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20
Fungal-Based Remediation: Treatment of PCP
Contaminated Soil in New Zealand
J.M. Thwaites1, R.L. Farrell1, S.D. Duncan1, R.T. Lamar2 and
R.B. White2
1Department of Biological Sciences, University of Waikato, Private Bag 3105,
Hamilton, NEW ZEALAND, Email: r.farrell@waikato.ac.nz; 2Earthfax Development
Corporation, Logan, UTAH
1. Introduction
Contamination of soil, water, and air with toxic chemicals is a serious and on
going problem facing the world today. Hazardous compounds, such as
polycyclic aromatic hydrocarbons (PAHs), pentachlorophenol (PCP),
polychlorinated biphenyls (PCBs), 1,1,1-trichloro-2,2-bis(4-
chlorophenyl)ethane (DDT), and trinitrotoluene (TNT) are persistent in the
environment and are known to have carcinogenic and mutagenic effects.
Removing these pollutants from the environment in an ecologically responsible,
safe, rapid, and cost-effective way is a priority for land management agencies.
Bioremediation, using microbial organisms, is one way to achieve this target.
Extensive laboratory studies have shown the capability of various organisms to
remediate contaminated soil and water. More research, however, to determine
the applicability and practicability of utilizing these microorganisms in
contaminated field sites needs to be achieved. This review and case study
presents an evidence of successful bioremediation of PCP and related dioxins
using fungal-based technology.
2. Fungal-based Remediation
Fungal-based remediation is an ex situ form of bioremediation, in which
hazardous organics are degraded or detoxified by fungi that are introduced into
the contaminated soil via a fungal inoculum (i.e. a lignocellulosic substrate
support which is colonized by the fungus). The soil inoculum mixture is then
treated in a forced aeration biopile, in which temperature and moisture conditions
are monitored and maintained to provide optimum fungal growth and activity.
J.M. Thwaites et al.
466
Fungi are robust organisms that are tolerant to high concentrations of
pollutants (Evans and Hedger 2001). Phanerochaete chrysosporium was the
first reported microorganism to show degradation of an extremely diverse group
of environmental pollutants (Bumpus et al. 1985; Eaton 1985). Since then, the
majority of research on bioremediation employed the pollutant-degrading
abilities of an ecological group of fungi, including P. chrysosporium, referred to
as white rot fungi. These fungi are saprophytes that obtain their carbon for
energy and biomass from the dead organic matter, and include members, such
as the common edible mushrooms, Pleurotus ostreatus (oyster mushroom),
Lentinulus edodes (Shitake), and Agaricus bisporus (white button mushroom).
White rot fungi degrade cellulose, hemicellulose and most importantly the
lignin component of the wood cell wall. After degradation, the residual wood is
typically fibrous with a whitish yellow to tan discoloration due to the removal
of lignin. Most white rot fungi are basidiomycetes, possessing dikaryotic
hyphae and clamp connections along the septate hyphae (Jennings and Lysek
1999). These fungi are uniquely equipped as soil remediation agents (Reddy
1995). As filamentous organisms, they have the natural propensity to extend
through soil in search of new substrates to exploit, and thus can colonize places
that bacteria are unable to reach. They possess the ability to oxidize extremely
hydrophobic substrates due to the highly oxidative nature of the enzymes that
comprise the extracellular component of their lignin-degrading system. These
extracellular enzymes extend the fungus degradative influence beyond the
hyphae. The fungus does not utilize the pollutant for grwoth, and thus, the
amount of pollutant degraded is not a function of the concentration of fungus
within the soil.
Lignin has a heterogeneous aromatic structure with many different types of
subunit linkages. Both direct (pollutant oxidation by lignin-degrading enzymes)
and indirect (pollutant mineralization during ligninolysis) evidences indicated
that the lignin-degrading systems of white rot fungi were involved in pollutant
degradation (Kirk and Farrell 1987). This enzymatic degradation, termed
“Enzyme Combustion”, is highly oxidative, extracellular and non-specific (Kirk
and Farrell 1987). In addition to oxidative agents, the fungi possess reductive
agents that are also involved in the degradation of aromatic sub-structures of
lignin that are produced from its depolymerization. It makes perfect sense then
that the aromatic pollutants, such as PCP and PAHs, that are degraded by white
rot fungi, closely resemble the aromatic sub-structures that are produced during
lignin depolymerization.
Much of the work on fungal-based remediation has been conducted using
PCP. PCP was used as a wide spectrum pesticide and wood preservative
throughout the world. In New Zealand, the primary use was in the timber
industry as a treatment for the interim protection of timber against sapstain
fungi and as a preservative in diesel oil. It was estimated that over a period of
forty years, 5000 tonnes of PCP was used in New Zealand alone (Yu and
Shepard 1997). Significant use of PCP, in New Zealand, ceased in 1988 and the
Fungal Bioremediation of PCPs 467
chemical was formally deregistered by the New Zealand Pesticides Board in
1991. It is currently banned in most countries and is listed as a priority pollutant
by the United States Environmental Protection Agency. Soil contamination
through accidental spillage and inappropriate disposal at many wood treatment
facilities plus the relative stability of PCP in the environment means that it
continues to be a major problem.
PCP is toxic to organisms because it is an inhibitor of oxidative
phosphorylation (Crosby 1981) and it is also hydrophobic with low water
solubility; this contributes to its persistence in the environment. The products of
the PCP manufacturing process, polychlorinated dibenzo-p-dioxins and
polychlorinated dibenzofurans (PCDFs), are also a problem at many sites where
PCP was manufactured or used (Eduljee 1999). The major dioxin congeners
identified are the hexa-, hepta-, and octachloro congeners (Eduljee 1999) and all
are highly persistant and toxic in the environment.
2.1 Fungal Degradation of PCP
White rot fungi have been shown to cause a considerable depletion of PCP from
soil in the laboratory experiments (Lamar et al. 1990a, b; Okeke et al. 1993;
Okeke et al. 1994; Lestan et al. 1996) and under field conditions (Lamar and
Dietrich 1990; Lamar et al. 1993; Lamar et al. 1994). In several small-scale
field trials (total weight of soil for all treatments < 24 tonnes) significant but
less extensive PCP decreases (88-91%) were reported (Lamar and Dietrich
1990). In this study, Phanerochaete sordida and P. chrysosporium were
inoculated into a sandy gravel soil pH 9.67, contaminated with a commercial
PCP containing wood preservative (250-400mg PCP/kg soil) and the reductions
were evident after 6.5 weeks.
2.2 Fate of PCP after Bioremediation
Numerous publications have shown that white rot fungi can efficiently deplete
PCP in the contaminated soil, but the question of the fate of the portion of PCP
that is converted to non-extractable products remains unanswered (Lamar et al.
1990a,b; Lamar and Dietrich 1990; Lamar et al. 1993). In liquid fungal culture,
chlorinated phenols, were shown to be degraded via a series of reactions that
remove all the chlorines, after which the aromatic ring is oxidatively cleaved
and further degraded to CO2 (Valli and Gold 1991; Joshi and Gold 1993). In
soil, mineralization (degradation to CO2 and H2O) was demonstrated to be a
minor fate for PCP. Work by Rüttiman-Johnson and Lamar (1997)
demonstrated that a large part of PCP present in the contaminated soil is bound
to the soil humic materials by the action of extracellular oxidative enzymes. The
covalent binding of pollutants to soil fractions is important as it may reduce
their bioavailability and therefore their toxicity. The oxidized transformation
J.M. Thwaites et al.
468
products of xenobiotics, like PCP (Rüttimann-Johnson and Lamar 1997), PAHs
(Eschenback et al. 1998; Kastner et al. 1999), and chlorocatechols (Stott et al.
1983), can be readily incorporated into soil humic materials (a phenomenon
referred to as humification) because of their structural similarity to natural
aromatic substrates originating mainly from lignin degradation. The underlying
mechanism of the humification process involves oxidation of the chlorinated
phenols or other aromatics to free radicals or quinone products that
subsequently couple directly to humic acids, fulvic acids and/or humin, all of
which possess oxygen-containing functional groups (hydroxyls, carboxyls, and
quinones) or to naturally-occurring phenols that are also subject to oxidation. If
the unpaired electron of a free radical is located at an aromatic carbon which is
substituted by a chlorine atom, dehalogenation occurs during the coupling
reaction (Hatcher et al. 1993). The occurrence of dehalogenation during
oxidative polymerization provides a direct evidence for the formation of
covalent bonds between the chlorophenol transformation product and humic
acid. Covalent binding is considered the strongest type of bonding, and
therefore, together with the dehalogenation process, is a desired reaction for the
removal of chlorinated phenols, and other aromatic xenobiotics from the
environment.
Phenol oxidases are produced by white rot fungi, are present in terrestrial
systems and enriched in fungal-based remediation systems (Stevenson 1994).
The ability of oxidases to mediate oxidative coupling reactions has been
demonstrated in a number of experiments in which selected xenobiotics were
combined with humic monomers or natural humic acids in the presence of a
phenoloxidase. Rüttmann-Johnson and Lamar (1996) reported that high
molecular weight humic polymers were produced when a reaction mixture
containing PCP and fulvic acid (a humic/fulvic acid precursor), in the presence
of a surfactant and H2O2, were mixed with a crude, concentrated supernatant
from cultures of P. chrysosporium. Pure polyphenoloxidases (manganese-
dependent peroxidase, lignin peroxidase, and laccase) also catalyzed the
reaction. Humification of oxidized transformation products of 2,4-DCP and
other chlorinated phenols in the presence of humic acid precursors (e.g. syringic
acid) and laccase was demonstrated by Bollag and colleagues (1980). They
showed the production of hybrid polymers in which the aromatic ring of
oxidized 2,4-DCP becomes an integral part of the humic polymer. As the
complexity and the size of humic polymers increases, the stability of the
aromatic ring that was associated with the original xenobiotic also increases, as
additional covalent bonds are formed with other aromatic structures through
further oxidative polymerization reactions.
While in vitro experiments do not provide a direct evidence for covalent
binding, they do demonstrate that fungal phenol oxidases are capable of
mediating the covalent bonding of xenobiotic transformation products with
humic materials. Direct evidence for the covalent nature of bonds formed
between oxidized transformation products of chlorinated phenols and humic
Fungal Bioremediation of PCPs 469
compounds produced in phenol oxidase-mediated reactions, comes from
analyses of the hybrid polymers by MS and NMR spectrometric methods.
Hatcher and colleagues (1993) investigated the horseradish peroxidase-
mediated bonding of 13C-labeled 2,4-DCP to natural humic acid. After
incubation, the humic acid was analyzed using 13C NMR spectroscopy. The
NMR spectrum of the humic acid displayed nine major signals that did not
appear in the spectrum of the free 2,4-DCP that was just mixed with the humic
acid. Evaluation of the signals revealed that the oxidized product of the 2,4-
DCP molecules were covalently bound to humic acid through ester, phenolic
ether and carbon-carbon linkages.
Covalent bonding of the oxidized transformation products of aromatic
xenobiotics, like the p-chlorobenzoquinones of chlorophenols, to humic
materials, a process that results in the aromatic structures of the original
xenobiotics becoming a part of the structure of natural humic polymers,
eliminates the bioavailability and thus the toxicity of the original xenobiotics.
The decreased bioavailability is a result of the high molecular weight of the
polymers which makes them to large to pass through cell membranes. Two
phases are observed during the microbial turnover of natural biopolymers in
soils. A first phase involving rapid metabolism of parent compounds (e.g.
lignin, cellulose, and in contaminated soils, aromatic xenobiotics) is followed
by a second phase of slow turnover of derived humic material. The
mineralization of formed humic substances decreases to mean turnover rates of
2-7% per year after 250-300 days (Alexander 1977; Führ et al. 1985). With
increasing age of the material, the formed molecules become increasingly inert,
leading to humic substances with long-term stability. Age analysis of humic
substances revealed residence times of several hundred years (Haider 1996).
Eschenback and colleagues (1998) investigated the fate and stability of non-
extractable residues of 14C-labeled PAHs in contaminated soils under
environmental stress conditions. The work described in their paper is worth
examining in detail, because it directly addressed the long-term stability of
humified aromatic xenobiotics in soils. The experiments were conducted using
non-extractable [14C]-PAH residues that were produced in the preceeding long-
term bioremediation experiments using the white rot fungus P. ostreatus
(Eschenback et al. 1995). Soil samples from these experiments bearing non-
extractable residues of oxidized transformation products of [14C]-naphthalene,
[14C]-anthracene, [14C]-pyrene, or [14C]-benzo[a]pyrene were treated by
biological, physical, or chemical treatments. The effect of the various treatments
was assessed by comparing first, 14CO2 mineralized; second, extractable 14C-
activity; and third, 14C associated with non-extractable residues in treated and
non-treated soils. Biological treatments consisted of inoculating the soils with
selected humus-degrading fungi or bacteria, or amending the soils with easily
metabolized carbon sources (glucose and starch to initiate a “priming” effect of
indigenous soil microbes). Neither, the addition of humus-degrading microbes,
or easily metabolized carbon sources, led to an increase in 14CO2 or extractables
J.M. Thwaites et al.
470
and thus potentially mobile [14C]-PAH residues. The transformation activity in
the various treated soils (including non-treated soils), as based on mineralization
activity, approached, in all cases, similar levels of continuous but very low
14CO2 release during the first 100 days of incubation. More importantly, the rate
of mineralization (release of 14CO2) was comparable to humus turnover rates in
soil (2-5 % per year (Saxena and Bartha 1983)). This information, coupled with
no change or decreases in the amount of extractable 14C, indicated that the 14C,
that was released from non-extractable residues was rapidly metabolized.
Physical stress factors including frost, rapid temperature variations, and drying
and rewetting of soil also did not have any remarkable effect on the stability of
the non-extractable 14C residue fraction. The effect of a chemical change on the
stability of the non-extractable [14C]-PAH residues was evaluated by using a
complexing agent (EDTA) to disrupt the metal-organic complex. EDTA
treatment was accomplished by extracting dry soil samples containing non-
extractable [14C]-PAH residues, with EDTA solutions of varying
concentrations. Application of EDTA did result in the release of 14C-activity
into the soil solution from the non-extractable fraction. It was shown that this
activity was due to dissolved organic matter-[14C]-PAH residue complexes that
were released as a result of EDTA complex of metals, thus disrupting metallo-
humic complexes and releasing [14C]-residues into the soil solution. This type of
treatment would be extremely unlikely to occur naturally or even as a result of
anthropogenic activities. As discussed above, humic materials in soil are
extremely stable over time, their stability increases with age and they are
extremely difficult to mobilize due to their high molecular weights and
hydrophobicity.
It has been shown in previous studies both in the laboratory and in field
situations with P. chrysosporium and P. sordida that a small percentage of the
decrease in the amount of PCP was the result of fungal methylation of PCP to
pentachloroanisole (PCA). Both bacteria and fungi have been shown to
methylate chlorophenol compounds to their methylated derivatives. Chung and
Aust (1995) speculated that methylation of PCP to PCA may be a detoxification
mechanism since PCA is not an inhibitor of oxidation phosphorylation and less
toxic than PCP to wood rotting fungi, other microbes and fish. The solubility of
PCA (~0.2 ppm) is less than that of PCP (2.5 ppm), thus preventing the
contamination of groundwater. Lamar and Dietrich (1990) found that only about
9 to 14% of the PCP was converted to PCA. Thus, methylation was not the
major route of PCP depletion in the contaminated soil. It was reported that the
PCA was accumulated by these fungi during the initial bioremediation period
after which it decreases slowly (Lamar and Dietrich 1990; Lamar et al. 1990b
and 1993). The use of fungal stains, that do not produce significant amounts of
PCA, would be advantageous for bioremediation purposes. Only trace amounts
of PCA and 2,3,4,6-tetrachloroanisol (2,3,4,6-TeCA) were formed during the
laboratory based remediation of PCP contaminated soils with the fungus
Trametes versicolor (Tuomela et al. 1999). A part of the 14C-label was alkali-
Fungal Bioremediation of PCPs 471
extractable, indicating that it was bound to humic substances, but this in part
was apparently later attacked and mineralized by the fungus. The production of
the enzyme laccase by this fungus was thought to enhance the degradation of
PCP to other compounds rather than PCA (Tuomela et al. 1999). Lestan and
colleagues (1996) also found a negative correlation between manganese
peroxidase activity and PCA production by T. versicolor, indicating that this
enzyme may also be involved in desirable PCP removal from the contaminated
sites.
2.3 Biodegradation of Dioxins and Furans
White rot fungi have also been shown to degrade PCDDs and PCDFs in
aqueous culture (Bumpus et al. 1985; Valli et al. 1992; Takada et al. 1996;
Rosenbrock et al. 1997). Rosenbrock et al. (1997) evaluated the mineralization
of undifferentiated dibenzo-p-dioxins in soil using four species of white rot
fungi (P. chrysosporium,Pleurotus sp., Dichomitus squalens, and an
unidentified fungus isolated from the site). They found that mineralization of
PCDDs was greatly enhanced by inoculation with white rot fungi. Over a 70
day period, the extent of mineralization varied from 30 to 55% depending on
the soil and the fungal species. Valli and Gold (1991) also found that both 2,4
dichlorophenol and 2,4,5-trichlorophenol may be degraded by P. chrysosporium
by a complex pathway, involving oxidative displacement of chloride and O-
methylation with the formation of 1,2,4,5-tetrahydroxybenzene before ring
fission.
2.4 Fungal-based Remediation for Treatment of Contaminated Soil in New
Zealand
In a scientifically-controlled experiment, the ability of fungal-based
remediation to decrease the concentration of PCP and selected dioxin and
furan congeners was evaluated in soil samples collected from a former dip
tank wood-treating operation in Whakatane, New Zealand. The study was
conducted using two New Zealand strains of white rot fungi, and two other
fungi which were isolated from soil at the site (provided from The University
of Waikato Fungal Culture Collection). In addition, the degradation was
evaluated, for the purpose of comparison, using one strain of a fungus
typically used for the remediation of similar compounds in the United States.
All fungal strains were grown on locally available lignocellulosic substrates
(Pinus radiata or Eucalyptus sp. wood chips). Technology performance was
based on the percent decreases in the concentrations of 1,2,3,4,6,7,8-
heptachlorodibenzo furan (HpCDF), 1,2,3,4,6,7,8-heptachlorodibenzo dioxin
(HpCDD), and octachlorodibenzo dioxin (OCDD). These contaminants will
be collectively referred to as the “analytes.” The following factors were
J.M. Thwaites et al.
472
evaluated: fungal species, inoculum application rate, and surfactant addition.
The effectiveness of the various treatments was evaluated by determining the
concentrations of the analytes immediately after treatment application, after
14, 28 and 56 days of treatment.
A “representative” soil sample was obtained from the site. The soil was
air-dried and sieved to pass a 2-mm screen and thoroughly mixed. The soil
was then stored dry in a sealed container until use. The concentrations of
target chemicals were determined on soil sub-samples using appropriate
extraction and analytical techniques. As the regulatory drivers in this soil were
the dioxin congeners and these compounds are extremely hydrophobic, we
evaluated the effect of amending the contaminated soil with a surfactant to
enhance their degradation. The surfactant that was evaluated was emulsified
soybean oil (ESO). The ESO was mixed with the water that was used to adjust
the moisture content of the soil to provide homogeneous distribution of the
surfactant and was applied at a rate of 3% (weight of oil to dry weight of soil).
A total of five fungal species were evaluated. The four New Zealand strains
(provided by The University of Waikato Fungal Culture Collection) were
Phanerochaete gigantea, Resinicium bicolor, and two fungi isolated from the
site soil referred to as the “East side” and “West side” strains based on where
the soil they were isolated from was in relation to the former dip tank
location. For comparative purposes, we also evaluated a United States strain,
Pleurotus ostreatus. The latter fungus has a demonstrated ability to degrade
PCP and dioxins in soil. Fungal inoculum was prepared by cultivating pure
cultures of each of the fungi on sterilized P. radiata and/or Eucalyptus sp.
wood chips. The moisture content of the chips was adjusted to 60% (wet
weight basis) and then they were sterilized by autoclaving at 15 psi and 121oC
for one hour on two successive days. The chips were inoculated with a
mycelial slurry inocula produced from liquid cultures (2% glucose and 2%
malt extract) of each fungal species. The inoculated chips were incubated at
30oC until they were thoroughly colonized (about 2 weeks).
The soil treatments were conducted in 8 oz. (272 ml) canning jars with lids
modified to allow adequate air exchange. Each jar contained approximately 30
g of the test soil (i.e. wet weight) and the appropriate amount of fungal
inoculum and amendments. Three replicates were prepared for each treatment
for each sample time, with the exception for day 0. For day 0, a sample was
prepared on the side for each treatment from which 2 sub-samples were taken
for analysis. The cultures were incubated at 30oC (this would be the optimum
biopile temperature) under high relative humidity to prevent moisture loss.
Soil moisture content was maintained as needed. Target compound
concentrations were evaluated on the following days: 0, 14, 28, and 56.
Soil and soil inoculum mixtures from each experimental unit were air-dried
in plastic weigh boats and then ground to a fine powder using a commercial
coffee grinder. The ground samples were stored dry in sealed glass containers.
To determine the concentrations of PCP, HpCDD, HpCDF, and OCDD, 3 g
Fungal Bioremediation of PCPs 473
sub-samples from each sample were extracted with a 50:50 mixture of hexane
and acetone with a Dionex Accelerated Solvent Extractor. Sub-samples of the
extracts were then analyzed using GC/ECD methods to determine extract
concentrations of the analytes. PCP was analyzed as the trimethylsilyl
derivative. PCP in extract sub-samples was derivitized using Sylon BTZ
(Supelco Co.). GC/ECD analyses of derivatized extracts were performed on a
Hewlett-Packard model 5890 gas chromatograph equipped with a 63Ni
electron capture detector, a model 7673A autosampler, and a split-splitless
capillary column injection port. Gas flows were: column flow 2 ml/min; total
flow 60 ml/min. Operating temperatures were: 220oC (injector) and 300oC
(detector); the carrier and makeup gas was nitrogen. The column was a DB-5
fused silica capillary column (30 m by 0.321mm; film thickness 0.25 um).
The temperature program was as follows: initial 60oC; hold for 1 min; split off
for 0.5 min; ramp A, 10oC/min for 9 min (60 to 150oC); ramp B, 2oC/min for
20 min (150 to 190oC); hold at 190oC for 5 min. GC/ECD analysis of extracts
for HpCDD, HpCDF and OCDD were performed on the same instrument
using the following conditions: Gas flows were: column flow 2 ml/min; total
flow 30 ml/min. Operating temperatures were: 280oC (injector) and 300oC
(detector); the carrier and makeup gas was nitrogen. The column was a DB-5
fused silica capillary column (30 m by 0.321mm; film thickness 0.25 um).
The temperature program was as follows: initial 185oC; hold for 2 min; split
off for 0.5 min; ramp A, 8oC/min for 8 min (85 to 285oC); hold at 285oC for 8
min.
Analyses of variance (ANOVA), using a = 0.05, were performed on the
percent difference between concentrations of the analytes on day 0 and day
56. The main effects included in the ANOVA were fungal treatment,
inoculum application rate and surfactant addition.
Initial concentrations after treatment applications are given in Table 1.
There was significant variation in initial analyte concentrations among the
treatments for all four analytes. This is an indication of the heterogeneity of
the soil with respect to contaminant concentrations.
Fungal inoculation had a significant effect on the mean percent decreases
of all four analytes among fungal inoculation treatments (Table 2). In all
cases, inoculation with any of the tested fungi resulted in a significantly
greater decrease than no inoculation (control). Among the tested fungi, the
greatest percent PCP decrease occurred in soils inoculated with the “East
side” strain. There were no significant differences among the fungal
treatments in the degradation of HpCDF and HpCDD. Average percent
decrease of these compounds was greater than 90% in all fungal treatments.
Degradation of OCDD was greatest in soils inoculated with P. gigantea
(Table 2). The percent OCDD decrease in all other fungal inoculated soils was
less, significantly so, in soils inoculated with R. bicolor.
J.M. Thwaites et al.
474
Table 1. Initial concentrations of PCP, HpCDF, HpCDD, and OCDD immediately after
treatment application
Treatment PCP
(mg/kg) HpCDF
(Pg/kg) HpCDD
(Pg/kg) OCDD
(Pg/kg)
Control 83 313 135 472
P. ostreatus 92 262 189 508
“East side” 182 340 351 743
“West side” 115 378 331 1045
R. bicolor 154 323 307 644
P. gigantea 136 356 380 792
Table 2. Effect of fungal inoculum and control treatments on mean1 percent decrease of
PCP, HpCDF, HpCDD, and OCDD after 56 days of treatment
Treatment PCP HpCDF HpCDD OCDD
Control 15.6c 5.4b (33.3)b (22.4)c
P. ostreatus 75.2b 98.5a 97.8a 82.1ab
“East side” 90.3a 95.7a 95.9a 69.3ab
“West side” 75.7b 97.0a 92.4a 81.0ab
R. bicolor 83.5ab 95.0a 91.6a 68.2b
P.gigantea 76.6ab 93.7a 91.5a 86.2a
1Means within columns followed by the same letter are not significantly different
Mean concentrations of all four analytes after 56 days of treatment were
significantly less in fungal inoculated treatments compared to control
treatments (Table 3). The lowest residual PCP concentration occurred in soils
inoculated with the “East side” fungus. There were no significant differences
among the fungal treatments in residual concentrations of HpCDF and OCDD.
The lowest residual concentration of HpCDD occurred in soil inoculated with
P. ostreatus. However, as with HpCDF, and OCDD, all the fungal treatments
resulted in very extensive decreases in the concetration of HpCDD. The rate
of fungal inoculation did not have a significant effect on the average percent
decrease of any of the four analytes (Table 4). Application of ESO had no effect
Table 3. Mean1 fungal inoculum treatment concentrations of PCP, HpCDF, HpCDD,
and OCDD after 56 days of treatment
Treatment
(mg/kg) PCP
(Pg/kg) HpCDF
(Pg/kg) HpCDD
(Pg/kg) OCDD
(Pg/kg)
Control 70c 263b 264c 557b
P. ostreatus 28b 4a 3a 98a
“East side” 13a 12a 12ab 210a
“West side” 28b 15a 30b 196a
R. bicolor 22ab 14a 24b 188a
P. gigantea 32b 21a 30b 95a
1Means followed by the same letter are not significantly different
Fungal Bioremediation of PCPs 475
Table 4. Effect of inoculum application rate on mean treatment percent decrease for
inocuolum application rate of PCP, HpCDF, HpCDD, and OCDD after 56 days of
treatment
Decrease (%)
Inoculum application
rate (wt inoc/wt soil) PCP HpCDF HpCDD OCDD
10 83.7 96.1 92.6 78.9
20 77.1 95.9 95 75.8
on the mean inoculum application rate percent decrease of PCP, but
significantly decreased the percent degradation of HpCDF, HpCDD, and
OCDD (Table 5).
Table 5. Effect of ESO application rate on mean percent decrease of PCP, HpCDF,
HpCDD, and OCDD after 56 days of surfactant treatment
Decrease (%)
ESO addition rate PCP HpCDF HpCDD OCDD
0 71.8a 91.3a 82.9a 79.0a
3 76.8a 84.2b 81.6b 57.6b
The treatment combination, that resulted in the greatest overall percent
decreases of the four analytes (386.4%), was inoculation with P. ostreatus using
an inoculum application rate of 10% and augmentation of the soil with 3% ESO
(Table 6). The second most effective treatment, with a total percent decrease for
the four analytes of 371.7%, was inoculation with the “East side” strain at an
inoculum application rate of 10% in the presence of 3% ESO (Table 6). Based
on the degradation of PCDD/PCDFs only, the most effective treatments were
inoculation with P. ostreatus at a rate of 10% with or without ESO and
inoculation with the “West side” isolate at a rate of 20% with or without ESO.
Because similar results were obtained with or without ESO, it would not be
necessary to use it in the field.
3. Conclusion
Inoculation of PCP/PCDD/PCDF-contaminated soils with selected isolates of
white rot fungi and fungi from New Zealand, grown on locally available radiata
pine or eucalyptus pulpwood chips, resulted in the rapid and extensive
decreases in the concentrations of the contaminants. In particular, treatment
with either of two fungal species isolated from PCP/PCDD/PCDF-contaminated
soil from around the former dip tank at the Whakatane site, effectively
decreased the concentrations of the PCP, HpCDF, HpCDD, and OCDD. Based
on these results, the use of fungal-based remediation of the treatment of
New Zealand soils contaminated with PCPs and associated PCDDss/PCDFs has
J.M. Thwaites et al.
476
Table 6. Percent decreases in PCP, HpCDF, HpCDD, and OCDD concentrations in
fungal inoculation/inoculum application rate/surfactant addition rate treatments
Treatment PCP HpCDF HpCDD OCDD % sum
A11 %sum
B22
No inoculation/0/0 13.6 45.6 (12.4) 6.7 53.5 39.9
No inoculation/0/3 17.6 (34.8) (54.3) (51.5) (123.0) (105.4)
P. ostreatus/10/0 77.3 99.4 96.4 86.0 359.1 281.8
P. ostreatus/10/3 87.1 99.8 99.8 99.7 386.4 299.3
P. ostreatus/20/0 65.2 96.2 91.3 93.0 345.7 280.5
P. ostreatus/20/3 71.2 98.3 99.8 49.5 318.8 247.6
East side/10/0 98.3 96.9 91.5 64.5 351.2 252.9
East side/10/3 99.7 98.1 97.2 76.7 371.7 272.0
East side/20/0 77.2 94.2 97.1 83.9 352.4 275.2
East side/20/3 86.2 93.5 97.6 52.2 329.5 243.3
West side/10/0 75.5 94.0 89.6 85.3 344.4 268.9
West side/10/3 82.5 98.2 93.1 52.3 326.1 243.6
West side/20/0 70.1 99.4 96.5 90.6 356.6 286.5
West side/20/3 74.8 96.4 90.3 95.8 357.3 282.5
R. bicolor/10/0 73.8 97.3 92.7 87.6 351.4 277.6
R. bicolor/10/3 88.7 90.7 86.5 54.9 320.8 232.1
R. bicolor/20/0 92.3 95.9 91.2 74.0 353.4 261.1
R. bicolor/20/3 79.4 96.2 96.2 56.2 328.0 248.6
P. gigantea/10/0 72.2 89.3 84.1 99.8 345.4 273.2
P. gigantea/10/3 77.7 96.6 95.0 73.4 342.7 265.0
P. gigantea/20/0 74.9 96.1 94.2 89.8 355.0 280.1
P. gigantea/20/3 80.3 92.9 92.6 73.6 339.4 259.1
1the sum of the percent decreases of all four analytes
2the sum of the percent decreases of HpCDF, HpCDD, and OCDD
excellent potential. Work has been undertaken to demonstrate the effectiveness
of fungal-based remediation, using the fungal strain ‘Eastside’, at pilot-scale
and further developmental work is underway to upscale this technology for
application at a full scale commercial basis. On the basis of investigations,
following conclusions may be drawn:
1. Fungal inoculation greatly stimulated the degradation of PCP, HpCDD,
HpCDF, and OCDD in the Whakatane soil.
2. Two New Zealand wood decay basidiomycetes and two unidentified fungi
isolated from site soil performed similarly to a US strain (i.e. P. ostreatus)
that has proven previously to be an effective degrader of PCP and
PCDD/PCDFs.
3. Based on PCP-degrading performance alone, the most effective treatment was
inoculation with the ”East side” strain at a rate of 10% without addition of ESO.
Fungal Bioremediation of PCPs 477
4. Based on PCDD/PCDFs-degrading performance alone, the most effective
treatment that included a New Zealand fungal strain was inoculation with the
“West side” strain with or without ESO. It would not be worthwhile to use
the ESO given the added cost in materials and labor.
Acknowledgements. We thank sincerely for his expertise and contributions to
the project Harold H. Burdsall, Jr., PhD, Mycologist Expert of Black Earth
Wisconsin, formerly with Forest Products Laboratory, Madison, Wisconsin,
USA.
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