc sulfate, ZnS04-7H20; 0.005 g sodium molybdate, Na2Mo04-
IHiO; 0.435 g dibasic potassium phosphate,
0.028 g calcium chloride, CaCl2-
2H2O; 440 fj.1 of 0.1% ferrous chloride, FeCl2-6H20; 1.0 g ammonixim chloride, NH4CI,
adjusted to pH 7.0. The acidity produced during 2,4-D degradation causes the eosin blue
to stain the colony dark purple. Previous studies have confirmed that the purple colony
appearance is indicative of 2,4-D degradation.
It is generally believed that organic degradation is inhibited in the presence of
metals presxmiably due to metal toxicity, however, few studies have actually addressed
this issue. The ultimate goal of this research was to find cadmium-resistant bacteria that
could either tolerate cadmium and degrade 2,4-D (used as a model organic) or could
detoxify cadmium to allow for another, metal-sensitive organism to carry out the organic
degradation. To examine this potential, four of six characterized bacterial populations,
resistant to greater than 12 jj.g ml'^ cadmiima, were used in 2,4-D degradation studies.
Solution Studies
In order for the cadmium-sensitive 2,4-D degrading A. eutrophus JMP134 to
survive and metabolize 2,4-D in the presence of cadmium, bioavailable cadmium
concentrations must be rendered nontoxic. The ability of six cadmium-resistant soU
isolates, Arthrobacter D9, Pseudomonas HI, Bacillus H9, Pseudomonas LI,
Arthrobacter E92 and Pseudomonas Ila (Table 1), to detoxify cadmium such that A.
eutrophus JMP134 could degrade 500 |i,g ml"^ 2,4-D was determined first in broth (Fig.
1) and subsequently in soil. Experiments showed that A. eutrophus JMP134 alone in the
presence of >3 jig ml*' cadmium did not degrade 2,4-D, presumably because of cadmium
toxicit}'. Likewise, in the presence or absence of cadmium, none of the cadmiumresistant isolates could degrade 500 ^xg ml"' 2,4-D. In MSM broth with 500 pig ml"' 2,4D and 10"^ CFU ml'' of one of the cadmiimi-resistant isolates, up to 12 ^g ml"' cadmium
was detoxified by isolate Pseudomonas Ila and 24 |ag ml"' cadmium was detoxified by
isolates Arthrobacter D9, Bacillus H9 and Pseudomonas HI within 120 hr, allowing for
complete degradation of the 2,4-D. Pseudomonas LI and Arthrobacter E92 (the least
cadmium resistant as presented in Table 1) were imable to detoxify cadmium imder these
The results of this experiment imply that these cadmium-resistant
populations detoxify cadmium through a sequestration or precipitation mechanism since
the toxic or soluble concentration of cadmium had to be reduced to allow for A.
eutrophus JMP\3A survival and metabolism.
Laboratory Soil Microcosms
In soil microcosms in the laboratory, artificially-contaminated Brazito soil was
amended with 500 fxg ml"' 2,4-D and 60 (ig ml"' cadmium. The same four isolates above
successfully detoxified cadmium, thereby protecting A. eutrophus JMP134 from
cadmium toxicity. Figs. 2 and 3 show the specific rates of degradation for each isolate.
Within 50 days, cadmium-resistant Pseudomonas HI and Pseudomonas Ila allowed for
the complete degradation of 500 |j.g ml"' 2,4-D (Fig. 2). Upon addition of cadmiumresistant Bacillus H9 and Arthrobacter D9, degradation occurred within 35 days (Fig. 3).
Interestingly, neither the indigenous microbial flora nor the cadmium-resistant isolates
could degrade 2,4-D in the soil system within the 50 day time frame. Additionally, A.
eutrophus JMP134, without the assistance of the resistant isolates, did not degrade 2,4-D
when cadmium was present, indicative of cadmium toxicity. In Brazito soil amended
solely with 2,4-D, complete 2,4-D degradation by A. eutrophus JMP134 occurred in 5
days. The difference in degradation rates for the 2,4-D only versus the 2,4-D and
cadmium amended soils was also indicative of cadmium stress. It should be noted that
the Brazito soil used in this study was not sterile and consequently presented competitive
challenges for the introduced organisms and, yet, degradation still occurred in the cocontaminated soils upon inoculation with the cadmium-resistant isolates.
In each of the degradation experiments when cadmimn was present, the 48 hr
incubation period between inoculation with a cadmiima-resistant isolate and A. eutrophus
JMP134 was critical. When both a cadmium-resistant isolate and A. eutrophus JMP134
were co-inoculated at the same time, no degradation occurred. Degradation was only
detected under co-contaminated conditions when at least 48 hr elapsed between
inoculations. The 48 hr period appeared to be necessary to allow for adequate cadmium
detoxification to occur.
Intermediate Field Scale
One problem with bench-scale studies is that they often do not translate
successfully to the field.
Therefore we tested the dual bioaugmentation strategy in an
intermediate field trial. In the intermediate field scale trial, as expected a great deal more
variability than in the bench-scale studies was evident (Table 2; Fig. 4). This variability
was due to the difficulty of ensuring uniform mixing of 2,4-D and cadmium into the large
soil mass utilized in each reactor. In spite of the variability, several conclusions can be
drawn from the field data. When 2,4-D was added to the Brazito soil without cadmiimi,
slow rates of degradation ultimately occurred without bioaugmentation with A. eutrophus
JMP134 (Fig. 4a), as the indigenous populations acclimated to the 2,4-D. However, even
after 10 weeks, degradation was incomplete and 2,4-D levels did not decrease between
weeks 5 and 10. No degradation (up to 50 days) was observed without augmentation in
the laboratory soil microcosm study, therefore the apparent degradation observed in the
absence of inoculation in the field may be the result of incomplete amendment mixing
that resxilted in some soil with low initial levels of 2,4-D. However, in the presence of 60
fj.g ml"' cadmium, indigenous degradation appeared to be inhibited in the absence of
bioaugmentation (Fig. 4b) even though some 2,4-D degraders were viable within the soil
after 8 weeks. Under all treatment conditions, in the absence of A. eutrophus JMP134
inoculation, 2,4-D degrading organisms did not appear until week 8. In the reactors
inoculated with A. eutrophus JMP134, during weeks 1 through 3, the number of 2,4-D
degrading organisms fell from the inoculated 10'^ CFU g"' of A. eutrophus JMP134 to
<10^ CFU g"'. By week 4 or 5, however, the nimiber of 2,4-D degraders increased
dramatically to 10^ CFU g"' (Figs. 4c-e). The introduction of A. eutrophus JMP134 into
2,4-D amended soil enhanced degradation in previous studies, found to be the restilt of
transfer of the pJP4 plasmid encoding 2,4-D degradation to indigenous soil recipients
(DiGiovanni et al., 1996), and so in this study, the rapid increase in 2,4-D degrading
organisms may correlate with the death of A. eutrophus JMP134 and subsequent gene
transfer. As observed in the soil microcosms, the reactors with 2,4-D and cadmium, coinoculated with cadmium-resistant
HI and^. eutrophus JMP134, exhibited
substantial degradation in conjunction with the appearance of > 10^ CFU g"' dry weight
2,4-D degrading organisms, suggesting that Pseudomonas HI confers a protective effect
(Fig. 4e).
Slow rates of degradation were observed in the Treatment 1 reactors (2,4-D only)
after 5 weeks with the appearance of 10^ indigenous 2,4-D degraders g'^ at week 8 (Fig.
4a). With cadmium present (Treatment 2, Fig. 4b) complete inhibitition of degradation
seemed to take place, even though indigenous 2,4-D degraders appeared at week 8
(approximately 10^ CFU g"'). Note that the EMB medium used to select for 2,4-D
degrading organisms did not contain cadmium. Consequently, while 2,4-D degrading
organisms appeared in the soil with 2,4-D and cadmium, 2,4-D degradation was
inhibited, presumably due to cadmium toxicity. For Treatment 3 (Fig. 4c) with 2,4-D and
A. eutrophus JMP134, as observed in all the reactors inoculated with A. eutrophus
JMP134, a 2,4-D degrading population was evident by week 5 (10^ CFU g"^). The
concentration of 2,4-D was reduced from 500 to 300 mg Kg"' in the first four weeks and
then the 2,4-D concentration remained stable.
Thus, degradation behavior did not
correlate with population development. It appears that the increase in 2,4-D degraders
occurred after 2,4-D degradation took place. We hypothesize that genetic transfer of
pJP4 occurred to indigenous recipients (DiGiovanni et al., 1996). These recipients,
however, lacked the chromosomal component for complete 2,4-D degradation resulting
in degradation to an aromatic intermediate still detectable by spectrophotometric
measxirements at 230 nm. In Treatment 4 reactors (Fig. 4d), in the presence of cadmium
and A. eutrophus JMP134, degradation was less apparent, though by week 6, 10^ 2,4-D
degraders g"' could be recovered, though cadmitmi appeared to have no significant effect
on 2,4-D degradation. As expected from both the pure culture and soil microcosm
experiments, cadmium-resistant Pseudomonas HI (Reactors 5 and 6) did not degrade (or
facilitate the degradation of) 2,4-D within the 70 day time frame of the field study (data
not shown).
In Treatment 7 reactors (2,4-D, cadmium, A. eutrophus JMP134 and
Pseudomonas HI), the extent of degradation was noticeably enhanced upon addition of
the co-inoculants (Fig. 4e). Degradation from 500 to 100 mg Kg'^ 2,4-D occurred by
week 6 in conjunction with the appearance of 10^ 2,4-D degraders g"'. Thus it appears
that initial inoculation with the cadmium-resistant isolate Pseudomonas HI allowed for
cadmium detoxification.
Once detoxified, less cadmium was available to inhibit
degradation by A. eutrophus JMP134 up to week 6. After week 6, 2,4-D levels remain
constant as a result of incomplete degradation by indigenous transconjugants. In the
presence of Pseudomonas HI, A. eutrophus JMP134 appeared to survive longer to
degrade more 2,4-D than in the treatments without HI.
This study has demonstrated the potential for a dual bioaugmentation strategy for
remediation of co-contaminated systems. This strategy involves co-inoculation of a
metal-resistant microbial population to allow detoxification of metal with a degrader to
facilitate removal of the organic. The primary mode of action being metal detoxification
such, that organic degradation was no longer inhibited. Examined in a field trial, we
demonstrated that bioaugmentation using co-inoculants may be a viable option for the
remediation of metal- and organically-contaminated soils.
Special thanks to Christine Stauber and Miriam Eaton for their assistance in the
field study. This work was supported in part by grant number 5 P42 ES04940-07 from
NIEHS, Superfund Program.
Table I. Cadmium-resistant soil bacterial isolates.
Soil Isolate
Arthrobacter sp. E92
Pseudomonas sp. LI
Pseudomonas sp. Ila
Arthrobacter sp. D9
Pseudomonas sp. HI
Bacillus sp. H9
Maximum Cadmium Resistance
(Hg ml"')
Table 2. Description of field bioreactor treatments.
(500 fag g"')
Amendments/I noculants
(10'^ C F U g ' )
Figure 1. In broth, cadmium-detoxification by isolates Arthrobactex D9, Bacillus H9,
Pseudomonas HI and Pseudomonas I la allowed for 2,4-D degradation by cadmiumsensitive A. eutrophus JMP134 in the presence of 12 |ig ml*' cadmium for isolate Ila and
24 ^.g ml"' cadmivun for isolates D9, HI and H9. Within 120 hr, all isolates allowed for
the degradation of 500 jj.g ml"' 2,4-D to undetectable levels. Each graph represents
replicate trials. Times as indicated were 48 hr following inoculation with the cadmiumresistant isolate.
Figure 2. Detoxification of 60 ji.g g"' cadmium by resistant Pseudomonas Ila and HI
allowed for 2,4-D degradation by cadmium-sensitive A. eutrophus JMP134 in laboratory
soil microcosms. Degradation of 500 |j,g g"' 2,4-D to undetectable levels occurred within
50 days.
Figure 3. Detoxification of 60 \ig g"' cadmium by resistant Bacillus H9 and Arthrobacter
D9 allowed for 2,4-D degradation by cadmium-sensitive A. eutrophus JMP134 in
laboratory soil microcosms. Degradation of 500 ^ig g"' 2,4-D to undetectable levels
occurred within 35 days.
Figiire 4. The degradation of 500 [ig g"' 2,4-D and the appearance of 2,4-D degrading
microbial populations with time in weeks as detected ia pilot field scale bioreactors
containing Brazito soil amended with (a) 2,4-D only; (b) 2,4-D and 60 p.g g"' cadmium;
(c) 2,4-D and lO'^ CFU g"' A. eutrophus JMP134; (d) 2,4-D, 60 \ig g"^ cadmiirai and lO'*
CFU g"^ A. eutrophus JMP134; (e) 2,4-D, 60 jag ml'^ cadmium, lO'^ CFU g"' A. eutrophus
JMP134 and lO'* CFU g"' cadmium-resistant Pseudomonas HI.
24 Hrs
48 Hrs
72 Hrs
120 Hrs
500 -
400 -
I 300
u 200 n
Q 100 -
200 100 -
24 |.ig m
12 |xg ml
)ig ml
12 lag ml"
Cadmium-Resistant Isolate
Figiire 2.
600 -
Pseudotnonas Ila
Pseudomonas HI
500 J
400 -
300 -
4 200
: r-:|
v-o >•
"*• *1
Figxrre 3.
Arthrobacter D9
600 -
I Bacillus H5
500 -
400 -
I 300
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100 -
(miitM up,I
CI f't
§ g § § 9 S g 8
u >
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DiGiovanni, G.D., J.W. Neilson, I.L. Pepper, and N.A. Sinclair. 1996. Gene transfer of
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While this research has raised many interesting questions, some of the more
significant ones that merit much further research are mentioned here. For example, why
did the cadmium-resistant isolates from
the metal-contaminated soils become
unculturable? We hypothesized that some of the microbial populations present in these
soils were sublethally injured to the extent that they could not be cultured on traditional
laboratory media. We looked to the high number of viable but nonculturable organisms
present in these soils as further evidence for their injury. It has been established that we
(as microbiologists) can culture <1% of the total number of microorganisms. However,
the extent of nonculturability, the cause of nonculturability (in this research, we assumed
metal toxicity), and the implications of nonculturability in the environment are not clear.
Another intriguing question is the environmental significance of metal resistance
mechanisms capable of responding to specific metal exposures. In this research, several
cadmium-resistant populations were became more resistant to cadmiimi as cadmiimi
toxicity increased, such that we were able to recover a higher number of resistant
organisms at higher cadmium concentrations. We attributed this observation to the
"activation" of a specific cadmium-directed mechanism of resistance at the higher
cadmium levels. As of yet no direct evidence for this has been foxmd, although a similar
observation has been made in organic degradation studies where a threshold level of
organic is necessary for activation of organic degradation genes. If a single organism can
use multiple mechanisms of metal resistance in response to metal stress, this has
important implications for metal remediation. If a specific organism is to be used to
remediate a site, then consideration of the metal concentration in the environment and the
needed metal "activation" concentration for the organism's resistance genes need to
match or remediation will not work.
Just how many mechanisms of metal resistance are available to microorganisms?
This research found four distiactly different mechanisms within one soil's community
alone and all in response to cadmium.
Doubly interesting is that each of these
mechanisms was found in isolates from
an uncontaminated soil with no previous
exposure to cadmium.
This finding alone raises the questions, "Why do microbial
populations with no apparent exposure to metal in their natural habitat have metal
resistance?", "How did/does microbial metal resistance originate, e.g., did it come about
with the evolution of life or did it come about with the relatively recent pollution of the
environment?" Finally, if at least four distinct mechanisms of cadmium resistance were
found in an imcontaminated environment, what would be the diversity of metal resistance
in a metal-contaminated environment? Answers to any of these questions would lend
valuable information to our understanding of microbial adaptation to the environment.
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sites encompassing large areas of land or where lower contaminant levels exist.
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