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Table of Contents
Abstract
1. Introduction
2. Background
2.1 EDCs in the San Francisco Bay
2.2 EDC transportation through water treatment plants
2.3 How mycoremediation removes toxins
2.4 How phytoremediation removes toxins
2.5 How other remediation techniques remove toxins
2.5.1 Bacteria bioremediation
2.5.2 Plankton (phyto- and zoo-)
2.5.3 Chemical/Physical
2.5.3.1 Adsorption
2.5.3.2 Fenton Reagent
2.5.3.3 Hydrogen Peroxide
2.5.3.4 Membrane processes
2.5.3.5 Ozone
2.5.3.6 Photocatalysis
2.5.3.7 Sonolysis
2.5.3.8 Ultraviolet light
3. Methods
4. Results and Discussion
4.1 Which EDCs remediation will focus on
4.1.1 BPA
4.1.2 NP
4.1.3 TCS
4.2 Analysis of mycoremediation methods
4.2.1 Direct fungal application methods
4.2.2 Enzyme application methods
4.2.2.1 Free Enzymes
4.2.2.2 Immobilized Enzymes
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4.2.3 Metabolites found
4.3 Analysis phytoremediation methods
4.3.1 Plant application methods
4.3.2 Metabolites found
4.4 Best myco- and phytoremediation methods
4.4.1 Viability of mycoremediation system
4.4.2 Viability of phytoremediation system
4.5 Pilot studies
4.5.1 Mycoremediation pilot studies
4.5.2 Phytoremediation pilot studies
5. Conclusion and Recommendations
5.1 Summary of findings
5.2 Limitations and future work
5.3 Practical implications
6. Appendices
6.1 Appendix 1: EDC removal by fungi and their enzymes from aqueous solutions
6.2 Appendix 2: EDC removal by plants from aqueous solutions
7. Acronyms
8. References
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Abstract
Endocrine disrupting chemicals (EDCs) discharged from wastewater treatment effluent
are an environmental and anthropogenic concern. Aqueous EDCs can be removed in wastewater
treatment plants (WWTPs) by bioremediation. The goal of this paper is to compare
mycoremediation and phytoremediation treatment methods for use in a San Francisco bay area
WWTP. Reviewing mycoremediation studies on removal of the EDCs bisphenol A (BPA),
nonylphenol (NP), and triclosan (TCS), it is concluded that immobilized laccase in the form of
cross-linked enzyme aggregates (CLEAs) from the fungi Coriolopsis polyzona would be the best
mycoremediation method. Analysis of studies examining phytoremediation of BPA, NP, and
TCS suggest that constructed wetlands (CWs) with the species Portulaca oleracea, Landoltia
punctata, and Lemna minor would be the superior phytoremediation method. After a comparison
of these two methods by modeling with Hydrologic Engineering Center-River Analysis System,
it is recommended that mycoremediation be used for its superiority in efficiency, speed, and
maintenance.
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1. Introduction
Clean water is a vital resource for anthropogenic activity. As time progresses and
population increases, humans utilize more water, resulting in increasing contamination. A
significant portion of water used in households is discharged via sewer systems to wastewater
treatment plants (WWTPs) (Cunningham, Cunningham, & Saigo, 2007). Wastewater is typically
treated to remove solids, organics, and nutrients, before the effluent is released back into the
environment (USEPA, 1998; Jones et al., 2007). Although wastewater treatment systems remove
or neutralize some of the organic matter in wastewater, some pollutants, such as androgens,
detergents, and estrogens are not degraded in this manner due to their chemical stability
(USEPA, 1998; Caliman & Gavrilescu, 2009). Water quality of WWTP effluent varies,
depending upon local treatment requirements as well as effectiveness of individual treatment
methods (Schröder et al., 2007).
Of particular concern are Endocrine Disrupting Chemicals (EDCs), compounds that
disrupt an organism’s ability to bind, eliminate, metabolize, secrete, or synthesize hormones
important in the developmental, homeostatic, and reproductive systems. Effectiveness of a
particular EDC is a function of dose, age of the individual at exposure, latency, and synergistic
effects of the particular chemical. EDCs are highly bioactive because they mimic physiological
metabolites, thus there are no known immunities (Diamanti-Kandarakis et al., 2009). EDC-
related problems can include genetic damage or legacy pollutant exposure, for example, males
being born with cryptorchidism or hypospadias (Diamanti-Kandarakis et al., 2009; Bhatia et al.,
2005).
Traditional toxicological assumptions do not hold for EDCs. One of these assumptions is
when concentration increases, damage to health increases. However, certain EDCs have toxic
effects at relatively high or only high and low concentrations so that their dose-response curves
are J or U shaped. Another toxicity generalization is that toxins cannot have opposite effects, but
it can have multiple ones. EDCs can have opposite effects, for example, in dealing with different
groups in a population. The final assumption is that as the amount of a chemical approaches zero,
the effects of the chemical will become basically non-existent. Even minute amounts of EDCs
can cause significant health effects (Caliman & Gavrilescu, 2009).
EDCs in the environment affect aquatic wildlife, especially fish such as Cyprinus carpio,
which have been known to take up EDCs readily from their environment (Petrovic et al., 2002).
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Once in the individual, EDCs can cause problems by disrupting steroids, e.g. mimicking estrogen
or changing estrogen receptor status (Goksøyr, 2006; Roberts et al., 2010). Additional research
would be beneficial to determine the full extent to which this group of chemicals affects wildlife,
especially when more than one of these chemicals is present.
Caliman & Gavrilescu (2009) propose the following criteria to determine whether
medicines would cause environmental problems: the drugs in the environment produce active
EDCs or compounds that may become active; the drugs have high enough concentrations to
cause harm to endocrine systems; and active EDCs are portable and not easily degraded once in
the environment. For the purpose of this paper this criteria will be used.
A significant source of EDCs in the environment comes from wastewater effluent
(Jackson & Sutton, 2008; Jones et al., 2007; Oppenheimer et al., 2007). Household and business
EDCs entering wastewater treatment plants include drugs and drug additives, cleaning agents,
fire-retardants, personal care products and natural hormones (Caliman & Gavrilescu, 2009;
Jackson & Sutton, 2008). A Southern California phone survey by Kotchen et al. (2009)
interviewed 1005 residents and revealed 28.0% of all respondents disposed of unwanted
pharmaceuticals in the toilet or sink. 23.2% respondents who were aware of the problem of drugs
in surface water and treated wastewater still chose to dump their old medications into the toilet or
sink. This is compared to 31.3% that were unaware of the problem and used the same disposal
methods. The small difference suggests that it is unlikely that the problem can be fixed with
education. Thus, it would be beneficial to implement steps in wastewater treatment that reduce
EDCs in wastewater, rather than rely on changing consumer habits. Furthermore, excretion of
pharmaceuticals is the main method of WW contamination, and this is will not likely change
(Kotchen et al., 2009).
Pharmaceuticals that have endocrine disrupting properties have been used for many years
to ensure human and animal health. Continued use of these drugs is important in individual and
economic health. Thus, it is improbable that negative effects of these chemicals can be lessened
through reduction in use (Jones et al., 2007). EDCs can be neutralized during wastewater
treatment processes by multiple technologies including bioremediation or “advanced post-
treatment” (Caliman & Gavrilescu, 2009).
Bioremediation uses fungi, plants, bacteria, or other microbes such as plankton to break
down EDCs (Hai et al., 2006; Schöder et al., 2007; Liao et al., 2010; Ishihara & Nakajima,
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2003). Other technologies, termed “advanced post-treatment” include adsorption, advanced
oxidation, and membrane processes (Caliman & Gavrilescu, 2009).
This project focuses on bioremediation of WWTP effluent using mycoremediation and
phytoremediation. Mycoremediation is a type of Mycorestoration. Mycorestoration is the process
of using fungi to help solve environmental problems in one of four ways. Mycofiltration uses
fungi to filter organisms, pollutants, and silt from water. Mycoforestry is the process of utilizing
fungi to restore or sustain forest health. Mycoremediation utilizes fungi to degrade or remove
toxins from the environment. Mycopesticides uses fungi to combat pests, mostly insects
(Stamets, 2005). Mycoremediation shows promise in its efficiency and economy. Fungi are
decomposers that break down waste such as wood, so they are naturally adept at breaking down
organic matter. Unlike plants, fungi do not require sunlight to grow. Finally, estimated costs of
mycoremediation are relatively low compared to other methods of remediation, biological and
otherwise (Stamets, 2005). Mycoremediation can be effective in wastewater treatment for the
removal of EDCs. The fungi are able to break down the chemicals by either extracellular
oxidation or intracellular initial attack (Harms et al., 2011).
Phytoremediation is the process of using plants to filter out unwanted chemicals in soil or
water, and is effective for removal of pesticides, polycyclic aromatic hydrocarbons (PAHs),
landfill leachates, solvents, crude oil, explosives, and metals (Kadlec & Wallace, 2009). Plants
do this by absorbing the toxins or remediating them in the soil or water. Plants are also
considered a low cost method, although a time consuming one (Kadlec & Wallace, 2009;
Chaudhry et al., 2002).
This paper compares fungi and plant based bioremediation in order to determine their
feasibility for elimination of EDCs in San Francisco Bay Area (SFBA) wastewater. Bacterial and
other remediation are outside the scope of this thesis.
2. Background
This section discusses EDCs in the San Francisco bay, EDC transfer through wastewater
treatment, and different types of remediation methods.
2.1 EDCs in the San Francisco Bay
The effects of EDCs, including phenolic ones, are understudied globally. The San
Francisco Bay Area is no exception (Thompson et al., 2007; Brooks et al., 2011). Studies that
look at EDCs in the SFBA often look at phthalates. Hwang et al. (2006) found that Stege Marsh
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in SFBA had the most phthalates out of all tidal salt marsh sediments tested in the study, which
included Carpinteria Marsh, China Camp, Tom’s Point and Walker Creek. Stege Marsh is in the
east San Francisco Bay (SFB), China Camp is in the north SFB. Tom’s Point and Walker Creek
are located along the coast about 45 km north of SFB. Carpinteria Marsh is near the city of
Carpinteria, which is 15 miles south of Santa Barbara. Locations of phthalates may correlate to
where phenolic compounds are found.
A pilot study by Jackson and Sutton (2008) identified sources of EDCs flowing into East
Bay Municipal Utilities District’s (EBMUD) WWTP in Oakland. Facilities tested included a nail
salon, pet wash, residential coin laundry, diaper service laundry, two industrial laundries,
hospital, veterinary clinic, medical clinic, adhesives manufacturer, beverage manufacturer, paper
products manufacturer, pharmaceutical manufacturer, plastic bag manufacturer, two samples in a
residential area, two samples of wastewater treatment (WWT) influent, and three samples taken
after WWT. The five phthalates tested were Butylbenzyl phthalate (BBzP), Di-n-butyl phthalate
(DBP), Di-2-ethylhexyl phthalate (DEHP), Diethyl phthalate (DEP), and Di-n-octyl phthalate
(DOP). Four other known and suspected EDCs were also tested. They were Bisphenol A (BPA),
4-Nonylphenol (NP), Triclosan (TCS), and Tris(2-chloroethyl) phosphate (TCEP). Because of
the information available on this WWTP, the recommendations will be applied to it.
Figure 1 is based on the results of the study. It does not contain NP nor does it include
results from the beverage manufacturer because all of the tests done for these were below
detection levels. BBzP, BPA, and DEHP were found in one of the two field blanks.
Contamination of these chemicals in commercial laboratories is prevalent. It is also possible that
contamination occurred during sampling. BBzP and DEHP concentrations in some samples were
below two times the values of the field blank. DBP, DEHP, and TCEP had one or more
concentrations that were estimated because methods used to measure them did not have the
proper range, or there was interference due to water matrices (Jackson & Sutton, 2008).
The EDCs found in wastewater of the various facilities were overall consistent with
expectations based on products expected to be used there. For example, the greatest
concentration of DOP, used in producing plastics, was in the wastewater of the plastic
manufacturer. One exception to the norm was TCS. Even though it is in many household
products, it was not found in residential areas. One possible reason for no detection is that the
houses tested did not use triclosan-containing products. However, Jackson and Sutton (2008)
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state that they believe it is more likely that the testing methods did not have a low enough range.
NP was not detected in any facility effluent. NPs are alkylphenol ethoxylates (AE) metabolites. It
is possible that AE did not have enough time to degrade to NP for testing. Facilities may also not
use products with AE (Jackson & Sutton, 2008).
Residen
tial 1
Residen
tial 2
Nail sal
on
Industrial
laundry
1
Industrial
laundry
2
Residen
tial coin lau
ndry
Diaper
servic
e
Pet wash
Veterin
ary cli
nic
Hospital
Medica
l clinic
Pharmace
utical m
anufac
turer
Plastic b
ag man
ufacturer
Paper
products
man
ufacturer
Adhesive
s man
ufacturer
Pre-tre
atmen
t influen
t 1
Pre-tre
atmen
t influen
t 2
Treate
d wastew
ater 1
Treate
d wastew
ater 2
Treate
d wastew
ater 3
0.1
1
10
100
1000
10000
Concentrations of EDCs in wastewater samples
DEPDBPBBzPDEHPDOPBPATriclosanTCEP
Facilities
Conc
entr
ation
s of E
DCs (
μg/L
)
Figure 1: Concentrations of EDCs in WW samples before and after water passed through the
EBMUD WWTP (Based on Jackson & Sutton, 2008).
Jackson and Sutton (2008) found that all medical facilities had TCS in their WW effluent.
However, TCS was also found in other effluent, so addressing the contamination in just medical
facilities would not be enough to reduce it in WW. They also noted that laundry facilities had a
wide range of EDCs in their WW, with DEHP being the only one found in all four.
Jackson and Sutton (2008) recommend public outreach to encourage producers of EDCs
to stop using products that contain them. They state that prevention may be cheaper and more
effective than treatment at the WWTP level. This is in conflict with consumer reliability as found
at Kotchen et al. (2009).
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2.2 EDC transportation through water treatment plants
Loraine et al. (2006) measured pharmaceuticals and personal care products (PPCPs) in
raw and treated drinking water of San Diego County. Benzophenone, butyl benzyl phthalate,
butylated hydroxyanisole, clofibrate, clofibric acid, di(ethylhexyl) phtahalate, diethyltoluamine,
dibutyl phthalate, diethyl phthalate, dimethyl phthalate, hydrocinnamic acid, ibuprofen,
ibuprofen methyl ester, octyl methoxy cinnamate, surfynol, and triclosan were found in raw
drinking water. All of the PPCPs except clofibrate, clofibric acid, and diethyltoluamine were
found in treated water. At the wastewater reclamation plant where samples were taken
biochemical oxygen demand (BOD), BOD removal efficiencies, pH, and temperature did not
change significantly during the different seasons.
It was found that some PPCPs presence and amount varied seasonally. 77% of the
PPCPs were found in higher concentrations in the summer both in WWTP influent and reclaimed
wastewater (WW); 44% were only found during summer. The authors hypothesized that some of
this is due to the increase of sunscreen and pesticide use during summer months. It was also
noted that pharmaceuticals and phthalate esters also increased in concentration at this time. The
increase in PPCPs was found in drinking water before it became WW. They predict this rise is
due to a smaller volume of water in waterways to dilute pollutants. The highest concentrations
were found during the dry season. This is when imported water use increases. It was found that
raw imported water organic pollutant concentrations during summer were almost as high as those
of nonpotable reclaimed WW (Loraine et al., 2006).
Other studies outside SFBA have examined BPA, NP, and TCS remediation through
WWTPs. Giger et al. (2009) did a study analyzing phenolic EDCs, including BPA and different
forms of NP, in WWTP influent and effluent. They also looked at concentrations in the River
Glatt, near Zurich, Switzerland, which is where WWTP effluent was discharged. BPA median
influent concentration was 414 ng/L and its median effluent concentration was 24 ng/L. In the
Glatt River BPA median was 9.4 ng/L. NP median influent value was 473 ng/L and the median
effluent concentration was 123ng/L. NP concentration in Glatt River was 64 ng/L.
During this study two large rain events occurred. During this time WWTP influx had to
be dumped into River Glatt because the WWTPs were above capacity. Some EDCs had an
increase in concentration in river water during this time. This suggests that EDCs do not
necessarily dilute with greater flow (Giger et al., 2009). This result deviates from other studies
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which found that greater flow did dilute pollutants (Loraine & Pettigrove, 2006). Giger et al.
(2009) calculated expected concentrations of EDCs to compare to real world samples.
Calculations suggest that during days of low flow all EDC sources were known. However,
during high flow events authors concluded not all EDC sources were accounted for. This could
account for the odd result of similar environmental concentrations with greater flow. NP
degradation in Glatt River was insignificant, although BPA did decrease (Giger et al., 2009).
It was found that River Glatt had a decrease in alkylphenolic compounds by more than
one order of magnitude since 1984. This was suggested to be the product of reducing alkylphenol
polyethoxylates as well as improvements in wastewater treatment technology. EDCs are partially
degraded in WWTP activated sludge treatment. This can result in the presence of metabolites in
effluents as well as in receiving water bodies. Giger et al. (2009) found that it is necessary to
conduct field studies as modeling can be inaccurate when trying to calculate WWTP EDC
removal.
2.3 How mycoremediation removes toxins
Fungi are adept at remediation of toxins because they are decomposers. Fungi are able to
break down chemicals by either extracellular oxidation or intracellular initial attack. Fungi have
specialized enzymes for decomposition. These enzymes are not substrate specific. Extracellular
oxidation occurs when fungi produce oxidase enzymes which are used outside of the fungal cell
walls to break down organic molecules. A commonly studied enzyme that contains copper is
called laccase. Peroxidases are also enzymes that have the capacity to break down EDCs. These
lignin modifying enzymes (LMEs) may be more useful in breaking down the organic molecules
than those produced by bacteria because they work outside the organism, which make them more
effective on chemicals that have small concentrations, low bioavailability, or unique structural
elements (Harms et al., 2011). White rot fungi (WRF) and LMEs have shown promise as
treatment because they have successfully degraded a variety of organic pollutants in various
environments (Cabana et al., 2007(1)).
WRF is a large and varied classification. It contains basidiomycetes and decomposing
fungi that are capable of breaking down lignin aerobically. WRF put out one or multiple LMEs,
which are extracellular, as well as mediators with low molecular weights. The latter increases the
types of chemicals WRFs can oxidize. LMEs are produced when fungi are in low nutrient
substrates. Because WRF are filamentous, they can reach environmental toxins that bacteria may
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fail to contact. The key LMEs fungi produce are laccases (Lac), lignin peroxidases (LiP),
manganese-dependent peroxidases (MnP), and versatile peroxidases (VP). WRF species make
different forms and amounts of LMEs. They also vary in different growth conditions (Cabana et
al., 2007(1)).
Intracellular enzymes can also be effective in remediating toxins and also are not specific
to the substrate they grow on or in. Toxins pass through the mycelium cell wall and are then
attacked by enzymes such as cytochrome P450, which can break down a range of chemicals,
such as dioxins and PAHs. Intracellular activity can motivate extracellular activity such as
Fenton reactions (Harms et al., 2011).
The application types of mycoremediation are: direct interaction between fungi and
contaminant, use of fungal enzymes in free form, and use of fungal enzymes in immobilized
form. Direct interaction can be done through a solid medium like sediment, or in an aqueous
solution. Free enzymes can be used in a solid medium, aqueous solution, or organic solvents.
Immobilized enzymes are used in aqueous and organic solutions (Cabana et al., 2007 (1)).
Mostly true rules about mycoremediation (Cabana et al., 2007(1)):
Lignin oxidation gives no net energy to fungi, so LMEs are excreted during a second
metabolism
LMEs are produced when carbon, nitrogen or other nutrients are restricted
LiP and MnP have high production with oxygen partial pressures that are high
LiP and MnP have low production when in a submerged WRF culture that is agitated
Lac has high production when in submerged WRF culture that is agitated
WRF can make multiple types of isoforms of LMEs with varying growth conditions and
species strains
LME degradation of EDCs can be influenced by additives, heavy metals, inorganic salts,
organic chemicals, pH, temperature, and other wastewater contaminants. These variables will
manipulate enzyme stability, activity and substrate specificity.
2.4 How phytoremediation removes toxins
Phytoremediation has also shown promise as a viable technology. As with
mycoremediation, there are multiple mechanisms possible. Contaminants may be absorbed by
plants, and once absorbed, plant material and contaminants may be physically removed from the
site. Conversely, contaminants may be altered when plants release chemicals through roots or
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microbes at the roots that break down contaminants. In this second mechanism, vegetation
obtains trace chemicals the way it obtains nutrients, through the roots themselves (Kadlec &
Wallace, 2009). Plants can be used indirectly as a support of microorganisms. Plants can also
volatilize PPCPs or their metabolites (Schröder et al., 2007). Vegetation is grown in situ, either
in sediment or in water and is replanted until the contaminant is reduced to acceptable levels.
However, one drawback of phytoremediation is the possibility of a new route of exposure to the
toxin, as animals may consume plants used in remediation. Another drawback is time;
phytoremediation can take years to clean a site (Kadlec & Wallace, 2009; Schröder et al., 2007;
Imai et al., 2007). Phytoremediation is publicly popular and low cost compared to
physiochemical or bacterial remediation. However, environmental conditions will influence plant
capabilities of targeting undesired chemicals (Imai et al., 2007).
There are multiple types of phytoremediation applications: cell culture, constructed
wetland, phyto-degradation, phyto-exraction, phyto-volatiliztion, and rhizosphere degradation.
(Schröder et al., 2007; Shimoda et al., 2009). One application of phytoremediation is plant cell
culture, which can be done in a free or immobilized form. The ways chemicals are altered in this
method varies significantly and consists of esterification, glycosylation, hydrolysis,
hydroxylation, isomerization, methylation, oxidation, and reduction (Shimoda et al., 2009).
Constructed wetlands (CWs) are fully or partially man-made wetlands built to clean
water. CWs can remove nutrients, organic matter, and pathogens. They are more capable of
shock loads than submerged biomass treatment systems. High efficiency along with low
maintenance and low operating energy and costs make CWs an attractive technology. Although
this technology has some problems, increased use will help operators perfect its application.
Overall, CWs are good at removing PPCP and EDCs, but the break down can vary based on
retention time and rhizosphere chemistry (Schröder et al., 2007).
Vertical flow beds (VFBs) are CWs with a single or multiple vertical flow or reverse
vertical flow chamber(s). Their vertical orientation increases root surface and pollutant contact,
and increase diversity of the rhizosphere which allow for different bacteria to be present. In both
VFBs and CWs plants that increase oxygen to the rhizosphere can be used for indirect
remediation assistance (Schröder et al., 2007).
Hydroponic systems (HSs) grow plants without soil. This is an attractive method because
pollutants might react with the soil in an undesired manner, for example, pollutants with a high
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Kow will have a greater affinity for soil. Another motivation for hydroponic growing is a system
designed so that as the plant absorbs toxins, it can be easily removed after optimal absorption has
occurred. Hole filled plastic or rubber material can be the growing medium in HSs. This
technique is more reliant on the plant because there is not a rhizosphere where reactions can take
place. Helophytes are good candidates for this system. Some aeration may help increase growth
rates (Schröder et al., 2007).
A mix of these techniques would be beneficial as more than one species of plants make
the system more effective in the quantity of PPCPs degraded, and make the system more
resistant to stress (Schröder et al., 2007).
Positives of phytoremediation include (Schröder et al., 2007):
They can degrade a range of organic toxins to safe chemicals
The do not add unwanted compounds (like chemical treatment might)
They are moderately simple to maintain, not requiring much education
They can be changed to fit local needs relatively simply
They do not have high upfront costs
They can be effective in a variety of scales
They can be installed in modules which increases reliability
Phytoremediation can have negative aspects. One negative is roots tend to absorb and
hold more hydrophobic compounds and foliage prefers hydrophilic ones, so plants will not
absorb toxins equally. This technique is likely to be less effective where there are very low
concentrations of contamination. The efficiency of the system is dependent on climate, soil and
growth medium factors, water quality and nutrients. The rates at which toxins are degraded or
absorbed are limited, even with optimal plant growing conditions (Chaudhry et al., 2002).
How well plants uptake compounds is dependent on a variety of factors including
application technique, chemical properties, local climate, soil characteristics, and species of
plant. Pollutants can travel to different parts of the plant than the location where the pollutant
entered, generally the roots, by the xylem. Transportation is more common for chemicals that are
a little hydrophobic with a log Kow of about 1.8. Pollutants that are more hydrophobic, especially
with a log Kow >3, tend to stay in the roots bound to lipids (Chaundry et al., 2002). BPA, NP, and
TCS have log Kows of 3.40 (measured) – 3.82 (calculated), 4.48, and 4.76 respectively (Staples et
al., 1998; Ahel & Giger, 1993; Park et al., 2009). Hydrophilic chemicals can have a low
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absorption through the waxes of leaf cuticles (Chaudhry et al., 2002). Ideally, phytoremediation
in the form of extraction has a high amount of pollutants per mass of the plant (Zhang et al.,
2008).
Mineralization of toxins is the ideal outcome in phytoremediation. However, some
processes can change the chemical composition enough so that the toxin becomes
environmentally benign. These processes include degradation, synthetic transformation, and
rearrangement (Chaudhry et al., 2002).
The efficiency of phytoremediation can be dependent on the technology applied. One
way to increase efficiency is the use of genetic modification. Public opinion on the subject may
be more positive than genetically modified food because of the improved pollution cleaning
ability. Another method for increasing organic contaminant absorption is by augmenting root
growth in the plants. It is suggested that this is caused by a boost in root bulk and cell division
(Chaudhry et al., 2002).
2.5 How other remediation techniques remove toxins
Along with myco- and phytoremediation there are other biological and physiochemical
techniques. The methods are summarized in the following section.
2.5.1 Bacterial bioremediation
Bacterial bioremediation is the process of using bacteria to transform pollutants. Bacteria
do this as a survival tactic. Immobilization, mineralization, and transformation are mechanisms
used by bacteria to eliminate toxins. Bacteria are able to respond quickly to change due to their
high growth rates, ability of horizontal gene transfer, and metabolic versatility. How well
bacteria degrades a pollutant, such as petroleum, is dependent on moisture, nitrogen, oxygen, pH,
phosphorus, and temperature. Different bacteria are capable of degrading in aerobic or anaerobic
environments. This makes bacteria versatile (Sinha et al., 2009).
Bacterial bioremediation is commonly used in WWTPs as activated sludge. Jones et al.
(2005) suggests that instead of adding another type of treatment, sewage treatment plants (STPs)
could increase their sludge retention times and use both nitrification and denitrification. This
allows for more bacterial remediation. Methanotrophic bacteria and nitrifying bacteria both
produce monooxygenase enzymes which may help the degradation of EDCs. Furthermore, using
both nitrification and denitrification methods exposes waste to a greater range of bacteria which
use different enzymes. A greater number of enzymes are more likely to break down a greater
11
number of EDCs. By increasing sludge retention time bacteria that grow slowly are able to
participate in the remediation process. It is important to note that neither of these treatments
reduced all of the EDCs in the 2005 study.
2.5.2 Plankton (phyto- and zoo-)
Plankton have been known to reduce EDC levels in the environment (Caliman &
Gavrilescu, 2009; Ishihara & Nakajima, 2003). A lab experiment by Ishihara and Nakajima
(2003) introduced substrates containing one of four EDCs into an “eco-system” to see how the
EDC was decomposed and recovered by phytoplankton and zooplankton. The four EDCs were
bis(2-ethylhexyl)phthalate, bisphenol A, p-tert-octylphenol, and 4-n-octyloxyphenol.
Phytoplankton used were Chaetoceros gracilis and Nannochloropsis sp. Zooplankton used were
Artemia sp., Brachinous sp.
The first set of tests measured reactions of various EDCs using Nannochloropsis sp. and
C. gracilis in a test tube. Bis(2-ethylhexyl)phthalate and BPA eco-systems showed a reduction in
used saltwater toxin concentrations and an increase in phytoplankton toxin concentrations. This
suggests that EDCs accumulate in the phytoplankton. In p-tert-octylphenol and 4-n-
octyloxyphenol ecosystems both saltwater and phytoplankton toxin concentrations were low,
which suggests that the EDCs break down after they are consumed by the phytoplankton
(Ishihara & Nakajima, 2003).
The second set of tests measured different combinations of phytoplankton and
zooplankton concentrations using BPA. The eco-system consisted of an acrylic pipe with nylon
mesh at the end submerged in the synthetic saltwater medium, with a substrate solution
containing the EDCs added. Phytoplankton were in both the acrylic pipe and the salt water,
because they could pass through the nylon mesh. Zooplankton were only in the acrylic pipe
because they could not fit through the nylon mesh. The whole system was gently stirred. EDCs
were consumed by phytoplankton, which were consumed by zooplankton. Alone phytoplankton
took up 46% of the EDC. When both types of plankton were used together zooplankton took up a
little over 40% and phytoplankton accumulated 9 or 10%. The zooplankton species used alone
only accumulated 5 or 6% (Ishihara & Nakajima, 2003).
2.5.3 Chemical and Physical
There are a variety of chemical and physical means of removing EDCs from wastewater.
Advanced treatment such as granular activated carbon (GAC) and ozone (O3) are problematic
12
due to their energy requirements; moreover, a treatment plant powered by conventional energy
creates atmospheric carbon dioxide. By trying to reduce organic water pollution, these processes
may be contributing to global climate change, making them unsustainable. They may become
sustainable if renewable energy were to be used. There are also other construction and operation
environmental costs. Furthermore, these systems can have increased sludge production when
water purity increases. This cost is significant, but often not included during cost/benefit analysis
(Jones et al., 2005).
2.5.3.1 Adsorption
One method to remove EDCs from effluent is adsorption. Adsorption may be
accomplished using activated carbon (AC), either GAC, powdered activated carbon (PAC), or
other materials such as Mg-Al layered double hydroxide (LDH). AC is good in removing
estrogen, steroids and other hydrophobic chemicals. However, it is not as adept at removing
other chemicals such as ibuprofen and diclofenac. AC is available in many forms, and some
forms are more absorbent and less expensive than others. Humic material (HM) can make AC
units less efficient, so it is recommended that HM is removed before going through this process.
LDHs are also useful in removing estrogen mimickers. They can be used in slurries or in packed
columns (Caliman & Gavrilescu, 2009).
A study done by Gong et al. (2009) examined a new way to remove BPA from water by
sorbing it to hemimicelles. Esterified carboxyl cotton (ECC) was used as a sorbent and cetyl
trimethyl ammonium bromide (CTAB) was used as a surfactant. The maximum amount of BPA
could be extracted from the water with a pH between 4 and 10. Equilibrium took about 4 hours.
The thermodynamics of the method appeared to be exothermic and spontaneous. Other
substances have shown to be sorbents for BPA: calix[4]crown derivative, carbon nonomaterial,
carbonaceous material, chitosan-bearing β-cyclodextrin, Fe(III)/Cr(III) hydroxide, hydrophobic
zeolite, mineral, molecularly imprinted polymer, polyetherssulfone-organophilic
montmorillonite, and polysulfone bead (Gong et al., 2009).
2.5.3.2 Fenton Reagent
The Fenton process uses iron ions combined with hydrogen peroxide (H2O2) to make
hydroxyl radicals (˙OH). The radical is then used to destroy EDCs. The effectiveness of this
process is dependent on pH of the water as well as the ratio of oxidant to Fe. Adding ultraviolet
13
light (UV) to the process increases the amount of ˙OH, which makes the system more efficient
(Caliman & Gavrilescu, 2009).
2.5.3.3 Hydrogen Peroxide
H2O2 can be used for remediation alone, but is mostly used in the Fenton process and
ozone (see sections 2.5.3.2 and 2.5.3.5) (Caliman & Gavrilescu, 2009).
2.5.3.4 Membrane processes
Membrane filtration is a method whereby water is passed through a membrane that only
lets chemicals with certain properties through it. Permeate is the liquid that passes through the
membrane, and the retentate is the materials that are blocked. The main types of membrane
filtration are, in order of low to high efficiency, microfiltration, ultrafiltration, nanofiltration, and
reverse osmosis. The removal of EDCs depends on chemical, membrane, and water properties.
Divalent cations in the water can reduce the EDCs and drugs that are removed from the water
because the ions stop the pollutants from binding with organic material. Even though
nanofiltration and reverse osmosis are effective, EDCs can contaminate permeate with pH
variation and membrane backwashing. Thus, it is recommended to use membrane filtration as a
secondary treatment. These systems have a high energy use (Snyder et al., 2007). It is possible to
improve EDC removal by membranes with the use of chemical treatment (Caliman &
Gavrilescu, 2009).
2.5.3.5 Ozone
Ozone is a strong oxidant that can directly break down PPCPs. It can also be used in
combination with H2O2, UV, or both to produce ˙OH. Combinations tend to be the most efficient
oxidizers. ˙OH is capable of oxidizing multiple chemicals in a short amount of time without
preference. These processes work best in waters with low dissolved organic carbon (DOC). O3
oxidizes many drugs, but it favors activated aromatic systems and non-protonated amines. It
oxidizes most EDCs in less than 100 seconds (Caliman & Gavrilescu, 2009).
2.5.3.6 Photocatalysis
Photocatalysis shows promise in remediating EDCs in WW, especially with titanium
dioxide (TiO2). In this process semiconducting materials are illuminated and go into excited high
energy states that move to particle surfaces and are capable of many redox reactions. Studies
show that this treatment used with TiO2 and UV is useful in removing unwanted organics from
environmental systems (Caliman & Gavrilescu, 2009).
14
2.5.3.7 Sonolysis
Sonolysis is the process of using ultrasound to form ˙OH to break down chemicals. The
ultrasound builds pressure in the water that makes and collapses bubbles which makes high
temperature and pressure. This causes dioxygen and water to make hydrogen radical (H˙) and
˙OH. The radicals then cleave organic pollutants. It is used best in acidic water (Caliman &
Gavrilescu, 2009).
2.5.3.8 Ultraviolet light
H2O2 and UV, alone or together, can also break down pollutants non-selectively and
rapidly. UV is also commonly used Fenton processes or ozone (see sections 3.5.3.2 and 3.5.3.5)
(Caliman & Gavrilescu, 2009).
3. Methods
To determine the best remediation method for EDC removal from a SFBA WWTP
effluent a literature review was conducted. Remediation methods were compared and an optimal
technique was identified.
4. Results and Discussion
A literature analysis was done examining myco- and phytoremediation methods of BPA,
NP, and TCS. The findings are summarized in this section.
4.1 Which EDCs remediation will focus on
EDC types include phenols, phthalates, and natural and synthetic hormone substances
(Cabana et al., 2007(1)). BPA, NP, and TCS are phenolic endocrine disruptors. BPA and TCS
have been found in SFBA WWTP effluents (Jackson & Sutton, 2008). Detectable amounts of NP
have been found in SFBA waters, oysters, and mussels (Hoenicke et al., 2007). Because of their
presence in SFBA and their similar chemical nature these three EDCs were chosen to be the
focus of this paper’s remediation analysis.
4.1.1 Bisphenol A (BPA) (2,2-bis(4-hydroxyphenol)propane)
BPA is a commonly used industrial chemical. One application of BPA is manufacturing
processes of polycarbonate plastics which are found in items such as food and drink packaging.
It is also found in epoxy resins which are often used as anti-corrosion liners for food and drink
metal packaging. BPA is commonly found in polyvinyl chlorides and also in electronic devices
(USEPA, 2013; Loffredo et al., 2012; Imai et al., 2007). In 1936 BPA was identified as an
artificial estrogen (NTP, 2008). It was used as a growth hormone for poultry and cattle. For a
15
short time in the 1930s it was used as an estrogen replacement for humans. However,
diethylstilbestrol succeeded it quickly (Kwon et al., 2007). BPA is hard to remove from
wastewater by commonly used physiochemical means. This supports the practicality of using
biological alternatives (Imai et al., 2007). Cabana et al. (2007 (1)) reviewed several studies that
measured BPA concentration in effluent. They reported 2.2 µg/L is the highest concentration of
BPA found in sewage effluents.
BPA can have significant environmental effects. While most environmental
concentrations are not high enough to produce teratogen effects, BPA can still disrupt endocrine
systems. Modification of the gonadal function, sex determination, and commencement of liver
vitellogenin production can all be caused by environmental BPA exposure. When lesser amounts
of BPA are present, the gonadal functions of vertebrates can be damaged. For example, fish
testicular systems can be altered. Sperm count can be lowered, which can affect reproduction
timing in fish. Female fish can also have their gonadal processes affected. BPA can delay or
completely stop ovulation in female trout. Insects can also have their gonadal systems damaged.
When animals with temperature-dependent sex determination, such as reptiles and amphibians,
are exposed to high amounts of BPA they can have the sex of their offspring affected, mostly
increasing the number of females. Vitellogenin is a protein that is precursory to egg yolk
proteins. It is not generally produced in males. However, with high levels of BPA, or other
estrogen mimickers, it can be found in amphibian, fish, or insect males (Crain et al., 2007).
Xiphophorus helleri (swordtail) have been shown to have cell death in their testes when exposed
to BPA (Kwak et al., 2001). BPA has also been shown to kill off Oryzia latipes (medaka) eggs
(Shioda and Wakabayashi, 2000) and hinder aquatic midge growth (Hahn et al., 2002; Watts et
al., 2003; Lee et al., 2006).
BPA is an EDC that was used in hormone therapy (Kwon et al., 2007). The National
Toxicology Program has “some concern” that BPA can have negative effects in fetuses, children,
and infants regarding their behavior, brain, and prostate gland. They have “minimal concern”
about exposure for fetuses, children, and infants having problems in mammary glands and
reaching puberty at an early age. They also have “minimal concern” for employees who interact
with a significant amount in their job. The National Toxicology Program has “negligible
concern” about teratogenic effects, as well as adults who do not have BPA in the workplace
16
(NTP, 2008). It should be noted that environmental exposure is not considered a risk to humans
(USEPA, 2013; NTP, 2008; Loffredo et al., 2012; Imai et al., 2007).
4.1.2 4-n-nonylphenol (NP)
There are multiple types of NP, including 4-n-nonylphenol and p353 nonylphenol. For
this study it is assumed that all NPs have the same properties and react the same way to
treatment. NP is used to color fuel oil and to make oxime, which isolates copper. It is common
for NP to be an intermediate during manufacture of other compounds (USEPA, 2005). A greater
concern than direct dumping of NP and other alkylphenols into wastewater is their degradation
from alkylphenol ethyloxylates (AEs) (USEPA, 2005; Jackson & Sutton, 2008). AEs are
common chemicals used in detergents, paints, PCPs, and pesticides (Jackson & Sutton, 2008). A
subgroup of AEs is nonylphenol ethoxylates (NPEs), which are used industrially as surfactants.
In addition they are in households as antistatic compounds, detergents, and solubiliers (EPA,
2010; Soares et al., 2008). STP effluents are the main source of environmental NP (Soares et al.,
2008; Petrovic et al., 2002). Naylor et al. (1992) found that NPEs were degraded more in US
STPs than European ones. NP is found in relatively large concentrations in US STPs and rivers
(Pryor et al., 2002).
It was found that in fish NP is metabolized quickly, with a half-life of 24 to 48 hours of
digestion or waterborne exposure (USEPA, 2005).It is suspected that aquatic species moderately
bioaccumulate NP (USEPA, 2005). Oryzias latipes and Xiphophorus maculates are two fish
species that sustained male reproductive harm when exposed to NP (Tabata et al., 2001;
Kinnberg et al., 2000). However, in Oncorhynchus mykiss NP commenced the assembly of
female proteins (Jobling et al., 1996). Overall, NP is very acutely and chronically toxic to aquatic
invertebrates, aquatic plants, and fish (EPA, 2010). The maximum concentration of NP found in
surface water was 6.86 µg/L. NP readily adsorbs to sediment and accumulates in organisms
because it is highly hydrophobic (Cabana et al., 2007 (1)). It has been found that concentrations
of 10 μg/L of NP can cause damage to the endocrine systems of aquatic organisms (Jobling et al.,
1996). The EPA recommends that a water’s 4-day concentration average not be higher than 6.6
μg/L (USEPA, 2005).
NP in drinking water is not considered to be a significant means of exposure compared to
other sources such as detergents, PCPs and food packaging (CEPA, 2000; Guenther et al., 2002).
The EPA gives NP a low acute dermal and oral toxicity. It is likely to corrode and irritate eyes
17
and skin. However, it is unlikely to be a skin sensitizer (EPA, 2010). It was found that cell
proliferation in breast tumors could be commenced by NP (Soto et al., 1991). NP can compete
with 17β-oestradiol for binding sites by mimicking it (White et al., 1994). Lee et al. (2003) found
that NP can disrupt male reproductive systems in humans.
4.1.3 Triclosan (TCS) (5-chloro-2(2,4-dichlorophenoxy)phenol)
TCS is compound that is commonly used as a preservative, and for antibacterial and
antifungal means. It has low water solubility and is likely to adsorb to soil and bioaccumulate.
Commercial and industrial uses include conveyor belts, dye bath vats, equipment that makes ice,
and fire hoses. TCS is a material preservative that is used in a variety of products including:
adhesives, clothing, caulking compounds, garbage cans, mattresses, plastics, rubber, shower
curtains, and toilet bowls. As a bacteriostat and fungicide it is used in products such as anti-
bacterial soaps, deodorants, and toothpaste (EPA, 2008).
TCS has been found to be slightly toxic to birds, highly toxic to freshwater fish, and
highly toxic to freshwater invertebrates. There is not enough data to know the toxicity of TCS to
terrestrial and aquatic plants (EPA, 2008 (2)). Veldhoen et al. (2006) found that Rana
catesbeiana (frog) tadpoles exposed to environmental concentrations of TCS had affected
development. Effects included decreases in body weight and T3-mediated TRβ mRNA
expression. Increases in hindlimb development and levels of brain PCNA were other symptoms.
Furthermore, TCS disrupted cerebral α transcript concentrations. TCS increases bacteria
mortality at environmental concentrations of 0.21 μg/L. It also reduces the effectiveness of
photosynthesis in algae (Ricart et al., 2010). Adolfsson-Erici et al. (2002) found TCS in fish bile
living near WWTPs. 4.1 µg/L is the highest concentration of TCS found in effluents. It is also
highly hydrophobic and tends to adsorb to sediment and bioaccumulate in organisms (Cabana et
al., 2007 (1)).
The EPA has given TCS moderate dermal irritant, inhalation, and eye irritation ratings. It
also designated TCS as having low acute dermal and oral toxicity. It was not found to be a skin
sensitizer. The EPA does not consider TCS in drinking water a problem currently, nor is it likely
to become one. TCS exposure in the workplace does not seem to be a problem except for
occupations that deal with paint or pulp on site and paper production (EPA, 2008). A study in
Sweden found TCS in 3 out of 5 human milk samples (Adolfsson-Erici et al., 2002). A study was
done looking at the effects on the thyroid after adult humans used 0.3% TCS toothpaste. After 4
18
years, no negative effects were found (Cullinan et al., 2012). However, Clayton et al. (2011)
found that individuals were more likely to have allergies or hay fever if they had a greater
concentration of TCS in their urine.
4.2 Analysis of mycoremediation methods
As mentioned above mycoremediation is a possible bioremediation method for EDCs.
This section reviews different studies that have been done on this topic.
4.2.1 Direct fungal application methods
One type of mycoremediation is direct fungal application. This is where the mycelium is
used to remove toxins.
In Kim et al. (2008 (2)) the authors took a sample of Phlebia tremellosa from a forest
near Seoul and grew it on potato dextrose agar (PDA) with multiple EDCs. The authors then
grew the P. tremellosa in a liquid medium which contained phthalic EDCs. They also purified
laccase from the fungi and reduced estrogenic activity in the EDCs. The best temperature and pH
for laccase activity was 20°C and 4.0 respectively. In the presence of kojic acid laccase activity
was reduced 90%. When EDCs were added, laccase activity was reduced almost 90%. Laccase
lowered estrogenic activity at a lower rate when kojic acid was added.
Trametes versicolor has shown to be very effective in EDC remediation. A study by
Loffrendo et al. (2012) looked at how three different fungi species, Trametes versicolor, Stereum
hirsutum, and Pleurotus ostreatus, degraded BPA. Mycelium was grown on PDA disks. After
the colonies were established, mycelium absorbed BPA from water. Another variable examined
was the effect of two humic acids (HAs), leonardite (L-HA) and compost(C-HA), on the growth
of the mycelium. All fungi reduced the concentrations of BPA significantly. In the controls there
was a 20% reduction. T. versicolor was the best overall at removing the BPA, especially with
addition of L-HA and C-HA.
Heterobasidium insulare, Pleurotus ostreatus, and Stereum hirsutum, as whole-cell
biocatalysts are all capable of eliminating BPA from aqueous solutions (Lee et al., 2005; Hirano
et al., 2000). Only some strains will remove BPA. H. insulare and S. hirsutum used enzymes
other than LMEs to break down BPA. BPA’s side chain was broken down with dehydroxylation,
then carboxylation, ending with hydroxylation. 40% of BPA estrogenic activity was removed
from a 22.8 mg/L solution within 1 day by S. hirsutum. With S. hirsutum after 3 days all the
19
activity was removed. H. insulare eliminated all estrogenic activity within 1 day (Lee et al.,
2005; Hirano et al., 2000).
NP isomers were eliminated from aquatic medium with WRF Bjerkandera sp. BOL 13,
Cunninghamella sp., Phanerochaete chrysosporium, and Trametes versicolor. NP isomers were
also degraded with Clavariopsis aquatica (an aquatic hyphomycete) UHH 1-6-18-4 (a
mitosporic strain), and Fusarium and Mucor strains. The concentrations of the NP isomers
removed were between 11.02 mg/L and 99.16 mg/L (Cabana et al., 2007 (1)). Degradation
method and efficiency relies on growth conditions, type of NP isomer, and strain used.
Individual strains broke down NP isomers using various strategies. T. versicolor used laccase
and Bjerkandera sp. did not. T. versicolor worked better with agitation, where as Bjerkanera sp.
did better when kept still (Soares et al., 2006).
4.2.2 Enzyme application methods
Using free or immobilized enzymes has the benefit of not having to grow and stabilize
the fungi organisms. MnPs and laccases both show promise in the elimination of EDCs, although
their success is reliant on which WRF they come from. This process can be hindered by
environmental matrix complications like particulates or sludge slurries. These can lower EDC
bioavailability which can hinder remediation. This is due to the fact that in general phenolic
EDCs have high hydrophobicity and low water solubility. Matrix complications can also alter
enzymes (Zoungrana et al., 1997).
Cabana et al. (2007(1)) suggest that producing enzymes for remediation will probably
require WRF cultivation to have an increase in technology efficiency before it can be used for
WWT. More studies are needed to determine which isoenzymes are best for a particular job and
what EDCs metabolites are formed. The strain of WRF, growth conditions, and enzyme
preparation can all effect remediation success. Waste biomass, which is low cost, has been used
to make enzymes with a high activity.
4.2.2.1 Free Enzymes
Kim et al. (2008)(1) cloned laccase cDNA from T. versicolor and analyzed its
remediation properties with multiple EDCs. The authors also exposed EDCs (BBP, BPA, DEP,
NP) in a medium to cells of T. versicolor and noted their degradation. When cloned laccase was
used, it showed more activity and expression with benzylbutylphthalate (BBP) and DEP than it
did with BPA and NP. However, both BPA and NP showed higher activity and expression than
20
the control. T. versicolor cells in the liquid medium degraded over 95% of all EDCs. This result
suggests that multiple enzymes in T. versicolor are responsible for breaking down EDCs. The
authors showed that laccase is one of the remediating enzymes by exposing BPA and NP to it.
This was compared to a control and laccase with kojic acid, which is an inhibitor of laccase.
Laccase alone broke down the most BPA and NP followed by laccase and kojic acid. The control
degraded almost none of the EDCs. EDCs exposed to T. versicolor were analyzed for estrogenic
activities with yeast. After 48 hours the remediated solutions were almost free of estrogenic
activity, and after 72 hours there was no activity. NP takes the longest to degrade, up to 72 hours.
Hirano et al., (2000) describes an experiment where BPA is removed with MnP from
Pleurotus ostreatus O-48. The EDC had a concentration of 91 mg/L and was degraded with 10
U/mL MnP, 68.02 mg/L H2O2 and 302 mg/L manganese sulfate (MnSO4). This was done at room
temperature with a pH of 4.5. Tsutsumi et al., (2001) also used MnP to degrade 50 mg/L BPA as
well as 50.68 mg/L NP. H2O2, the co-substrate for MnP, was added by glucose oxidase. 100 U/L
of MnP and 7.55 mg/L of MnSO4 were used at 30°C and pH 4.5. Both EDCs had total
elimination in 1 hour.
Treatment conditions of BPA, NP, and TCS have been compared with the use of laccase
from Coriolopsis polyzona and Trametes villosa, as well as laccase from T. versicolor, which is
commercially available. In general, it was found that the best pH was between 5 and 6 and the
optimal temperature was between 45°C and 60°C. In general laccase stability increased at higher
pHs and catalytic activity increased at higher temperatures. Laccase from T. versicolor is
significantly stable between pH 4 and 8 (Kim & Nicell, 2006). In the Kim & Nicell (2006) study
laccase lost its activity faster when metabolizing BPA than with a buffer. In Mai et al., (2000)
study it was found that laccase stability was improved by phenols. Nitrite, sulfide, sulfite, and
thiosulfate anions hinder BPA remediation by T. versicolor laccase (Kim & Nicell, 2006).
Sulfide and sulfite compete with laccase for dissolved oxygen (DO). Acetone, formaldehyde and
methanol also decrease the ability of laccase to degrade BPA, possibly because they denature the
LME. High levels of ἄ-caprolactam, isoprene and phenol around 1000 µM and urea at 50 mg/L
did not hurt or help BPA metabolism. Fe3+ and Cu2+ hindered BPA degradation (Torres et al.,
2003). Calcium chloride, cobalt chloride and zinc chloride also hurt BPA conversion (Cabana et
al., 2007(1)). Cyanide lessened BPA remediation (Kim & Nicell, 2006).
21
4.2.2.2 Immobilized Enzymes
Free LMEs are problematic because they are not reusable and are quick to denature.
Additives or immobilization can correct these problems (Kim & Nicell, 2006; Modaressi et al.,
2005; Duran et al., 2002). Immobilization can keep enzymes active for a longer period of time,
increase LME efficiency, and make enzymes reusable (Duran et al., 2002). One problem is a low
activity to weight ratio. Additives have been found to have the same benefits (Cabana et al.,
2007; Kim & Nicell, 2006; Modaressi et al., 2005).
Cabana et al. (2009) conducted an experiment where they metabolized EDCs with
laccase onto diatomaceous earth pellets in a batch cylinder. The purpose of this study was to
immobilize laccase on solid support pellets and see how well it functioned under different
conditions. Immobilized laccase formed by simultaneous actions with glyoxal (GLY) kept more
than 95% of its original activity. Immobilized laccase formed by simultaneous actions with
glutaraldehyde (GLU) kept 90% of its original activity. Immobilized laccase made by sequential
procedures maintained around 30% of its original activity. Laccase in free form kept about 10%
of its original activity. Stabilizers in the form of natural proteins were also added.
The second part of the study tested the ability of the immobilized laccase to remove BPA,
TCS and p353NP in a packed bed reactor with different factors. The pellets that contained
immobilized laccase were put into a glass column that was 20 cm in height and 1.6 cm in
diameter. Then solutions with 5 mg/L or 100 mg/L BPA, TCS, and p353NP entered at a rate of 1
ml/min. 5 g of immobilized laccase CR633-GLU was used for BPA and TCS and 2.5 g was used
for p353NP. Five consecutive batches containing an EDC were treated for 200 min at 20°C and
pH 5. Between the cycles the pellets were cleaned with 100 ml of buffer. All of the batches were
remediated with about the same success. Chemical hydrophobicity directly impacted how well
the chemical adsorbed. In each batch the EDC had complete removal. Variance in temperatures
and pH were also tested. The different temperatures were 30°C, 40°C and 50°C. The different
pHs were 3, 4 and 5. The most effective removals were at 40°C and pH of 4, as well as 50°C and
pH of 5. All other combinations were found to be less effective (Cabana et al., 2009).
Adsorption ratio for 5 mg/L was between 40 and 60%. The ability of EDCs to adsorb to
the inactive laccase pellets was derived from their hydrophobicity and solubility. Immobilized
laccase seems to remove EDCs through adsorption and metabolism. The systems show
promising results in reusability (Cabana et al., 2009).
22
Cabana et al. (2011) looked at the ability of laccase from T. versicolor to eliminate
triclosan. This was done by conjugating with chitosan using 1-ethyl-3-(3-dimethylaminopropyl)
carbodiimide hydrochloride (EDCH). This strategy was tested with various glucosamine
monomer (NH2)/protein ratios, as well as with different EDCH/laccase ratios. To increase
stability from chemical and heat denaturation, immobilization techniques were developed. These
techniques also improved the storage and reusable properties of the conjugate. By combining
laccase and EDCH the efficiency improved from 12 to 60 times more than laccase alone. The
biocatalyst removed triclosan from aqueous solutions. It was indicated that the process formed
triclosan oligomers.
BPA can be metabolized with laccases from different strains of WRF including C.
polyzona, T. versicolor, T. villosa, and strain I-4. BPA was eliminated in 4 hours with
concentrations ranging from 5.02 to 502 mg/L, but only during certain conditions. BPA was
eliminated quickly when 10 to 1500 U/L was used. NP concentrations ranging from 5.07 to
70.51 mg/L have also been degraded by laccases. The strains used to produce the LME were C.
aquatic, C. polyzona, T. versicolor, I-4 and UHH 1-6-18-4. The most effective laccase (activity
of 1 U/L) was that from C. polyzona, which achieved total elimination in under 4 hours (Cabana
et al., 2007 (1)). All laccases successfully removed all NP under various conditions. Laccase
from C. polyzona and T. versicolor also removed TCS. However, with C. polyzona LME the
process was less efficient compared to BPA and NP (Cabana et al., 2007 (1); Kim & Nicell, 2006
(1)).
It is possible to increase laccase stability and activity by using stabilizing compounds.
These include compounds like alkyl betaine, Ficoll, polyethylene glycol (PEG), and polyvinyl
alcohol (Kim & Nicell, 2006; Modaressi et al., 2005). PEG increased BPA and TCS removal rate
(Kim & Nicell, 2006; Modaressi et al., 2005). With BPA it is thought to be done by BPA-PEG
coupling and enzyme protection. T. versicolor laccase still did not degrade TCS to the extent
BPA did. In the presence of cyanide and fluoride ions PEG does not increase laccase stability
(Kim & Nicell, 2006). Residual toxicity may depend on metabolites produced, which will vary
based on treatment, including stabilizing compounds. TCS had higher end toxicity when PEG
was added to the process (Kim & Nicell, 2006(2)).
Low-molecular weight oxidizable substances (LMWOS) are also capable of increasing
laccase activity. These include 1-hydroxy-benzo- triazole (1-HBT), 2,2′-azino-bis-(3-
23
ethylbenzthiazoline-6-sulfonic acid) (ABTS), 2,2,6,6-tetramethoxypiperidine 1-oxyl (TEMPO),
and violuric acid (VLA). LMWOS work by transferring electrons to the EDC (Bourbonnais &
Paice, 1990). BPA elimination by T. versicolor laccase was enhanced by the use of ABTS and
VLA. 1-HBT assisted P. ostreatus laccase in BPA removal (Tsutsumi et al., 2001; Kim & Nicell,
2006). ABTS also helped laccase of C. polyzona metabolism of BPA (Cabana et al., 2007 (1)).
Trametes sp. laccase has been immobilized on glass and put in a packed bed reactor
(PBR) to metabolize BPA (Iida et al., 2002). No activity was lost up to 50 doses of BPA (Iida et
al., 2003). This is better than the enzyme when it is free. When used with an electrolysis device
the immobilized laccase was more efficient at removing BPA (Iida et al., 2002). The electrolysis
made the system better at remediating shock loads (Iida et al., 2003).
Diano et al., (2007) study attached T. versicolor laccase to nylon-poly-(glycidyl
methacrylate) membrane and put it in a non-isothermal reactor to remediate BPA. The use of a
non-isothermal system increased the rate at which BPA was metabolized.
Enzymes may also be attached to a “physically defined matrix” like a dialysis tube
(Hoshino et al., 2003). The test was done with BPA. Remediation was successful.
Cross-linked enzyme aggregates (CLEAs) are formed by the immobilization of enzymes
using chemical cross-linking of the enzyme. This precipitation is done with the assistance of
bifunctional compounds. Possible additives recommended are bovine serum albumin (BSA) and
polyionic polymers (Cabana et al., 2007(2); Wilson et al., 2004). Laccase had the best activity
after 16 hours when precipitated with 1000 g/L of PEG. 200 μM of GLU was found to be the
cross linking agent that produced the greatest laccase activity and overall recovery. When BSA
was added the CLEAs had a greater size, but less activity (Cabana et al., 2007(2)).
Cabana et al. (2007(2)) tested the removal of BPA, NP, and TCS in a fluidized bed
reactor (FBR) with continuous 5 mg/L contaminated solution being treated. During the
experiments the water had a pH of 5 and was kept at room temperature. 5 U/L of catalase was
added to the water to eliminate H2O2. The FBR was a glass column with a height of 20 cm and a
diameter of 1.6 cm which held 0.5 mg of CLEAs. The solution had a flow rate of 1.5 ml/min.
After 50 min of reaction 30% of BPA and 90% of p353NP and TCS were eliminated. After 150
min of reaction about 90% or more of all EDCs were eliminated (Cabana et al., 2007(2)).
24
Figure 2: 0.5 mg laccase CLEA in FBR treatment of p353NP (triangle), BPA (circle), and TCS
(square) at concentrations of 5 mg/L. Taken from (Cabana et al., 2007(2)).
It is thought that PEG laccase had a higher activity than free laccase because PEG often
forms three dimensional structures with a lot of surface area which allows it to bind to more
water (Donato et al., 1996). Cabana et al. (2007(2)) suggest that adding BSA to the CLEAs
lowered activity because BSA covered the reaction center of laccase. Another explanation is that
a greater amount of BSA in the bulk material means less laccase that is available for reaction.
This can be seen in the scanning electron microscopy (Figure 3). However, even though BSA did
lessen laccase CLEA activity, it did increase the half-life. This may be because BSA encourages
cross-linking which helps enzymes stay folded (Cabana et al., 2007(2)). This has been found in
CLEAs with other enzymes (Shah et al., 2006).
CLEAs with laccase are more resistant than free laccase to heat damage because their
cross-linking helps keep them folded (Fernandez-Lafuente et al., 1995). This may not be true for
CLEAs with different enzymes (López-Gallego et al., 2005).
25
Figure 3: SEMs of CLEAs without BSA (A and B) and with BSA (C and D). Taken from
(Cabana et al., 2007(2)).
4.2.3 Metabolites found
TCS concentration of 72.38 mg/L was degraded using the strains T. versicolor SBUG-M,
DSM 11269, and DSM 11309, as well as Pycnoporus cinnabarinus SBUG-M 1044. Metabolites
formed were 2-O-(2,4,4′-trichlorodiphenyl ether)-β-D-xylopyranoside, 2-O-(2,4,4′-
trichlorodiphenyl ether)-β-D-glucopyranoside and 2,4-dichlorophenol. The last one was made
by phenoxy radicalization. 2,4,4′-trichloro-2′-methoxydiphenyl ether was also made. The
enzymes used were not LMEs (Cabana et al., 2007 (1)).
It is widely recognized that often WRF or their enzymes metabolize EDCs into
compounds that have little or no estrogenic activity (Cabana et al., 2007 (1)). BPA and NP
eliminated with laccase produce compounds with little to no estrogenic activity (Cabana et al.,
2007 (1); Tsutsumi et al., 2001). Metabolites of MnP degraded BPA and NP may have
estrogenic activity (Tsutsumi et al., 2001). It is still unknown how these compounds react in an
environmental setting and if they start to show EDC properties. Metabolism products of BPA
and TCS were found to be dimers, trimers and tetramers. NP metabolites were found in these
forms also, as well as pentamers (Cabana et al., 2007 (1)). The dimer made with T. villosa
laccase metabolized BPA was 5,5′-bis-[1-(4-hydroxy-phenyl)-1-methyl-ethyl]biphenyl-2,2′-diol
(Uchida et al., 2001). BPA can also break down to hexestrol, phenol, 4-isopropenylphenol and 4-
isopropylphenol (Fukuda et al., 2004; Uchida et al., 2001; Fukuda et al., 2001).
26
4.3 Analysis phytoremediation methods
As mentioned above phytoremediation is a possible bioremediation method for EDCs.
This section reviews different studies that have been done on this topic.
4.3.1 Plant application methods
Loffredo et al. (2010) looked at the capability of eight plants, five grasses and three
horticulture plants to remove BPA. The grasses were Agropyron fragile, Cynodon dactylon,
Festuca arundinacea, Lolium perenne, and Trifolium repens. The horticulture plants were
Cucumis sativus, Cucurbita pepo, and Raphanus sativus. The first set of tests measured BPA
uptake in relation to germination. A subset of this test was growth experiments in axenic and
septic conditions. The second set of tests measured the amount of BPA in the plants and growing
medium during a growth period of 16 days. Two concentrations of BPA were used for the tests,
4.6 mg/L and 46 mg/L (Loffredo et al., 2010).
In the first test the overall result is that germination was not hindered by BPA in the
growth medium. Exceptions were A. fragile and L. perenne which had their root growth inhibited
with 46 mg/L BPA. Also, C. pepo had more fresh mass with 4.6 mg/L. In this test all plants
removed a portion of BPA, while the controls saw basically no removal. This is consistent with
other findings (Imai et al., 2007). L. perenne (perennial ryegrass) and R. sativus (radish) were
among the top BPA removers in both concentration experiments. Overall, every species
eliminated about 10.6 times more BPA when exposed to the greater concentration. This implies
that plants absorb/degrade BPA proportionally to the quantity they are exposed to (Loffredo et
al., 2010).
C. pepo, F. arundinacea, L. perenne, and R. sativus, were germinated in both axenic and
septic mediums with a BPA concentration of 46 mg/L. In axenic conditions F. arundinacea root
mass, fresh mass, and C. pepo plant shoot length were hindered. F. arundinacea root length was
hindered in septic systems. L. perenne and C. pepo plants did not have significant BPA removal
abilities in one system over the other. However, F. arundinacea and R. sativus removed more
BPA in septic experiments. It is thought this is because microorganisms also absorbed BPA
along with these plants. When this test was done with 4.6 mg/L no species except L. perenne
showed favoritism to a medium. L. perenne degraded more BPA when germinated axenically.
R. sativus and L. perenne were grown in pots with 4.6 mg/L and 46 mg/L for 16 days to measure
plant growth and BPA reduction. At 46 mg/L half of the R. sativus died and the roots were
27
sustained harm. The fresh mass of both species were lowered. Plants had less fresh mass, live
plants, root length and shoot length after germination when exposed to BPA, especially at higher
concentrations. BPA concentrations decreased in all growth mediums that had plants. The
control with 4.6 mg/L BPA also lowered in concentration rapidly around day 10. The only
sample to keep most of its BPA was the control with 46 mg/L. It is thought that the control with
less BPA lowered in concentration because of microorganism activity. The authors suspect that
the greater concentration of BPA hindered microorganism activity. If the experiment had gone
on for more than 16 days this may have been changed. L. perenne removed 97% of 4.6 mg/L
concentration and 90% of 46 mg/L concentration. R. sativus removed 82% and 95.5% for the
same respective concentrations. In all four samples 79% to 99% of BPA was taken up and
degraded. 0.2% to 2.7% was amassed in the plants. 0.6% to 10.1% of BPA was found in the
solution (Figure 4). It is thought that BPA taken up by plants is quickly degraded. This is
supported in other literature. It is thought that microorganisms with the L. perenne feed off the
exudates that the plant gave off and helped it degrade BPA. It seems that the R. sativus with the
high concentration did the same, although at a lesser concentration the microorganisms fed off
exudates without helping remediate BPA (Loffredo et al., 2010).
Figure 4: BPA location in plants and medium after 16 days (Loffredo et al., 2010).
28
In Shimoda et al. (2009) the authors transformed BPA and benzophenone (BZP) into
glycosides. These were taken up by the cells which decreased the concentration in growth
medium. Cells used were Nicotiana tabacum. Tests where the cells were immobilized used
sodium alginate gel. The sodium alginate concentration was 2% because this was found to be the
most effective concentration.
The Imai et al. (2007) study had several parts to it. The first was testing plants to see
which was ideal at removing BPA. The authors found that Portulaca oleracea was one of the
best garden plants, out of about 100, to remove 50 µM of BPA in sterile conditions. It was
chosen as the main plant of study because it is a hardy plant. This is important for
phytoremediation candidates. Treated water was tested for endocrine disrupting abilities. As
BPA was removed so was the endocrine disrupting abilities of treated water (Imai et al., 2007).
Imai et al. (2007) tested P. oleracea ability to remove BPA and NP. NP was metabolized
at similar speeds to BPA. As EDCs were removed from solution their estrogenic abilities fell. P.
oleracea did not show a significant ability in removing phthalates. Concentration does not seem
to be a deciding factor for P. oleracea metabolism of BPA. With concentrations of less than 57.1
mg/L, BPA had almost total degradation in 24 hours. At 114.1 mg/L, the highest concentration,
over 95% was removed in the same amount of time. It was calculated that the average removal
rate was 1.25 µmol/gram of plant/hour. P. oleracea’s ability to remove BPA was also tested with
different light conditions: constant light, constant dark, and 8 hours of light with 16 hours of
dark. The test was done for over 90 hours. Metabolism of BPA did not change significantly with
light variances. Temperature can also affect plant metabolic rate. P. oleracea was grown in 15°C,
25°C, and 30°C for 24 hours. At the end of the test almost all BPA was removed. However,
during earlier times in the experiment BPA removal increased with temperature. This is not
thought to be too significant though. BPA removal ability was also tested with pH. It was found
that the removal is best between pHs of 4 to 7, with 6 being optimal. It is thought that BPA was
not degraded in alkaline solution because the alkalinity hindered enzyme activity. Overall, P.
oleracea seems like a good candidate for phenolic EDC removal. Degradation is high, although
it is hard to compare because there are few studies on the same topic. It is thought that Product A
(Figure 5, metabolite directly below BPA) is the first compound to be produced when BPA is
degraded, followed by other compounds made from glucosylated and/or hydroxylated processes
(Imai et al., 2007).
29
Authors tested the ability of P. oleracea to degrade multiple EDCs with 11 mg/L of BPA,
5.2 mg/L of OP, and 5.5 mg/L of NP. All EDCs experienced almost total metabolism at 24
hours. This suggests that the presence of multiple EDCs does not affect the capability of P.
oleracea to degrade them (Imai et al., 2007). After 12 hours about 90% of BPA was removed
from contaminated solution, at 24 hours over 95% was removed, and after 48 hours BPA was
essentially gone from the solution. This is a fairly rapid removal, especially since plants had no
assistance from microorganisms. P. oleracea was tested for BPA concentration at 3 hours and 24
hours. At the 3 hour mark 35% of BPA was removed from the water, but none was found in
plant tissue. When tested at 24 hours 95% had been removed from solution and a small amount
was measured in the plants. This suggests that P. oleracea does not significantly absorb or
adsorb BPA and accumulate it. When measured separately, with an end time of 100 hours, heat-
treated roots remediated less than shoots, which remediated less than untreated roots. This
implies that there is a root enzyme that is primarily responsible for metabolizing BPA into a non-
endocrine disruptor (Imai et al., 2007).
The authors in Nakajima et al. (2007) exposed 8 species of green algae to BPA. The
species were Carteria cerasiformis, Coelastrum reticulatum, Cyanophora paradoxa, Gonium
pectoral, Micracinium pusillum, Pseudokirchneriella subcapitata, Scenedesmus acutus,
Scenedesmus quadricauda. All species except P. subcapitata had unhindered growth at a BPA
concentration of 10 mg/L. At a concentration of 5 mg/L and lower, no algae showed negative
responses. All species of algae reduced solution concentration of BPA at varying levels. S.
acutus removed the most BPA. BPA concentration in algae cells were also measured to see how
much was being accumulated. This was done using radiolabeled BPA. G.pectorale and C.
paradoxa mostly removed BPA by accumulation. C. cerasiformis, C. paradoxa, M. pusillum, P.
subcapitata, S. acutus, and S. quadricauda algae degraded more BPA than they accumulated.
This indicates that they metabolized BPA. It is notable that the concentration of BPA leveled off
after 2 days. This indicates that algae metabolized or discharged absorbed BPA. These species
were put in a solution with a BPA concentration of 10 mg/L. The solution was analyzed after 10
days to identify metabolites.
Eucalyptus perriniana has shown to be a BPA remediation candidate (Hamada et al.,
2002). The processes used for this are glycosylation and hydroxylation. In the other tests only
three metabolites were found, but in this test two new ones were identified. BZP was
30
hydroxylated and reduced into different compounds.12 mg of BPA was put into flasks with
cultured E. perriniana cells. The flasks were incubated for 7 days in the dark at 25°C while being
shaken. The metabolites were extracted from the cells.
A study by Zhang et al. (2008) looked at phytoremediation of alkylphenols (APs) NP and
octylphenol (OP) in Moon Lake, China. They measured the concentration of the pollutants in
three locations progressively further away from a disused sewage discharge outlet. The amount
of APs in water and sediment got lower as distance from the outlet increased. Submersed aquatic
plants were also measured for AP concentration for locations further from the outlet. No flora
was tested directly near the outlet because no plants grew there. The species tested were
Ceratophyllum oryzetorum, Elodea nuttallii, Myriophyllum verticillatum, and Potamageton
crispus. The plant that accumulated the most NP was M. verticillatum, followed by E. nuttallii,
and C. oryzetorum, with P. crispus having the least. E. nuttallii absorbed the most OP, then M.
verticillatum, P. crispus and C. oryzetorum. P. crispus absorption of APs was measured during
the months of March, April and May. For both APs the highest concentration was found in May,
followed by April, then March. The authors concluded AP presence in the environment was due
to the old sewage discharge outlet and the local use of APs in nonionic surfactants (Zhang et al.,
2008).
Takahashi et al. (2005) tested 50 plants from an abandoned rice field for BPA
remediation abilities. This was done by testing the plants remediation of the dye Remzol Brilliant
Blue R both with and without microorganisms. It was found that Rumex crispus japonicus (curly
dock) is well adept at remediation of the dye. Based on this, the authors compared aseptic curly
dock and rice in removal of BPA from aquatic medium. Curly dock plants used were 0.2 g fresh
weight and rice plants used were 0.1 g fresh weight. BPA was added to the medium at a
concentration of 40 mg/L for curly dock and 20 mg/L for rice. This came out to a plant mass
ratio of 1000 mg BPA/kg plant.
It was found that both plants removed BPA from the medium relatively quickly. After 6
days curly dock removed about 40% of BPA and rice removed about 60%. After 15 days curly
dock almost completely removed BPA and rice removed about 80%. It is notable that curly dock
had twice the concentration of rice, which indicates that it may be a better remediator. After 7,
14, and 21 days plants were tested for BPA concentration. After 7 days curly dock had about 62
mg BPA/kg plant and rice had about 24 mg BPA/kg plant. After 14 days no BPA was found in
31
curly dock and about 12 mg BPA/kg plant was found in rice. After 21 days rice had marginal
BPA and curly dock had none. It is notable that amount found was less than 10% of BPA
removed from medium. This implies that absorbed BPA is metabolized. Furthermore, BPA in
plants may be made into a metabolite that is not capable of being removed from the plant
(Takahashi et al., 2005). This idea is supported in Noureddin et al. (2004).
Curly dock and rice were also tested in their capacity to remove environmental
concentrations of BPA. BPA was added to the medium at a concentration of 100 μg BPA/kg of
plant with 10 g of each plant being used. After 15 days plants and medium were tested for BPA.
No BPA was found in medium and BPA was found in both plants at about 70% (Takahashi et al.,
2005). While the results are consistent with Noureddin et al. (2004(2)) study in terms of rice, this
subject needs more comprehensive studies (Takahashi et al., 2005).
Reinhold et al. (2010) did a study examining the removal of eight pollutants by Landoltia
punctata and Lemna minor. The pollutants were atrazine, clofibric acid, 2,4-
dichlorophenoxyacetic acid, fluoxetine, ibuprofen, meta-N,N-diethyl toluamine, picloram, and
triclosan. Four of these, 2,4-dichlorophenoxyacetic acid, fluoxetine, ibuprofen, and triclosan
were found to be removed from contaminated solution. TCS was reduced from a 2.9 mg/L
solution by plants at a rate of 80% after 2 days, 90% after 6 days. TCS was also reduced in
chemically-inactive plant flasks in both light and dark conditions. Removal with active plants
was the quickest and most effective method. TCS was removed by degradation and
accumulation, as well as adsorption. The metabolite 2,4-dichlorophenol (2,4-DCP) was found in
chemically-inactive flasks, but not ones with healthy plants. This indicates that active L. punctata
and L. minor active plants absorb 2,4-DCP greater than inactive ones. 2-chlorophenol and 4-
chlorophenol are TCS metabolites that were analyzed for, but not found.
Genetically modified plants can be used for more efficient phytoremediation. This
process may be preferred over phytoremediation without genetic modification because
hydrophobic contaminants are not likely to be taken up by plants because remediation takes
place in the rhizosphere. Sonoki et al. (2005) produced transgenic tobacco plants (Nicotiana
tabacum) that released the laccase III enzyme that is produced in Trametes versicolor. Laccase
enzymes of fungi have been known to breakdown various organic contaminants such as BPA,
chlorinated hydrolyl biphenyl, nonylphenol, and polymerize chlorophenol. N. tabacum was used
because it has successfully been genetically modified to phytoremediate other toxins such as
32
glycerol trinitrate and dimethyl mercury. Two strains of Eschericha coli, JM109 and HB101,
were used to alter DNA.
A test was done to confirm that N. tabacum roots put active laccase enzymes in the
surroundings. If used in remediation, the roots can be used to alter the chemistry of contaminated
rhizospheres. When plants were two months old they were put into a hydrophonic culture. 10
μmol of BPA was added and effluent was monitored for BPA removal. The control plant
removed some of the EDC, but modified versions of tobacco removed considerably more. All of
the N. tabacum plants expressed the ability to remove BPA from the medium. This study did not
find reaction products. However, it is unlikely that products had estrogenic activity. Solution
from transgenic tobacco had less estrogenic activity than those of the control plants. It is thought
that the BPA may have been polymerized or degraded. Modified N.tabacum plants also removed
PCPs from solution in a greater quantity than control plants (Sonoki et al., 2005).
A test was done to confirm that N. tabacum roots put active laccase enzymes in the
surroundings. If used in remediation, the roots can be used to alter the chemistry of contaminated
rhizospheres. When plants were two months old they were put into a hydrophonic culture. 10
μmol of BPA was added and effluent was monitored for BPA removal. The control plant
removed some of the EDC, but modified versions of tobacco removed considerably more. All of
the N. tabacum plants expressed the ability to remove BPA from the medium. This study did not
find reaction products. However, it is unlikely that products had estrogenic activity. Solution
from transgenic tobacco had less estrogenic activity than those of the control plants. It is thought
that the BPA may have been polymerized or degraded. Modified N.tabacum plants also removed
PCPs from solution in a greater quantity than control plants (Sonoki et al., 2005).
4.3.2 Metabolites found
BPA may be degraded to glycoside, BPA’s hydroxylated form, or monophenols by plants
(Nakajima et al., 2007). Nakajima et al. (2002) found that tobacco seedlings and cells degrade
BPA into BPA glycoside (BPAGlc) by glycosylation. The plants then excrete metabolites into
medium, or they make BPAGlc into BPA-di-O-β-D-glucopyranoside or BPA-mono-O-β-D-
gentiobioside. Algae may be able to do this also, but the authors could not confirm this. In higher
plants phenol glucosyl transfers glycosylates BPA (Pridham, 1964; Kreuz et al., 1996). Algae
may also do this, and it was observed in S. quadricauda. BPAGlc has a third of the estrogenic
activity of BPA (Morohoshi et al., 2003). Metabolites formed from C. reticulatum,
33
P.subcapitata, and S. acutus are thought to be BPA-mono-O-β-D-glucopyranoside. S.
quadricauda metabolite is thought to be BPA-mono-O-β-D-galactopyranoside. None of the
metabolites in M. pusillum were identified. This shows that BPA is metabolized into BPA
glycosides by some algae species. The ability seems to be the deciding factor in which species
accumulate or reduce BPA. They did not find cleavage products of BPA (Nakajima et al., 2002).
BPAGlc may become BPA in animal gut (Kanank and Sullivan, 1966). It also can
accumulate in leaves (Nakajima et al., 2002; Noureddin et al., 2004). This makes environmental
BPAGlc a risk and should be taken into account.
The BPA metabolites from Kondo et al. (2006) identified were:
1) 2,2-bis(4-β-D-glucopyranosyloxyphenyl)propane
2) 2-(4-β-D- glucopyranosyloxy-4-hydroxyphenyl)-2-(4-β-D-glucopyranosyloxyphenyl)
propane
3) 2-(3-β-D-glucopyranosyloxy-4- hydroxyphenyl)-2-(4-β-D-glucopyranosyloxyphenyl)
propane
4) 2-(3-β-D-glucopyranosyloxy-4-β-D-glucopyranosyloxyphenyl)-2-(4-hydroxyphenyl)
propane
5) 2-(4-β-D-glucopyranosyloxy-3-hydroxyphenyl)-2-(3-β-D-glucopyranosyloxy-4-
hydroxyphenyl) propane
Metabolites numbers 3 and 4 had not been found in previous studies. Numbers 1 and 2
were not found in the same flasks. This indicates that 2-(4-hydroxyphenyl)-2-(3,4-
dihydroxyphenyl)propane is the intermediate between BPA and 2. Likewise, numbers 3 and 5
were not found in the same flask. This indicates that 2,2-bis(3,4-dihydroxyphenyl)propane is the
intermediate between BPA and 5 (Figure 5). The metabolites have low endocrine disruption
activity (Nishikawa et al., 1999; Hamada et al., 2002). More studies are being conducted on this
subject.
Over a period of 48 hours BZP was made into 4-O-β-D-glucopyranosylbenzophenone,
diphenylmethyl β-D-glucopyranoside, and diphenylmethyl 6-O-(β-D-glucopyranosyl)- β-D-
glucopyranoside. Cultured cells transformed 35% of BZP into glycosides. Immobilized cells
altered 64% of BZP. Overall, immobilized N. tabacum cells made 1.8 times more product than
cultured cells. Immobilized cells also converted BZP at a quicker rate (Shimoda et al., 2009).
Over a period of 48 hours BPA was changed to bisphenol 2,2-bis(4-β-D-
34
glucopyranosyloxyphenyl) propane, 2-(4-β-D-glucopyranosyloxy-3-hydroxyphenyl)- 2-(4-β-D-
glucopyranosyloxyphenyl) propane, and 2-(3-β-D-glucopyranosyloxy-4- hydroxyphenyl)-2-(4-β-
D-glucopyranosyloxyph enyl)propane. Cultured cells altered 29% of BPA. Immobilized cells
converted 50% of BPA into glycosides. Immobilized cells produced 1.7 times more glycosides
than cultured cells. Again, the immobilized N. tabacum cells transformed the EDC at a quicker
rate (Shimoda et al., 2009).
Figure 5: BPA metabolites from E. perriniana (Kondo et al., 2006).
4.4 Best myco- and phytoremediation methods
An analysis of mycoremediation methods (see Appendix 6.1) showed that a constructed
wetland with Portulaca oleracea, Landoltia punctata, and Lemna minor would be the most
effective at remediating BPA, NP, and TCS. After an analysis of phytoremediation methods (see
Appendix 6.2) it was found that a PBR filled with CLEAs with laccase would be the best
remediation for the three EDCs.
Treatment designs suggested here are meant to have the capacity of treating 864 m3/day.
The flow rates are 0.01 m3/s. The volumes of all sections are 1000 m3. The CW was given twice
35
as much area so that all EDCs have equal remediation volumes. By keeping these parameters the
same it is possible to compare the efficiency of EDC removal.
Water quality modeling was done to compare the remediation abilities between both
types of remediating systems. The program used was Hydrologic Engineering Center – River
Analysis System (HEC-RAS) version 4.1 from U.S. Army Corps of Engineers. Several
parameters were added to the system for the models. These were channel width, depth, and
height, original pollutant concentration, incoming pollutant concentration, flow rate, and
continuous flow. Decay constants were calculated and included in the model. The channel
widths, depths, and heights corresponded with the phytoremediation CWs and mycoremediation
tank. Both original and incoming pollutant concentration was set to 1 μg/L. This is a realistic
assumption based on the Jackson and Sutton study (2008) which found TCS in treated
wastewater at a concentration of 0.9 μg/L. Flow rate was input at 0.01 m3/s. Continuous flow was
chosen so that plants are able to continually remediate with less chance of a spike in
concentration. Continuous flow also gives the opportunity to have the mycoremediation tank
used to its fullest extent.
The decay constants were calculated under the assumption that the EDC remediation
follows first order kinetics. This is approximately true because the Michaelis-Menten constant
for immobilized laccase is significantly higher than the typical concentration of pollutant. The
Michaelis-Menten constant (Km) for laccase and BPA is 0.23 mM. Km of laccase and NP is 0.45
mM, and for laccase and TCS it is 0.12 mM (Cabana et al., 2009). This is significantly higher
than the BPA, NP, and TCS concentrations which are 0.00438 μM, 0.00454 μM, and 0.00345
μM respectively, which indicates the reaction is 1st order. Decay constants were calculated for
phytoremediation from Imai et al. (2007) for BPA and NP and Reinhold et al. (2010) for TCS.
All mycoremediation decay constants were calculated from Cabana et al (2007(2)). The
mycoremediation decay constant for BPA is 12.3/day, and for NP and TCS it is 55.3.
Phytoremediation decay constants for BPA and NP are 2.30/day, and for TCS it is 0.916/day:
K = ln
C0
C t
t1−t0
K = decay constant (1/day); C0 = initial concentration (mg/L); Ct = concentration after t time
(mg/L); t0 = start time (day); t1 = end time (day)
36
Mycoremediation Values calculated from Cabana et al. (2007 (2))
BPA: C0 = 5 mg/L; Ct = 3 mg/L; t0 = 0 hr; t1 = 1 hr
K = 0.510826/hr : K = 12.25982/day
NP: C0 = 5 mg/L; Ct = 0.5 mg/L; t0 = 0 hr; t1 = 1 hr
K = 2.302585/hr : K = 55.26204/day
TCS: C0 = 5 mg/L; Ct = 0.5 mg/L; t0 = 0 hr; t1 = 1 hr
K = 2.302585/hr : K = 55.26204/day
Phytoremediation Values calculated from Imai et al. (2007) and Reinhold et al. (2010)
BPA: C0 = 11 mg/L; Ct = 1.1 mg/L; t0 = 0 day; t1 = 1 day
K = 2.302585/day
NP: C0 = 8.8 mg/L; Ct = 0.88 mg/L; t0 = 0 day; t1 = 1 day
K = 2.302585/day
TCS: C0 = 2.9 mg/L; Ct = 1.16 mg/L; t0 = 0 day; t1 = 1 day
K = 0.916291/day
4.4.1 Viability of mycoremediation system
The mycoremediation concrete tank contains diatomaceous earth pellets coated with
CLEAs with laccase which is capable of remediating all three EDCs. The tank is designed to
have a width of 10 m, height of 1 m, and length of 100 m.
37
Figure 6: Mycoremediation PBR Layout. Blue circles represent pellets.
Mycoremediation degradation was modeled with HEC-RAS to see the change in
concentration as a function of position x in the tank (Figure 7). Parameters mentioned in section
4.4 were applied. The simulation was run until convergence was reached. It was found that NP
and TCS were remediated the quickest and more completely than BPA. The final concentration
for NP and TCS was 2.36 × 10-27 mg/L. Final concentration for BPA was 1.22 × 10-9 mg/L.
Models were checked with analytical calculations, using:
c x=co e−kx
u
and were found to be close to model outputs.
Cx = concentration after x distance (mg/L) C0 = initial concentration (mg/L); K = decay constant
(1/day); u = velocity (m/s); x = distance (m)
For all remediation models C0 concentrations are 0.001 mg/L, velocity (u) is 86.4 m/d, and tank
length is 100 m.
BPA: K = 12.25982/day
C100 = 6.87911 × 10 -10 mg/L
NP: K = 55.26204/day
C100 = 1.6681 × 10 -31 mg/L
TCS: K = 55.26204/day
C100 = 1.6681 × 10 -31 mg/L
38
Figure 7. Mycoremediation decrease in concentration of BPA (blue line), NP and TCS (black
line) and a conservative constituent (green line). Constructed from HEC-RAS.
4.4.2 Viability of phytoremediation system
The suggested constructed wetlands design for the EBMUD WWTP has two successive
parts to ensure optimal EDC degradation. Both tanks are made out of concrete. The first CW
(CW 1) contains Portulaca oleracea to remove BPA and NP. The length and width of CW 1 are
100m. The depth of the CW is 30 cm, with a 10 cm water depth. CW 1 contains plastic mesh
substrate that P. oleracea grows on. Plastic is preferable over sediment because EDCs have high
Kows, so they would bond to the sediment and be less likely to go through the remediation
process. BPA Kow is 3.40-3.82, NP is 4.48, and TCS is 4.76 (Staples et al., 1998; Ahel & Giger,
1993; Park et al., 2009).
The second CW (CW 2) contains the species Landoltia punctata and Lemna minor to
remediate TCS. CW 2 has a length and width of 100m and 200m respectively with a depth of 10
cm. Water flows through the CW at a height of 5 cm. Passive vertical turbulence increasing
mechanisms line the bottom of this section. Turbulence can be increased by attaching metal to
the floor at an appropriate angle (Figure 9). This mechanism would help ensure that pollutants
come in contact with the shallow roots. No substrate is needed for the species in this CW.
Figure 8: Constructed Wetland Layout: View from Top.
39
CW construction should include a wildlife barrier, such as hoops with netting, to protect
both wildlife and the remediation facility. This is especially important in CW 2 because the
plants used are known to be palatable to birds, especially ducks. One could implement a wildlife
barrier up by running supports down the length of the CW, and attaching hoops onto those
supports. A strong fabric net can then be laid over the hoops.
Figure 9: Constructed Wetland Layout: View from Side.
Portulaca oleracea, common name purslane, is a plant native to India. It is an annual
which grows about 15 cm high and about 30 cm laterally. Its leaves are 3 cm long (Missouri
Botanical Garden, 2013). This means that it would have to be replanted each year in the CW. P.
oleracea germination can take place 6 to 8 weeks in a greenhouse and then transplanted outside
after the last frost. It has flowers that bloom from June to frost, and range in color with red, pink,
orange, yellow, and white blossoms. P. oleracea grows in hardiness zones from 2 to 11
(Missouri Botanical Garden, 2013). SF east bay is 9b hardiness zone (National Gardening
Association, 2013), so purslane will likely thrive there. The species requires full sun and medium
to low soil saturation. This implies if it is planted in a CW, there would have to be a low flow
rate.
Landoltia punctata is an aquatic floating plant commonly called dotted duckweed. This
species does not have leaves; it has fronds, which can be 1 to 5 mm long. Thin roots, mostly in
groups of 2 to 4, grow under the frond and can be 2 to 10 mm long. This plant is commonly
found in thick mats. It does not require substrate, but prefers water with low flow (Global
Invasive Species Database, 2006). This species is known for its general water remediation ability
40
(Global Invasive Species Database, 2006). This supports the idea of L. punctata being used as a
remediator. This plant is found in California as an invasive (Global Invasive Species Database,
2006). Special care would have to be taken to ensure that this plant is not released into
susceptible ecosystems.
Lemna minor is a U.S. native that is considered invasive in some states. It is known as
common duckweed or lesser duckweed (Washington State Department of Ecology, 2013). L.
minor have fronds that are 2 to 5 mm in diameter with roots 5 mm deep (Missouri Botanical
Garden, 2013) (Washington State Department of Ecology, 2013). It is a perennial that grows in
hardiness zones 4 to 10, so it is suitable for growth in the SF east bay. It is a perennial that will
grow in full sun or partial shade. L. minor grows quickest in spring and fall. Like L. punctata, it
does not require substrate to grow. Birds, especially ducks, eat this species (Missouri Botanical
Garden, 2013). It may be required to have a barrier so birds do not enter the CW.
Phytoremediation degradation was also modeled with HEC-RAS to see the change in
concentration (Figure 10). Parameters mentioned in section 4.4 were applied. The system was
run until equilibrium was reached. The same contaminated effluent was assumed to flow through
both tanks for complete remediation. While in actuality this might result in P. oleracea
degrading some TCS, and L. punctata or L. minor remediating BPA or NP, no evidence was
found of this in the literature. Therefore, the model assumed the conservative estimate that no
pollutant remediation took place outside the tank intended for it. The final concentration for BPA
and NP was 7.11 × 10-5 mg/L. Final concentration for TCS was 3.47 × 10-4 mg/L. Models were
checked with analytical calculations, which were found to be consistent with model outputs:
BPA: K = 2.302585/day
C100 = 6.95973 ×10 -5 mg/L
NP: K = 2.302585/day
C100 = 6.95973 ×10 -5 mg/L
TCS: K = 0.916291/day
C100 = 3.46275 ×10 -4 mg/L
41
Figure 10. Phytoremediation decrease in concentration of BPA and NP (black line), and TCS
(blue line). Constructed from HEC-RAS.
4.5 Pilot studies
This section reviews pilot studies done with myco- and phytoremediation methods.
4.5.1 Mycoremediation pilot studies
Unfortunately, mycoremediation is not yet a popular method for WW remediation, so
there are few to no pilot studies. However, some of the maintenance required for such a system
can be assumed based on knowledge from lab studies. One such procedure is CLEA cleaning
with buffer for reuse (Cabana et al., 2009). Furthermore, the system would need general
maintenance required for any filtration system.
4.5.2 Phytoremediation pilot studies
Zarate et al. (2012) examined TCS and triclocarban (TCC) bioconcentrations in a CW in
Denton, Texas. A metabolite of TCS, methyl-triclosan, was also examined, but was under
detection limit in all of the samples. The CW treats about 1% (4.54 * 105 L) of the WWTP’s total
daily effluent. The wetland is 46 m * 46 m. It has an average retention time of 4.3 days and an
average flow rate of 2968 L/hr. The species tested were Pontederia cordata, Sagittaria
graminea, and Typha latifolia. Other species in the CW, that were not tested, are Ceratophyllum
demersum, Lemna sp., Paspalum sp., and Potamageton sp. Concentration analysis was done for
both the roots and shoots of all the species, and in the beginning, middle, and end of the CW.
42
Comparison of different species root concentration took place at the WW effluent inlet.
P. cordata concentrated TCS in the roots at 15.0 ng/g. S. graminea had a TCS root concentration
at 25 ng/g. TCS root concentration for T. latifolia was 40.3 ng/g. Shoot concentrations were 19
ng/g for P. cordata, 18 ng/g for S. graminea, and 17 ng/g for T. latifolia. S. graminea and T.
latifolia shoot concentrations were determined with 3 or more substituted measurements.
Overall, root concentrations decreased as distance from inlet increased. TCS concentrations in
the roots and shoots were measured for T. latifolia at the inlet, half-way between the inlet and
outlet, and outlet. At the inlet the root concentration was 40 ng/g. At the half-way point the root
concentration was 28 ng/g. At the outlet the root concentration was 12 ng/g. All shoot
concentrations were measured at about 17 ng/g, which is the detection limit. Shoot
concentrations were determined with 3 or more substituted measurements. Overall, this study
shows that plants accumulate TCS at different concentrations, and that roots accumulate the EDC
at higher concentrations than the shoots (Zarate et al., 2012).
Zarate et al. (2012) suspect their paper is the first to identify species and tissue TCS and
triclocarban (TCC) concentration variations in a working CW. Their suspicion points out the
need for further study in this field, especially with different species. They also point out that high
TCS concentrations can harm plants and their subsequent wetlands.
Park et al. (2009) did a study looking at the reduction of PPCPs and EDCs in a CW that
remediates WW effluent. The CW has an average width of 30 m, length of 120 m, and depth of
0.13 m. The flow rate is 1800 m3/day with about a 6 hr retention time. The CW has different
sections. One section has Acorus sp. and another has Typha sp. The study looked at the ability of
both plants to reduce atenolol, carabamazepine, diazepam, diclofenac, dilantin, naproxen,
sulfamethoxazole, TCEP and TCS from WW.
TCS was measured in WW effluent at about 15 ng/L in May 2007 and about 20 ng/L in
August 2007. It was calculated that in May Acorus sp. removed 74% of TCS and Typha sp.
removed 80%. The wetland effluent had a total of 75% removal. In May, both species removed
atenolol, carbamazepine, diclofenac, and naproxen over 50%. Diazepam had an increase in
concentration with Typha sp. and wetland effluent. Dilantin had an increase in concentration
with Typha sp. In August Acorus sp. removed -14% and Typha sp. removed 100%, with total
wetland removal being 100%. No hypothesis was given as to why concentrations increased in the
Acorus pond in August. August concentration also increased during Acorus treatment for
43
carbamazepine and TCEP, as well as total wetland effluent for carbamazepine. Authors tried to
find a pollutant removal relation to pKa and log Kow but found none. Authors suggest more
studies to examine the anoxic bioremediation processes by which PPCPs and EDCs are removed
(Park et al., 2009).
Waltman et al. (2006) examined the Denton municipal WWTP and its adjoining CW
abilities to remove TCS. The treatment plant uses conventional activated sludge methods. It uses
UV as its disinfectant and has the ability to remediate 21 million gallons/day. The CW is an
experimental unit that is 46 m by 46 m and can hold a maximum of 570,000 L. It is split into
four with-flow sections with depths ranging from a couple to 60 cm, increasing with distance
from inlet. The CW has a clay bottom to keep seepage from groundwater.
TCS concentrations were measured at the WWTP influent and effluent, the wetland
influent and effluent, and a section of Pecan Creek about 300 m downstream of where the
effluent enters it. It was found that the average concentration of TCS of the influent was 7.32
μg/L with a minimum value of 2.7 μg/L and a maximum value of 26.8 μg/L. Seasonal variation
was not found in the influent, although results may have been skewed because the samples had a
large amount of suspended solids. It is thought that TCS in WW influent will tend to adsorb to
solids (Thomas & Foster, 2005). The effluent average concentration was 0.11 μg/L with a
minimum value of 0.03 μg/L and a maximum concentration of 0.25 μg/L. Seasonal variation of
TCS was also measured (Figure 11). The WWTP before the CW removed an average of 98.3%
of TCS through its activated sludge technique. No effect was found on the use of UV to remove
TCS, although authors note that this finding had little data and studies were performed on low
concentrations of TCS which may give biased results. It is thought that this variation is due to
either less flow which increased the concentration, or differences in product utilization based on
seasonality. The wetland average concentrations were 0.09 μg/L for the influent, with a
minimum of 0.00 μg/L and a maximum of 0.29 μg/L. The average concentration was 0.04 μg/L
for wetland effluent. The minimum concentration was 0.00 μg/L and the maximum was 0.13
μg/L. The downstream average concentration was 0.12 μg/L, with a minimum of 0.03 μg/L and a
maximum of 0.29 μg/L. The downstream concentration was not statistically different from
WWTP effluent. This is different from other studies (Boyd et al., 2003; Federle et al., 2002). The
authors think that this is because other studies looked at streams that had lower WW effluent
concentration, and higher degradation, dilution and sorption in the streams. It was found that the
44
CW did significantly remove TCS from the WW effluent. This confirms CW use as a useful
technique to improve WW effluent quality before it is released into the environment (Waltman et
al., 2006).
Figure 11: Variations in TCS concentrations in the Denton Municipal WWTP effluent. Summer:
June to September, Fall: October to November, Winter: December to January, Spring: April to
May. Taken directly from Waltman et al., 2006.
Ávila et al. (2010) examined a pilot small scale horizontal subsurface flow CW in its
ability to improve general water quality, remove BPA and various PPCPS: diclofenac, ibuprofen,
naproxen, and tonalide. This CW treated sewer water as opposed to previously treated effluent.
Influent went through a course screen before entering the CW. First it entered a hydrolytic
upflow sludge bed reactor (HUSBR). Then it was inserted into two distribution tanks after which
an electrovalve was used to bring it to the CWs. The CWs were made in three different sections:
B1, B2, and B3. The first two sections had a surface area of 0.65 m2 and the water passed in one
of the two sections before moving into B3 which had a surface area of 1.65 m2. The treated water
then entered a holding tank where it was monitored. The CW sections were filled with 30 cm of
gravel and planted with Phragmites australis. The water depth was kept at 25 cm with a flow of
84 L/day and a theoretical hydraulic retention time of 3.5 days. Pollutant removal was tested by
continuous injection. The process was monitored over a period of 22 days.
BPA was injected into the system at 1 mg/L. It was found that after HUSBR remediation
BPA was reduced to a concentration of 1.5 μg/L.CW B1 reduced it to 0.3 μg/L and B2 reduced it
45
to 0.2 μg/L. B3 was the final reduction down to 0.05 μg/L. All chemicals had a reduction
between 97 and 99%. Carboxy-BPA was found as a metabolite after B3, which suggests that
aerobic remediation was effective in BPA removal. Figure 12 shows how the concentration of
BPA found in the CWs as a fraction of original concentration of HUSBR from day 13 to day 22.
The final concentration decreases after all treatment (Ávila et al., 2010).
Figure 12: Concentration of BPA found in the CWs as a fraction of original concentration of
HUSBR. Black circles are HUSBR, white circles are B1, black triangles are B2, and white
triangles are B3. Taken directly from Ávila et al., 2010.
Overall, it was found that the majority of remediation happened during HUSBR. This
coincides with the removal of particulate matter and conditions are anaerobic. However, authors
believe that the mix of aerobic and anaerobic conditions contributed to such high rates of
degradation. Authors suggest further studies examining metabolites that form during horizontal
subsurface flow CW (Ávila et al., 2010).
A benefit of constructed wetlands is that they can be used for commercial purposes.
Zurita et al. (2009) tested the use of horizontal and vertical subsurface-flow CWs (HFCW and
VFCW respectively) treating domestic wastewater in their ability to produce different
commercial flower species. HFCWs had continuous flooding and VFCWs had periodic flooding.
The four different species used were Agapanthus africanus, Anturium andreanum, Strelitzia
reginae, Zantedeschia aethiopica. Two HFCWs and two VFCWs were studies. One HFCW and
VFCW had just Z. aethiopica growing, and the others had the other three plants. This was done
because Z. aethiopica already had studies examining its remediation potential. It was found that
46
VFCWs had better removal for more pollutants: ammonium, biochemical oxygen demand,
chemical oxygen demand, organic nitrogen, total coliform, and total phosphorus. However, the
HFCWs were better at removing nitrate and total suspended solids. A. africanus and S. reginae
grew better in VFCWs. Z. aethiopica grew better in HFCWs. All of the plants except A.
andreanum survived 12 months. Vegetation used affected pollution remediation with BOD,
chemical oxygen demand (COD), TSS, and TP, with more plant species being more effective. A
variety of vegetation is thought to be beneficial because a variety of habitats are available at the
root level so different bacteria could be an active part of the system. Zurita et al. supports the
possibility of using CWs for economic as well as environmental reasons.
Belmont & Metcalf (2003) did lab studies examining the ability of Z. aethiopica to
remediate wastewater in small scale subsurface flow wetlands. Parameters measured were
ammonium, COD, nitrate and NPE, a precursor of NP. Remediation was compared to subsurface
flow units without plants. In these tests it was found that Z. aethiopica helped reduce nitrogen
levels. However, the plant did not help reduce COD or NPE in comparison to the non-vegetated
units. It was found that NP increased proportionally to NPE, likely because the former is a
metabolite of the latter. The plant’s lack of NPE reduction makes Z. aethiopica an unlikely
candidate for phenolic EDC remediation. However, it is possible that other ornamental plants are
good candidates for EDC remediation.
Giraudi et al. (2001) examined fungi found in pilot CWs and their ability to remove
anthracene (AA) and flourathene (FA). Two CWs were constructed and studied, one
contaminated with PAHs and one control. Both remediated domestic WW and were run for 6
months. 40 species were found in both CWs cumulatively. All 40 fungal species in the
contaminated CW where the control had 21 species. They were tested to find their AA and FA
remediation capabilities. Two species eliminated AA by over 70%. 33 species were able to
remove more than 70% of FA. Strains from the contaminated CW were better at reducing toxins.
New species were identified during this study.
This supports the idea of using fungi in CWs. It is possible to adapt species strains to be
better at removing certain toxins. However, some species will do better than others, and these
should be identified for efficient remediation cultivation. Fungi may work with plants and
bacteria during their remediation processes. It would be ideal to have a CW with bacteria, fungi,
and plants that assist each other in removing toxins (Giraud et al., 2001).
47
CWs work because of the living organisms’ biological processes which break down
toxins. However, using biota increases the maintenance required for treatment. Annuals used in
the system (Imai et al., 2007) would have to be germinated, or bought, and replanted every year.
Perennials would also have to be replaced when organisms expire (Zurita et al., 2009).
Furthermore, for species that accumulate EDCs (Takahashi et al., 2005), it may be necessary to
remove plants that have reached the limit of accumulation, or are accumulating toxins too
slowly. The waste this creates would have to be disposed of in a proper manner. Furthermore,
general water treatment maintenance would be required.
5. Conclusion and Recommendations
Based on the studies analyzed, it is recommended that the EBMUD WWTP located in
Oakland, California, use mycoremediation in the form of CLEAs.
5.1 Summary of findings
It is recommended that EBMUD WWTP use CLEAs because they provide greater decay
efficiency, higher decay speed, reusability, and require lower maintenance compared to CWs. As
shown in Figure 13, decay of pollutants by CLEAs is faster and more complete than by
phytoremediation.
Any WW treatment method has to be usable for as long as the WWTP is in operation. In
principle, CLEAs can be used repeatedly. Over time CLEA enzymes will become biologically
inactive. However, enzyme activity can be restored by running a buffer through the system. It
was shown in a lab study that enzyme activity could be fully recovered (Cabana et al., 2009).
Buffer waste would have to be managed in a responsible manner. CWs are also reusable when
maintained properly. Annuals have to be replaced every year (Zurita et al., 2009). Furthermore,
plants used to accumulate toxins or their metabolites will need to be replaced periodically as well
(Takahashi et al., 2005). Therefore, the maintenance of CWs would require more time and effort
than that of CLEAs because of the amount of replanting required for a large tank. While both
treatments are reusable, the CLEA method is preferable because it requires less maintenance.
As mentioned above, mycoremediation of WWTP effluent is promising technology,
however, no known pilot studies exist in literature. This constitutes a barrier to entry for facilities
interested in mycoremediation because without examples of past treatment design or treatment
outcomes, unexpected problems may be encountered during construction and operation.
Furthermore, while results can be modeled, in actuality they may be significantly different, as
48
reported in Giger et al. (2009). Facilities would rather invest in technologies that were already
shown to have high remediation capabilities. Thus, conducting pilot studies on the efficacy of
WW mycoremediation are essential to increase knowledge in the field and promote adoption of
this promising technology.
There are multiple pilot CWs in use today to treat wastewater as discussed in section
4.5.2. Studies conducted on these systems report the details of construction, the success of
operation, and list any problems encountered. Facilities looking to invest in EDC removal
technology may be likely to use this method due to the successes reported in literature (Waltman
et al., 2006; Park et al., 2009).
Figure 13. Efficiency comparison of phytoremediation and mycoremediation of three EDCs. A
conservative constituent is included to show no degradation. Constructed with HEC-RAS.
5.2 Limitations and future work
Several assumptions in this study require further verification. One significant issue is that
WRF and LMEs have been tested to remove EDCs in much higher concentrations than those
found in the environment (Cabana et al., 2007 (1)). Environmental concentrations of these
chemicals tend to be in μg/L (Jackson & Sutton, 2008). However, tests are often done with
concentrations in the mg/L range (Cabana et al., 2009; Reinhold et al., 2010; Imai et al., 2007). It
is not clear if techniques that remediate higher concentrations will be as successful at removing
lower concentrations (Cabana et al., 2007 (1)). In Cabana et al. (2009) a significant portion of
49
each EDC that was eliminated was adsorbed to the support. These EDCs may be released from
the support when the buffer is used to recover enzymatic activity, thereby contaminating the
buffer waste. As stated above, there are multiple types of NPs. It was assumed that all NPs would
be enzymatically degraded with similar efficacy. This may not be true.
In addition, it would be useful to have more complete answers to the following questions:
What are the environmental effects of BPA, NP, and TCS?
Where in the SFBA are these chemicals found in the ecosystem and geographically?
How much NP is in wastewater in the SFBA?
What are all metabolites formed from degradation of BPA and NP by P. oleracea?
What are all metabolites formed from degradation of TCS by L. punctata and L. minor?
What are all metabolites formed from degradation of BPA, NP, and TCS by CLEAs?
What is the endocrine disrupting potential of all metabolites?
Are remediation products likely to be transformed back into EDCs in the environment?
Are there other immobilized enzymes, such as MnP, that would be more efficient at
remediation than laccase?
Would laccase from another fungal species be more efficient at remediation?
Are there other plants that would be better at degrading selected EDCs?
What other EDCs or pollutants can be remediated with this technique?
Would adding oxygen to the system make it more efficient?
Would adding nutrients to CWs make them more efficient? Would this cause nitrification
in the local area?
What enzymes are P. oleracea, L. punctata, and L. minor using to degrade pollutants?
In what organs are P. oleracea, L. punctata, and L. minor accumulating the toxins?
5.3 Practical implications
This paper focused on application of EDC treatment methods for the EBMUD WWTP in
Oakland. However, this information can be applied to many other facilities. Any WWTP within
or outside of the SFBA can use this information. Manufacturing, laundry, and hospital facilities
can also use these systems to reduce the pollution concentration in their WW. In Europe, the use
of phytoremediation is encouraged to clean WW (Schröder et al., 2007).
Lastly, while this study focused on the removal of three EDCs from aqueous solution,
both mycoremediation and phytoremediation can also be applied to pollutants in other media,
50
such as soil or sludge (Schröder et al., 2007; Molla & Fakhru’l-Razi, 2012; Sethunathan et al.,
2004).
51
6. Appendices
6.1 Appendix 1: EDC removal by fungi and their enzymes from aqueous solutions (Adapted from
Cabana et al., 2007 (1))
EDC FungiTreatment method
EDC concen.
Rate of EDC removal
Rate of estrogenic activity removal
BPACoriolopsis polyzona Free laccase 5 mg/L
40% after 1 hour, 100% after 4 hours
35% after 1 hour, 95% after 4 hours
BPACoriolopsis polyzona
Immobilized laccase in CLEA form 5 mg/L
50% after 40 min, 90% after 150 min
BPATrametes versicolor Confined laccase
456 mg/L
50-90% after 24 hours, 100% after 96 hours
BPA
Trametes versicolor in Nicotiana tabacum Phytoremediation
22.8 mg/L
90-275 μM/gram of plant/2 months Less than start
BPAHeterobasidium insulare Fungi
200 mg/L
77% after 3 days, 100% after 14 days 100% after 1 day
BPAPhanerochaete chrysosporium Free MnP 50 mg/L
90% after 30 min, 100% after 60 min
40% after 4 hours, 90% after 6 hours (0.88mM)
BPAPleurotus ostreatus Fungi 91 mg/L
80% after 12 days, 85% after 21 days
BPAPleurotus ostreatus Free MnP 91 mg/L 100% after 1 hour
BPA Russula delica Fungi200 mg/L
68% after 3 days, 100% after 14 days
40% after 1 day, 100% after 3 days
BPA Trametes sp.Laccase and activated sludge
5-100 mg/L Varies Varies
BPA Trametes sp.Immobilized laccase
23 – 685 mg/L
134 μg of BPA after 30 min
BPATrametes versicolor Free laccase 50 mg/L
50% after 30 min, 70% after 60 min
40% after 1 hour, 60% after 6 hours (0.88 mM)
BPATrametes versicolor Free laccase
27.4 mg/L Varies Varies
BPATrametes versicolor
Immobilized laccase
0-1.1 g/L Varies Varies
52
BPA Trametes villosa Free laccase502 mg/L
100% after 3 hours
Completely removed
BPA Strain I-4 Free laccase 1.1 g/L
95% after 1 hour, 100% after 3 hours
All estrogenic activity was removed after 24 hours.
NP Bjerkandera sp. Fungi 45 mg/L95% after 5 days (9.7 mg/L d)
NPClavariopsis aquatica Fungi
22 – 55 mg/L
50% after 11 days, 60% after 26 days (250 μ solution, t-NP)
NPClavariopsis aquatica Free laccase
44.91-51.80 mg/L
22.1% of 4nNP after 24 hours, 14.0% of t-NP after 24 hours
NPCoriolopsis polyzona Free laccase 5 mg/L
80% after 1 hour, 100% after 4 hours
80% after 1 hours, 95% after 4 hours
NPCoriolopsis polyzona
Insolubilized as CLEA 5 mg/L
80% after 40 min, >95% after 60 min
NPCunninghamella sp. Fungi 11 mg/L
Half-life: 1 day (4nNP), 2 days (NP)
NP Fusarium sp. Fungi 11 mg/L
Half-life: 1-2 days (4nNP), >8 days (NP)
NP Mucor sp. Fungi 11 mg/L
Half-life: 1.5-2 days (4nNP), 3-5 days (NP)
NPPhanerochaete chrysosporium Fungi 11 mg/L
Half-life: 6 days (4nNP), 3 days (NP)
NPPhanerochaete chrysosporium Free MnP 51 mg/L
90% after 30 min, 95% after 60 min
60% after 1 hour, 80% after 5 hours (0.92mM)
NPTrametes hirsuta
Pilot-scale fungal/UF system 2.9 μg/L
>94% after 1.5 days
NPTrametes versicolor Fungi 45 mg/L
90% after 15 days (2.8 mg/L d)
NPTrametes versicolor Fungi 11 mg/L
Half-life: 1 day (4nNP), <1 day (NP)
NP Trametes Free laccase 51 mg/L 10% after 30 min, 10% after 4 hours,
53
versicolor 60% after 60 min60% after 9 hours (0.92 mM)
NPTrametes versicolor Free laccase 5 mg/L
90% after 5 min, 100% after 90 min (4nNP)
NP UHH 1-6-18-4 Fungi22 – 55 mg/L
75% after 11 days, 100% after 26 days (t-NP, 250 μM)
NP UHH 1-6-18-5 Free laccase
43.34-70.75 mg/L
46.2% of 4nNP after 24 hours, 63.5% of t-NP after 24 hours
NP Strain I-4 Free laccase 1.1 g/L
70% after 1 hour, 100% after 6 hours
TCSCoriolopsis polyzona Free laccase 5 mg/L
15% after 1 hour, 60% after 8 hours
TCSCoriolopsis polyzona
Insolubilized as CLEA 5 mg/L
80% after 40 min, >95% after 60 min
TCSPycnoporus cinnabarinus Fungi 72 mg/L
TCSTrametes versicolor Fungi 72 mg/L
60% after 1 week, 90% after 4 weeks
TCSTrametes versicolor Free laccase
5.8 mg/L
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6.2 Appendix 2: EDC removal by plants from aqueous solutions (Adapted from Takahashi et al., 2005; Loffredo et al., 2010; Shimoda et al., 2009; Imai et al., 2007; Nakajima et al., 2007; Gattullo et al., 2012(1); Iimura et al., 2007; Okuhata et al., 2010; Saiyood et al., 2013; Saiyood et al., 2010; Zhang et al., 2008; Gattullo et al., 2012 (2); Park et al., 2009; Reinhold et al., 2010; Chen et al., 2009)
EDC PlantTreatment method
EDC concentration Rate of EDC removal
BPARumex crispus japonicus Accumulation 40 mg/L
40% after 6 days, almost 100% after 15 days
BPA Agropyron fragileAccumulation and metabolism 4.6 mg/L 98% after 7 days
BPA Nicotiana tabacumImmobilized cell metabolism 28.8 mg/L 50% after 2 days
BPA Portulaca oleraceaAccumulation and metabolism 11 mg/L
>90% after 24 hours(90% after 24 hours)*
BPA Scenedesmus acutusAccumulation and metabolism 10 mg/L 50% after 4 days
BPAMonoraphidium braunii
Accumulation and metabolism 2 mg/L
19% after 2 days, 40% after 4 days
BPA
Populus seiboldii x Populus gradientata with Trametes versicolor expression
Accumulation and metabolism 29 mg/L
155 μM BPA /g dry root after 1 week
BPASalvia sclarea var. turkestanica
Mostly metabolism with some accumulation 11 mg/L
80% after 1 day, 98% after 2 days
BPABrugiera gymnorhiza
Accumulation and metabolism 20 mg/L
50% after 10 days, 100% after 51 days
BPADracaena sanderiana (sterile)
Accumulation and metabolism 4.6 mg/L 50% after 4 days
NPMyriophyllum verticillatum
Accumulation in lake Up to 26.4 μg/L Not applicable
NP Portulaca oleraceaAccumulation and metabolism 8.8 mg/L >90% after 24 hours
NP Raphanus sativus
Mostly metabolism with some accumulation 1 mg/L
73% after 1 day, 95% after 2 days
NP Loium perenne
Mostly metabolism with some accumulation 1 mg/L
71% after 1 day, 96% after 2 days
55
TCS Acorus sp.Constructed wetland around 20 ng/L
74% in May, -14% in August
TCS Typha sp.Constructed wetland around 20 ng/L
67% in May, 100% in August
TCSLandoltia punctata and Lemna minor
Accumulation and metabolism 2.9 mg/L
80% after 2 days, 90% after 6 days
TCSPhragmites australis
Accumulation and metabolism
1400 ng/g (sludge) 43% after 13 months
* Endocrine activity removal
6.2
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7. Acronyms
˙OH – Hydroxyl Radical
1-HBT – 1-hydroxy-benzo-triazole
2,4-DCP – 2,4-dichlorophenol
AA – Anthracene
ABTS – 2,2′-azino-bis-(3-ethylbenzthiazoline-6-sulfonic acid)
AC – Activated Carbon
AE(s) – Alkylphenol Ethoxylates
APs – Alkylphenols
BBP – Benzylbutylphthalate
BBzP – Butylbenzyl phthalate
BOD – Biochemical Oxygen Demand
BPA – Bisphenol A
BPAGlu – Bisphenol A Glycoside
BSA – Bovine Serum Albumin
BZP – Benzophenone
CLEAs – Cross-Linked Enzyme Aggregates
COD – Chemical Oxygen Demand
CTBA – Cetyl Trimetyl Ammonium Bromide
CW(s) – Constructed Wetland(s)
DBP – Di-n-butyl phthalate
DEHP – Di-2-ethylhexyl phthalate
DEP – Diethyl phthalate
DO – Dissolved Oxygen
DOC – Dissolved Organic Carbon
DOP – Di-n-octyl phthalate
EBMUD – East Bay Municipal Utilities District
ECC – Esterified Carboxyl Cotton
EDC(s) – Endocrine Disrupting Chemical(s)
EDCH – 1-ethyl-3-(3-dimethylaminopropyl) carbodiimide hydrochloride
FA – Flourathene
57
GAC – Granular Activated Carbon
GLU – Glutaraldehyde
GLY – Glyoxal
H˙ - Hydrogen Radical
H2O2 – Hydrogen Peroxide
HA(s) – Humic Acid(s)
HEC-RAS – Hydrologic Engineering Center - River Analysis System
HFCW – Horizontal subsurface-Flow Constructed Wetland
HM – Humic Material
HSs – Hydroponic Systems
HUSBR – Hydrolytic Upflow Sludge Bed Reactor
Kbiol – pseudo first order bio degradation constant
Klsw – Low Sludge Water distribution coefficient
Koc – Organic Carbon normalized coefficient
Kow – Octanol/Water partition coefficient
Lac – Laccase
LDH – Layered Double Hydroxide
LiP – Lignin Peroxidase
LME(s) – Lignin Modifying Enzymes
LMWOS – Low-Molecular Weight Oxidizable Substances
MnP – Manganese-dependent Peroxidase
MnSO4 – Manganese Sulfate
NP – Nonylphenol
NPEs – Nonylphenol Ethoxylates
O3 – Ozone
OP – Octylphenol
PAC – Powdered Activated Carbon
PAH(s) – Polycyclic Aromatic Hydrocarbons
PBR – Packed Bed Reactor
PCP(s) – Personal Care Product(s)
PDA – Potato Dextrose Agar
58
PEG – Polyethylene Glycol
PPCP(s) – Pharmaceuticals and Personal Care Product(s)
SFB – San Francisco Bay
SFBA – San Francisco Bay Area
STP – Sewage Treatment Plant
TCC – Triclocarban
TCEP – Tris(2-chloroethyl) phosphate
TCEP – Tris(2-chloroethyl)phosphate
TCS – Triclosan
TEMPO – 2,2,6,6-tetramethoxypiperidine 1-oxyl
TiO2 – Titanium Dioxide
USEPA – United States Environmental Protection Agency
UV – Ultraviolet Light
VFBs – Vertical Flow Beds
VFCW – Vertical subsurface-Flow Constructed Wetland
VLA – Violuric Acid
VP – Versatile Peroxidase
WRF – White Rot Fungi
WW – Wastewater
WWT – Wastewater Treatment
WWTP – Wastewater Treatment Plant
59
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