1. i
OLIVER JAMES GRIEVSON
THE REMOVAL OF PERMETHRIN AND TRIBUTYLTIN FROM
WASTEWATER USING
ADVANCED OXIDATION PROCESS AND ADSORPTION
SCHOOL OF APPLIED SCIENCE
MSc WATER & WASTEWATER TECHNOLOGY
3. iii
Abstract
Permethrin and tributyltin (TBT) are a persistent problem in the aquatic
environment with annual environmental quality standard failures reported in the
waterways and coastal waters of the United Kingdom. Their presence has been
reported in wastewater treatment plant effluents, methods of their removal are
required.
This paper examines the degradation of permethrin and TBT with three
advanced oxidation processes; UV-photolysis, UV/H2O2 and Fenton’s reagent.
The effectiveness and cost of each of the three processes is compared to
adsorption onto granular activated carbon (GAC).
UV/H2O2 was able to reduce permethrin and TBT simultaneously in water with
permethrin being removed to below the limit of detection and TBT removed to
89% of its initial concentration. UV-photolysis was able to remove both
permethrin in spiked deionised water but wasn’t as effective for TBT removal.
Fenton’s reagent wasn’t as effective as either of the photolysis methods for the
removal of the target compounds and adsorption, experimentally, wasn’t as
effective as it should have been. This will be expanded upon in the results
section.
The final conclusion, is that the cost of treatment was significant and for the
AOP’s predictions were £8.14 - £28.95 per m3
, in comparison to treatment by
adsorption to GAC ranging between £0.02 – £0.07 per m3
.
4. iv
Executive Summary
Introduction
The introduction to this project introduces the issues of permethrin and TBT as
toxic micro-pollutants, that they are ubiquitous within the aquatic environment in
the United Kingdom and the failures of environmental quality standards,
particularly TBT. The levels present in UK wastewater treatment plants are then
addressed and the applicability of advanced oxidation processes to their
removal from wastewater.
Literature review
The literature review looks at the chemical behaviour, toxicity and
environmental fate of both permethrin and TBT and looks at levels found within
the aquatic environment, in particular levels within wastewater treatment plants.
Advanced oxidation processes (AOP’s) are then addressed with the
mechanisms of how they work and then case studies of their use in the
degradation of organic compounds, with particular reference to permethrin and
TBT.
Methods and Materials
The method and materials section of this report details the degradation
experiments undertaken and gives details of the ethylation and novel analytical
procedure for the simultaneous analysis of permethrin and TBT by gas
chromatography-mass spectrophotometry.
5. v
Results and discussion
The results and discussion section firstly reviews the method development used
for the determination of permethrin and TBT. Then it goes onto look at the
results of the degradation experiments in deionised and wastewater.
The discussion goes on to look at the mechanisms of the degradation of
permethrin, TBT and other compounds that are degraded by similar
mechanisms. The photolability of permethrin and the lack of photolability of TBT
is discussed. The degradation of TBT using the hydroxyl radical produced by
UV/H2O2, Fenton’s reagent and coincidentally by UV/Fe(III) is discussed.
Finally the costs of treating permethrin and TBT by AOP’s are looked at and
compared to the cost of treatment by adsorption.
Conclusion
The main conclusions that this study comes to is that amongst the AOP’s used
UV/H2O2 is the most effective technique for the simultaneous degradation of
permethrin and TBT but that its is an economically unfeasible technique when
compared to adsorption onto GAC. It also concludes that further work is needed
into the degradation products formed and whether different AOP’s maybe more
effective and cheaper to use. Another area of further study is in to whether other
types of adsorbents maybe more effective at removing the target compounds
and how applicable this work is to un-spiked samples in treatment plant scale
tertiary wastewaters.
7. vii
4.3.1 Method development....................................................................... 37
4.3.2 Control degradation experiments .................................................... 38
4.3.3 Direct & indirect photolysis .............................................................. 39
4.3.4 Fenton’s reagent degradation experiments..................................... 41
4.3.5 Adsorption experiments................................................................... 42
4.3.6 Rate constants, quantum yields & energy consumption................. 42
4.4. Discussion.............................................................................. 44
4.4.1 Mechanisms of degradation ............................................................ 44
4.4.2 Costs of treatment........................................................................... 47
4.5. Conclusion ............................................................................. 48
4.6 Acknowledgements .................................................................. 49
4.7. References ............................................................................. 50
4.8 Tables ..................................................................................... 55
4.9 Figures.................................................................................... 60
5.0 References........................................................................... 65
Appendix A : Methodology....................................................... 71
A 1.1 Performance characteristics of the method.............................. 72
A 1.2 Principle ........................................................................................ 73
A 1.3 Reagents ....................................................................................... 74
A 1.4 Apparatus...................................................................................... 76
A 1.5 Procedure...................................................................................... 77
A 1.5.1 Degradation. ........................................................................... 77
A 1.5.2 Ethylation................................................................................ 82
A 1.5.3 Instrumental analysis .............................................................. 83
A 1.6 Calculation .................................................................................... 84
8. viii
Appendix B: Uridine Actinometry............................................ 85
B 1.0 Aim...................................................................................... 86
B 1.1 Methodology ................................................................................. 86
B 1.1.1 Principle.................................................................................. 86
B 1.1.2 Equipment............................................................................... 86
AP 2.1.3 Reagents.............................................................................. 87
B 1.1.4 Procedure ............................................................................... 88
B 1.1.5 Calculations ............................................................................ 89
B 1.2 Results........................................................................................... 90
B 1.3 Conclusions .................................................................................. 94
Appendix C: Raw data .............................................................. 95
Appendix D: Water Research Guide for authors .................. 103
9. ix
List of tables
Table 2.1: Acute toxicity of permethrin to fish species...........................................6
Table 2.2: Oxidation potential of common species...............................................13
Table 2.3: Experimental conditions of UV degradation case studies. ...............16
Table 2.4: Experimental conditions of UV/H2O2 degradation case studies......19
Table 2.5: Experimental conditions of Fenton’s reagent degradation case
studies. ..........................................................................................22
Table 2.6: Experimental conditions of adsorption case studies. ........................24
Table 4.1: Experimental conditions for spiked deionised degradation
experiments..............................................................................................55
Table 4.2: Experimental conditions for spiked wastewater degradation
experiments..............................................................................................55
Table 4.3: Table of quantum yield, rate constants & EEO values for spiked
deionised water experiments...........................................................56-57
Table 4.4: Table of quantum yield, rate constants & EEO values for spiked
wastewater experiments ........................................................................58
Table 4.5: Quantum yields, rate constants and EEO values of selected
compounds...............................................................................................59
10. x
List of Figures
Figure 1.1 Pesticide environmental quality standard failures in 2005.................1
Figure 2.1: Chemical structure & environmental behaviour of permethrin ..... …5
Figure 2.2: Chemical structure & environmental behaviour of tributyltin........... 8
Figure 2.3: Mass fluxes of total organotin through a wastewater treatment plant
……………………………………………………………………………11
Figure 4.1: Example chromatogram from the GC-MS...........................................60
Figure 4.2: Permethrin and TBT reduction by UV and UV/H2O2 in spiked
deionised water ........................................................... ……………61
Figure 4.3: Permethrin and TBT reduction by UV and UV/H2O2 in spiked
wastewater ...............................................................................................62
Figure 4.4: Permethrin & TBT reduction by Fenton’s reagent in deionised
water..........................................................................................................63
Figure 4.5: Permethrin and TBT reduction by adsorption onto Norit GAC 1240
in spiked wastewater. .............................................................................64
11. xi
Nomenclature
Φ quantum yield
amu atomic mass unit
AOP’s advanced oxidation processes
COD chemical oxygen demand
d20 density at 20ºC
EQS environmental quality standards
g/d grams per day
GAC granular activated carbon
GC-ECD gas chromatography electron capture detector
GC-FPD gas chromatography flame photometric detector
GC-MS gas chromatography mass spectrometer
hv photon’s energy
Koc organic carbon absorption coefficient
Kow octanol water partition coefficient
kWh kilowatt hour
LC50 lethal concentration causing the death of 50% of the population
LD50 lethal dose causing the death of 50% of the population
PAC powdered activated carbon
TBT tributyltin
TCE trichloroethylene
UV ultraviolet
UV/H2O2 ultraviolet photolysis/ hydrogen peroxide advanced oxidation
process
W m-2
watts per square metre
12. xii
Acknowledgements
I would like to thank my supervisor Professor Simon Parsons for all of his efforts
and help in this project, especially for all of his reading of draft after draft and for
all of his editing skills and feedback.
I would also like to acknowledge the support of United Utilities in the funding of
this project as part of the STAMP scheme.
13. Page 1 of 109
1.0 Introduction
Tributyltin (TBT) and permethrin can be classed as two of many micro-
pollutants present in UK waters that cause problems in the natural environment.
The toxicity of TBT has been well documented especially its properties as an
endocrine disruptor and its previous impacts on the oyster industry (Alzieu,
1991). Permethrin is as well documented and its effects on aquatic eco-systems
can include invertebrate and fish deaths (Kamrin, 1997).
The prevalence of both of these compounds meant that between 2003-2005
there were a total of 53 failures of environmental quality standards (EQS) in
freshwaters from TBT and 17 failures in freshwaters from a combination of
cyfluthrin and permethrin (Environment Agency, 2007). The following figure
shows that these failures in 2005 were distributed throughout the United
Kingdom.
Figure 1.1 Pesticide environmental quality standard failures in 2005
(Environment Agency, 2007).
14. Page 2 of 109
TBT has been legislated since 1982 as a result of damage to the oyster
industry, with a standard set at 20 ng L-1
(Alzieu, 1991) and later in the United
Kingdom in 1986 (Clark et al, 1998). Current proposed legislation as part of the
Water Framework directive in the proposed directive on environmental quality
standards in the field of water policy is to see the maximum allowable
concentration of TBT set to fall to 1.5 ng/L and an average annual concentration
of 0.2 ng L-1
(Commission of the European Communities, 2006).
Environmental quality standards for permethrin are set at a local level by the UK
Environment Agency and permethrin is classed as a dangerous substance
under the Dangerous Substance Regulations (Environment Agency, 2007).
Permethrin and TBT have been measured in wastewater treatment plants at
concentrations up to 81 g L-1
for permethrin (Kupper et al, 2006) and 0.22
g L-1
for TBT (Fent & Muller, 1991) their removal from the wastewater
treatment process is imperative in order to comply with the proposed UK and
EU legislation.
Whilst a number of major studies have reported that the major route of both
permethrin and TBT within the wastewater environment is to be adsorbed to the
solids and get captured with the sludge (Fent, 1996; Plagellat, 2004) however
this will not remove either compound in the liquid phase to below the European
limit of 1.5 ng L-1
(Schafran, 2003).
As a result of this, it is necessary to look at treatment methods for the removal
of both permethrin and TBT from wastewater. Due to both compounds high
affinity to be adsorbed, the use of granular activated carbon (GAC) is a
possibility and has been studied previously (Schafran, 2003) and found to be
effective but not enough to remove TBT to below the regulated concentration of
50 ng L-1
, in the shipyard waters that were studied. Due to the GAC’s inability to
15. Page 3 of 109
reduce TBT to below the legislative limit Schafran also investigated the use of
UV/H2O2 advanced oxidation process and found it to be an effective process in
the removal of TBT from shipyard waters.
The aim of most advanced oxidation processes (AOP’s) is to produce the
hydroxyl radical in order to mineralise organic compounds to less toxic or ideally
harmless inorganic molecules (Parsons, 2004). In this study the advanced
oxidation process’s that are being used are UV photolysis in the UV-C
waveband at 254nm, a combination of UV photolysis and hydrogen peroxide
and finally Fenton’s reagent (a mixture of ferrous iron and hydrogen peroxide).
Their removal from tertiary wastewater streams either by degradation or by
adsorption onto GAC will be examined in terms of the effectiveness of the
techniques and their applicability to the wastewater industry.
16. Page 4 of 109
2.0 Literature review
This literature review will look at the two target compounds of this study,
permethrin and TBT, their uses, their environmental fate and their toxicity.
Finally, any relevant studies to their occurrence within wastewater treatment
plants and their environmental fate will be reviewed. The review will then look at
the four main removal methods that are going to be used in this study, UV-
photolysis, UV/Hydrogen peroxide, Fenton’s reagent and adsorption. The
mechanisms of the methods will be explained, and then studies of their use in
removing the target compounds reviewed.
2.1 Target compounds
In this study the two main compounds that are going to be reviewed are
permethrin and TBT.
2.1.1 Permethrin
Permethrin (or to use its chemical name: 3-phenoxybenzyl (1RS)-cis, trans -3-
(2,2-dichlorovinyl)–2,2-dimethylcyclpropanecarboxylate) is a synthetic
pyrethroid insecticide used in a wide variety of sheep dips (Cooke et al, 2004),
and on agricultural crops to control biting flies, cockroaches and ectoparasites
(Kamrin, 1997). Details of permethrin’s chemical structure and its behaviour in
the environment are contained in Figure 2.1.
17. Page 5 of 109
It is practically non-toxic to mammalian life with reported LD50 in rats through the
oral route of 430-4500 mg kg-1
. In
humans, permethrin is rapidly
metabolised and excreted and does
not significantly persist in the human
body (Kamrin, 1997).Permethrin
does have a significant impact on
aquatic ecosystems as it destroys
both the quality and quantity of
insects and invertebrates with LC50
concentrations of less than 1 g L-1
being fatal (Kamrin, 1997). It is also
highly toxic to most fish especially at
lower temperatures and especially for
smaller fish (Sánchez-Fortún &
Barahona, 2005). LC50
concentrations for aquatic organisms range from 0.075 g L-1
for Daphnia
Magna to 9.8 g L-1
for Rainbow trout (Imgrund, 2003). There are a variety of
toxicities of permethrin depending on the species of fish, some LC50 values are
in Table 2.1.
Figure 2.1: Permethrin
Chemical Formula: C21H20Cl2O3
CAS No: 52645-53-1
Molecular Weight: 391.288
Melting point (ºC): 34
Boiling point (ºC): 200
Density (g cm3
) 1.23
Solubility in water (mg l-1
) 0.2
KOC (mL g-1
) 100,000
Kow (Log P) 6.5
(Lide, 2005)
(Lee et al, 2002)
18. Page 6 of 109
Table 2.1: Acute toxicity of permethrin to fish species (Baser et al., 2003)
Species Duration of test (hrs) LC50 ( g L-1
)
Salmo salar 96 12
Oryzias lapites 48 41
Micropterus salmoides 96 8.5
Salvelinus fautinalis 96 3.2
Leopomis macrochirus 96 5
Cyprinodon variegates 96 7.8
Stripped bass 96 16.1
Menidia beryllina 96 0.062
Paleomonetes pugio 48 0.049
Tilapia zillii 48 49
Due to its high KOC value (Figure 2.1) the environmental fate of permethrin is to
be tightly bound in the soil or within a wastewater environment to the sewage
sludge. This is especially the case when the soils or sludge contain a high
organic matter content (Kamrin, 1997). Due to its low solubility in water and its
affinity for organic carbon, once it is bound in the sludge it is not very mobile
and will be readily broken down by micro-organisms (Kamrin, 1997). Permethrin
has a reported average half-life in sludge of between 30 & 38 days (Kamrin
1997 & Imgrund 2003).
Plagellat (2004) reported removal of permethrin from the final effluent of
wastewater treatment plants at greater than 94%. This identified that removal by
adsorption to the sewage sludge within the wastewater treatment process was
very dependant on the type of treatment process used but the adsorption
ranged from 4-15% of the incoming load. The rest of the removal (i.e. 85-96% is
not explained. This study also identified that private households were a
significant source of permethrin and TBT.
19. Page 7 of 109
Kupper et al (2006) reported influent concentrations of permethrin into a
conventional activated sludge wastewater treatment process up to 81 g L-1
. In
Kupper’s study water was spiked up to 544 ng L-1
, permethrin concentrations
after primary treatment were reduced by 20% and by the end of the wastewater
treatment process there had been a decrease in permethrin concentrations of
92% (to a concentration of 20 ng L-1
). The main removal mechanisms that were
identified were the adsorption of permethrin onto the sewage sludge and
biodegradation. The removal of 20% in the primary treatment stage was in a
similar range to Plagellat (2004) who reported a 15% removal due to adsorption
onto the sewage sludge in the primary treatment stage (Plagellat, 2004).
Contrary to the figure of 15-20% removal in the primary treatment process
(Plagellat, 2004, Kupper et al, 2006) Kirk et al (1989) reported a removal of up
to 61% of permethrin in batch studies on the laboratory scale, but with an
incoming concentration of 50 g L-1
. However once Kirk et al (1989) worked on
a larger scale with continuous flow activated sludge simulation the removal of
permethrin by adsorption was 10-30%. The study by Kirk et al (1989) also
looked at the degradation of permethrin during the sludge treatment processes
and found removal of permethrin from the sewage sludge over a 32 day
anaerobic digestion period could reach up to 96%.
Rogers et al (1989) study on the occurrence of permethrin in twelve UK sewage
sludges showed concentrations of permethrin up to 40.8 mg kg-1
(dry weight).
This shows “that permethrin is sorbed onto sludge solids during sewage
treatment” (Rogers et al. 1989).
20. Page 8 of 109
2.1.2 Tributyltin
TBT is a sub-group of the trialkyl organotin compounds (Etoxnet, 2007). TBT
was first used in 1959 as an anti-fouling additive in marine paints (Clark et al,
1988) until the practise was
stopped in January 2003 (Song et
al., 2005). TBT is also used in wood
treatment and preservation, as an
anti-fungicide in the textile industry,
in wood pulp and paper mill
systems, in breweries (Etoxnet,
2007), and as stabilizing additives
in poly-vinyl chloride and other
polymers (Hoch & Schweisig, 2004)
Details on TBT’s chemical structure
and behaviour in the environment
are contained in Figure 2.2.
TBT is moderately toxic via oral
ingestion and through the skin. Oral
LD50 values range between 55-87
mg kg-1
in mice & rats. In humans a strong irritating effect has been recorded
and if concentrations are high enough, irritated skin, dizziness and flu-like
symptoms (Etoxnet, 2007).
Where TBT is particularly toxic is in the aquatic environment. It has been known
to cause defective shell growth in oyster populations, and cause the
development of male genitalia in the female dog whelk, and cause
immunosupression in fish populations (Yebra et al, 2004). This was extensively
studied in two papers by Alzieu (1991 & 1998) where problems with oyster
growth, specifically Crassostrea gigas in the Bay of Archaron, prompted the first
Figure 2.2: TBT
Sn
H
Chemical Formula: C12H28Sn
CAS No: 688-73-3
Molecular Weight: 291.060
Melting point (ºC): 76
Boiling point (ºC): 113
Density (g cm3
): 1.103
Solubility in water (mg l-1
): 4
KOC (mL g-1
): 1600
Kow (Log P): 4.1
(Lide, 2005)
(Weidenhaupt, 1997)
21. Page 9 of 109
regulations on TBT in the aquatic environment. In 1982 the French government
initially passed legislation banning the use of TBTs on all boats under 25
tonnes. This was then extended to cover all boats under 25m length and fishing
apparatus.
In Alzieu’s study in 1998 it confirmed that the actions of the French government
had resulted in a decrease in TBT concentrations. In 1986 the UK government
instigated a safe water concentration of 20ng L-1
(Clark et al., 1988). In 1987 the
state of Virginia in the United States imposed a limit of 50 ng L-1
of TBT
(Messing et al, 1997 ; Prasad & Schafran, 2006). More recent draft proposals
from EU legislation has seen this maximum allowable concentration fall to 1.5
ng L-1
and an average annual value of 0.2 ng L-1
(Commission of the European
Communities, 2006).
Similarly to permethrin, the environmental fate of TBT is to be adsorbed on to
sedimentary and particulate material, however it also has a tendency to de-sorb
when degraded to either dibutyltin or monobutyltin as they have less of an
adsorption capacity to particulates (Hoch & Schweisg, 2004). Due to this
capability of adsorption and de-sorption from sediments, TBT is ubiquitous
within the aquatic environment (Clark et al, 1988).
Within the environment, TBT will breakdown by a number of breakdown
pathways, these include direct photolysis, biodegradation and differing types of
chemical degradation (Clark et al, 1988). Of these degradation mechanisms
photolysis utilising natural light has been shown to be a slow method of
degradation with a half life of greater than 89 days and thus is not considered to
be a significant route (Clark et al, 1988). Of the three breakdown mechanisms
biodegradation is seen as the most significant (Clark et al, 1988), where it
degrades to dibutyltin, which will further degrade to monobutyltin, and finally
inorganic tin.
22. Page 10 of 109
There have been a number of studies on the removal of TBT in wastewater
treatment and in industrial wastewaters (mainly to do with the treatment of
shipyard wastes). In addition to this other authors have looked at the fate of
TBT in the wastewater environment.
Schafran and Tekleab (2000) and Schafran (2002) looked at various methods of
removing TBT from shipyard waters. These methods included
coagulation/clarification at a range of different pH’s (using both aluminium and
ferric sulphates as a coagulant at pH 6, 8, 10 and at doses ranging from 41-164
mg L-1
Al2(SO4)3 and 60-240 mg L-1
Fe2(SO4)3), granular media filtration,
granular activated carbon filtration (Calgon F400), and with UV/hydrogen
peroxide (No information on UV lamp intensity, and hydrogen peroxide
concentrations between 0 – 200 mg L-1
). This study concluded that granular
media filtration had a poor affinity for removing TBT. The next least effective
method was coagulation/clarification with approximate removal rate of 45%.
Both granular activated carbon and UV/Hydrogen peroxide were assessed to be
effective methods for the removal of TBT from the shipyard wastewaters. Other
technologies for removal of TBT from industrial wastewaters include thermal
treatment at 1000ºC (Song et al., 2005) and sorption onto dolomitic sorbents
(Walker et al., 2005).
In wastewater treatment plants there have been a large amount of studies into
the occurrence of TBT, and its environmental fate and treatment. The
concentrations of TBT in the sewage treatment process influent range from
below the limit of detection of 1.8 ng L-1
(Donard et al., 1993) to 220 ng L-1
(Fent
& Mϋller, 1991; Fent 1996). A study by Fent in 1996 on organotin in municipal
wastewater in Switzerland showed 73% removal of organotins (including TBT)
in the primary effluent, this rose to 90% removal in the secondary effluent and
98% removal in the tertiary effluent. The major removal mechanism in this case
was adsorption onto the suspended solids in the wastewater treatment process.
23. Page 11 of 109
In this study Fent managed to map the mass balance of organotins throughout
the wastewater treatment plant (Figure 2.3):
2.5%
PC
AD
AS SC F
10%
90%
45%
48%
5%
22%
0.5%
5%
5%
Dissolved
Particulate
Key
122g/d = 100%
PC Primary Clarifier
AS Activated Sludge
SC Secondary Clarifier
F Filter
AD Anaerobic Digestion
Excess sludge 9.5%
Figure 2.3: Mass fluxes of total organotin through a wastewater treatment plant
(From Fent, 1996).
Fent & Muller (1991) had earlier reported this removal percentage. In this earlier
study 92% of organotin was associated with the suspended solid fraction of the
wastewater treatment plant and this fraction was removed with the sewage
sludge. The incoming concentration of TBT ranged from 64-217ng L-1
. Fent &
Muller concluded that adsorption into the sludge was the most important
process for organotin removal in sewage treatment. Biodegradation in the
activated sludge process only accounted for a 7.5% removal of the organotins.
Plagellat et al (2004) reported TBT levels in sewage sludge up to a maximum of
648.5 g kg-1
dry weight confirming that TBT will preferentially associate itself
with the suspended solids within a wastewater treatment process.
24. Page 12 of 109
Once in the sludge fraction of sewage treatment a study by Fent et al., (1991)
showed that TBT was not significantly degraded. Whilst, Stasinakis et al. (2005)
reported removal of 99.7% of TBT with a degradation time of 10 days in
activated sludge and a sludge spiked at a concentration of 100 g L-1
as Sn in
laboratory scale activated sludge batch reactors
2.2 Advanced Oxidation Processes & Adsorption
This study will look at four methods of removing permethrin and TBT from
wastewater. Three of these methods are advanced oxidation processes (AOP’s)
and are UV-Photolysis, UV-Photolysis combined with hydrogen peroxide and
hydrogen peroxide combined with ferrous sulphate, also know as Fenton’s
reagent. The fourth method of removal is adsorption onto granular activated
carbon (GAC).
UV-photolysis works on the principle of the absorption of UV light by a
compound in order to cleave the bonds of the compound. UV-
photolysis/hydrogen peroxide (UV/H2O2) and Fenton’s reagent both work on the
principle of generating the hydroxyl radical (HO•
). The hydroxyl radical is one of
the most reactive free radicals and one of the strongest oxidants (Huang et al.,
1993) with only fluorine being more reactive (see table 2.2)
25. Page 13 of 109
Table 2.2: Oxidation potential of common species (Parsons, 2004)
Species Oxidation potential (V)
Fluorine 3.03
Hydroxyl radical 2.80
Atomic oxygen 2.42
Ozone 2.07
Hydrogen peroxide 1.78
Perhydroxyl radical 1.70
Permanganate 1.68
Hypobromus acid 1.59
Chlorine dioxide 1.57
Hypochlorus acid 1.49
Chlorine 1.36
2.2.1 UV-Photolysis
The underlying principle of UV-Photolysis of a target contaminant such as
permethrin and TBT, as in this study, is related to Planck’s law of radiation and
the laws of photochemistry, insofar as a certain amount of energy is needed for
the photon energy to match the bond energy and cleave the bonds within a
compound (Parsons, 2004).
The effectiveness of UV-photolysis is governed by the first law of
photochemistry which states:
“that only the light that is absorbed by a molecule can be effective in producing
a photochemical change in that molecule” (Parsons, 2004).
26. Page 14 of 109
This is measured by the Beer-Lambert law “which states that the fraction of light
absorbed by the system does not depend on the incident spectral radiant power
and the amount of light absorbed is proportional to the number of the
constituent molecules absorbing radiation” (Parsons, 2004). This gives us the
molar absorption coefficient of a pure compound at a given wavelength and
governs how much will be absorbed. This shows whether a compound will
absorb UV light and the energy required in order to cleave the compounds
bonds.
In this study the UV-C band range (between 200-280nm) specifically UV254 will
be used, as this is the wavelength where both the pollutants and the
constituents within the water absorb radiation (Parsons, 2004).
The most common reaction equations that a compound undergoes are below:
RX + hv → RX* (1)
RX* → (R•
…•
X) cage → R•
+ •
X (2)
(R•
…•
X)cage → RX (3)
RX* → (R+
…X-
) cage → R+
+ X-
(4)
RX* + O2 → RX+•
+ −
2O •
(5)
RX* + 3
2O → RX + 1
O2 (6)
(Parsons, 2004)
The excited state RX* is generated through light absorption processes in
equation 1, this is highly energetic and either deactivates to the ground state of
the molecule or undergoes “dark” chemical reactions, as in the equations above
(Parsons, 2004). The bond scission that occurs in equation 2 is the predominant
chemical pathway. Once the radicals have escaped from the solvent cage they
undergo further oxidation/reduction reactions depending upon the chemical
structure (Parsons, 2004).
27. Page 15 of 109
The effectiveness of UV-photolysis as a degradation technique for pesticides
has its limitations. In his review of pesticide chemical oxidation Chiron et al.
(2000), identified that pesticide degradation using an artificial light source
requires long treatment times of high energy photons and “rarely achieve a
complete degradation of the pollutant, ”with the exception of vacuum UV
(Chiron et al., 2000). However, in combination with other degradation
techniques, such as ozonation, Fenton’s reagent and hydrogen peroxide the
efficiency of these techniques can be greatly increased (Chiron et al., 2000).
Gogate & Pandit (2004) also identify that UV can be used in a photo-catalytic
oxidation to effectively degrade compounds in wastewater.
UV was used by Esplugas et al. (2002) to degrade phenol however a maximum
of 24% degradation was observed after 30 minutes of treatment at pH 4.4. This
decreased to 14% when the pH was increased to 6.8 and only 5% when the pH
was further increased to pH 11.5 due to a decreased quantum yield with
increasing pH. Esplugas et al. (2002) also concluded that the effectiveness of
UV increased when combined with another degradation technique such as the
use of hydrogen peroxide. The initial concentration of phenol ranged between
94 and 114 mg L-1
and the flux of radiation of the two ultraviolet lamps were
26.6 & 21.1 W m-2
.
Experiments by Beltran et al. (1993) degraded atrazine in laboratory
experiments. Atrazine degradation using only direct UV photolysis had a half life
of between 3.5 – 11.7 minutes. When hydrogen peroxide was added to the test
solution the half life decreased 1.2-2 minutes. In the experiments using UV only
the concentration of atrazine ranged between 2.37 – 23.7 mg L-1
and the flux of
radiation from the ultraviolet lamp was 0.46 W m-2
.
Fukui et al. (1991) managed to degrade up to 76% of 3-chloro-4-
(dichlormethyl)-5-hydroxy-2(5H)-furanone (MX) after a treatment time of 60
28. Page 16 of 109
minutes. This proves that the efficacy of UV-photolysis is very much dependant
upon the compound itself and the length of time that it is treated. In this study
only 1mL of 1mmol solution of MX was degraded under an 18W low pressure
mercury lamp at a path length of 30cm. Actinometric tests were not conducted
in order to determine the flux of radiation.
The table below summarises the experimental conditions and results of the
studies detailed:
Table 2.3: Experimental conditions of UV degradation case studies.
2.2.2 UV-Photolysis & hydrogen peroxide (UV/H2O2)
UV photolysis when combined with hydrogen peroxide has been studied in
conjunction with numerous industrial wastewater effluents especially the textile
industry, olive oil industry and the paper & pulp industries (Parsons, 2004).
The principle of UV/H2O2 is, as with all AOP’s, the production of the hydroxyl
radical. With this process the UV photolysis cleaves the hydrogen peroxide,
producing two hydroxyl radicals as in equation (7)
H2O2 + hv → OH•
+ OH•
(7)
(Parsons, 2004)
Study Target
compound
Initial
concentration
UV Power % degradation
Esplugas et
al. (2002)
Phenol 94-114 mg L
-1 26.6 & 21.1
W m
-2
5-24% in 30 minutes
(depending on pH)
Beltran et al.
(1993)
Atrazine 2.37 -23.7 mg L
-1
0.46 W m
-2
99% in < 15 minutes
Fukui et al. MX 1mL of 1mmol
18W at a
path length of
30cm
76% in 60 minutes
29. Page 17 of 109
The problem with UV/H2O2 is that the molar absorption coefficient is low. This
means that relatively high concentrations of hydrogen peroxide are required.
The disadvantage of this is that if the concentration is too high then hydrogen
peroxide scavenges the hydroxyl radical using the following reaction equations
(equations 8-10):
OH•
+ H2O2 → HO•
2 + H2O (8)
HO•
2 + H2O2 → OH•
+ H2O + O2 (9)
HO•
2 + HO•
2 → H2O2 + O2 (10)
(Parsons, 2004)
In addition to this any alkalinity in the wastewater in the form of carbonate or
bicarbonate ions, also act as scavengers of the hydroxyl radical under the
following reaction equations (equations 11-12):
OH•
+ HCO3
-
→ •
CO3
-
+ H2O (11)
OH•
+ CO −2
3 → •
CO3
-
+ OH-
(12)
The bicarbonate or carbonate ions react with part of the hydroxyl radicals to
form carbonate ion radicals, which although they do react with the organic
compounds, are much more selective and have lower rate constants (Parsons,
2004).
There have been a number of studies optimising the use of the UV/H2O2
technique on a variety of compounds and a variety of industrial and domestic
wastewaters as well as synthetic solutions on a laboratory scale.
30. Page 18 of 109
As mentioned earlier, a study by Beltran et al (1993) managed to degrade
atrazine with half life of between 1.2 – 2 minutes, apart from one sample where
hydrogen peroxide was added in excess and thus acted as a hydroxyl radical
scavenger (as in equations 8-10) and the half life increased to 7.2 minutes. In
total 99% of atrazine was removed. Concentrations of atrazine ranged between
3.8 x 10-5
and 9.6 x 10-5
mol L-1
. The concentration of hydrogen peroxide used
ranged between 0.6 and 110 mmol L-1
and the UV incident flux radiation
0.52 W m-2
.
Weir & Sundstrom (1993) looked at the degradation of trichloroethylene (TCE)
in phosphate buffer on a laboratory scale. This study looked at the kinetics of
TCE degradation and concluded that TCE reaction followed first order kinetics.
The UV intensity followed an apparent first order kinetic rate and the
concentration of hydrogen peroxide followed first order kinetics up to a
maximum level. The experimental conditions for these experiments were an
initial TCE concentration of 26.8 mg L-1
, a hydrogen peroxide concentration
between 0.2mM and 20mM and a UV intensity between 0.8 and 2.88 W m-2
. As
these experiments looked at the rate or reaction no details on how much TCE is
removed from solution is given.
A recent study by Yonar et al. (2006) applied UV/H2O2 to wastewater samples
and found over a 95% reduction in COD concentrations. This paper also
identified an approximate cost in terms of electrical energy per kg of COD of
10kWh. Experimental conditions for these experiments were an initial COD
concentration of 336 ± 25 mg O2 L-1
, a hydrogen peroxide concentration
between 0.74 and 2.94 mmol L-1
and a UV intensity between 3 – 8.8 W m-2
.
The table below summarises the experimental conditions and results of the
studies detailed:
31. Page 19 of 109
Table 2.4: Experimental conditions of UV/H2O2 degradation case studies.
2.2.3 Fenton’s reagent (Ferrous sulphate & hydrogen peroxide)
Fenton’s reagent generates the hydroxyl radical by a number of complex
chemical reactions, it was first used by Henry J Fenton in 1894 and further
developed by Haber & Weiss in 1934. It was not until 1949 that Barb et al.
proposed the following set of chemical reactions describing the “dark” reaction
of Fenton’s reagent (Pignatello et al., 2006):
Fe(II) + H2O2 → Fe(III) + OH-
+ HO•
(13)
Fe(III) + H2O2 → Fe(II) + HO•
2 + H+
(14)
HO•
+ H2O2 → HO•
2 + H2O (15)
HO•
+ Fe(II) → Fe(III) + OH-
(16)
Fe (III) + HO•
2 → Fe(II) + O2H+
(17)
Fe (II) + HO•
2 + H+
→ Fe(III) + H2O2 (18)
HO•
2 + HO•
2 → H2O2 + O2 (19)
Study Target
compound
Initial
concentration
H2O2
concentration
UV
Fluence
%
degradation
Beltran
et al
(1993)
Atrazine
8.2 – 20.7
mg L
-1
0.6–110
mmol L
-1
0.52
W m
-2 99%
Weir &
Sundstrom
(1993)
TCE 26.8 mg L
-1 0.2mM -20
mmol L
-1
0.8-2.8
W m
-2
No
information
Yonar et
al. (2006)
Wastewater 336 mg L
-1
O2
0.74 – 2.94
mmol L
-1
3 – 8.8
W m
-2 95%
32. Page 20 of 109
In these reactions iron cycles between the +II and +III oxidation states where in
the absence of other oxidizable substances it acts as a catalyst to convert the
hydrogen peroxide to oxygen and water (Pignatello et al., 2006). The hydroxyl
radical is stoichometrically produced in reaction 1, but this produces a
stoichometric amount of Fe(III) which later precipitates as ferric oxyhydroxide,
as pH is increased, creating an undesirable sludge. In reaction 2 the generation
of the hydroxyl radical is catalytic in iron (Pignatello et al., 2006). Reactions 3 &
4 show the scavenging of the hydroxyl radical by both the hydrogen peroxide
and the ferrous iron, but, as iron is used catalytically this scavenging is kept to a
minimum, as is the production of ferric oxyhydroxide sludge. Although reaction
2 minimises the sludge production, it is also a lot slower in producing the
hydroxyl radical than reaction 1 (Pignatello et al., 2006).
The pH at which the Fenton’s reagent operates is vital to its effectiveness, as
this effects the species of iron present in solution, and thus the rate at which the
reaction progresses. Depending upon the target compounds, an ideal range for
the Fenton’s reactions is between pH 3-4, due to the formation of Fe(OH)2
which is approximately 10 times more reactive that Fe(II) (Pignatello et al.,
2006). Other authors have looked at the pH range and found that it is very much
target compound dependant and that this pH range can be extended to
approximately pH 5.5 (Arnold et al., 1995). Above pH 5.5, the effectiveness of
Fenton’s reagent declines rapidly due to the speciation of iron (Arnold et al.,
1995) and other compounds such as the bicarbonate ion, which is known to be
a strong scavenger of the hydroxyl radical (Beltran et al., 1993).
The degradation of a number of compounds such as atrazine, chloro-
benzenes, chloro-phenols & organo-phosphorus compounds, as well as the
treatment of domestic and industrial wastewaters with Fenton’s reagent, have
been studied in depth by a number of authors, although no studies on the
degradation of TBT or permethrin appear to have been performed.
33. Page 21 of 109
Arnold et al, (1993), have discovered the optimal conditions for atrazine
degradation using Fenton’s reagent showing that a pH of 3 and a 1:1 ratio of
2.69mM of FeSO4:H2O2 allowed a degradation of 30.2 mg L-1
of atrazine in
under 30 seconds. Gallard & De Latt in 2000 on kinetically modelled Fenton’s
like reactions using atrazine as a target compound showed that below a pH of 3
the degradation rate follows pseudo-first order kinetics. Atrazine was spiked at
a concentration of 0.15 mg L-1
, a reaction pH of between 1 and 3 was used,
hydrogen peroxide concentrations between 0.2 mM and 1M and a ferric iron
concentration of 0.2mM. Gallard & De Latt in 2001 found that the kinetic
approach was much more complicated, with radical intermediates formed from
the decomposition of their target compounds reduced back to the parent
compound. This study used chlorobenzenes and phenyl-ureas as target
compounds at a concentration of 1 M, a reaction pH of 3, and an excess of
ferrous iron and hydrogen peroxide.
Badaway et al, (2006) compared the “dark” Fenton’s reactions (Fenton’s regent
without the addition of UV) with light enhanced Fenton’s reactions (i.e. the
addition of UV to Fenton’s reagent) and found that for organo-phosphorus
compounds the photo-assisted Fenton’s reactions were significantly more
efficient in wastewater, although the transitivity of the wastewater will have been
a factor on the efficiency of the degradation process. Initial concentrations of
organo-phosphorus pesticides were 50mg L-1
, Fe2+
concentrations were
0.089mM and hydrogen peroxide concentrations were 8.99mM, the effect of pH
was studied with reaction pH’s between 2 & 5. This gave a 70% degradation in
a treatment time of 90 minutes.
Badaway & Ali in 2006 also studied the use of Fenton’s reagent in conjunction
with coagulation for industrial and domestic wastewater and found the
technique to very effective although quite expensive; however, this could be
34. Page 22 of 109
offset by lower consumption of disinfection chemicals. The experimental
conditions for this study were a COD concentration 1596 mg L-1
O2, pH 3, a
Fe2+
concentration of 400 mg L-1
and a hydrogen peroxide concentration of 550
mg L-1
. This managed a COD reduction of greater than 90%.
The table below summarises the experimental conditions and results of the
studies detailed:
Table 2.5: Experimental conditions of Fenton’s reagent degradation case studies.
2.2.4 Adsorption
Adsorption can be simply defined as “the process of accumulating substances
that are in solution on a suitable interface” (Metcalf & Eddy, 2005). The
substance being removed from solution, the target compounds permethrin and
TBT, are the absorbates. Due to the high affinity of both permethrin and TBT to
be adsorbed by organic matter, or organic carbon as measured by their Koc
values (figures 2.1 & 2.2) adsorption is a viable alternative to advanced
oxidation processes for their removal from wastewaters.
Study Target
compound
Initial
concentration
H2O2
concentration
Iron
concentration
%
degradation
Arnold
(1993)
Atrazine 30.2 mg L
-1
2.69 mM L
-1
2.69mM L
-1
100
Gallard &
De Latt
(2000)
Atrazine 0.15 mg L
-1 0.2 mM L
-1
– 1
mol L
-1 0.2mM L
-1 No
information
Gallard &
De Latt
(2001)
Chloro-
benzenes &
Phenyl-
ureas
1 mol L
-1
In excess In excess
No
Information
Badaway
et al.
(2006)
Organo-
phosphorus
pesticides
50 mg L
-1
0.089 mM L
-1
8.99 mM L
-1
70
Badaway
& Ali
(2006)
Wastewater
COD:
1596 mg L
-1
O2
16.18 mM L
-1
7.14 mM L
-1
>90
35. Page 23 of 109
There have been numerous studies on the removal of both TBT from
wastewaters with adsorption (with a particular emphasis on GAC), as well as
conventional wastewater treatment processes (especially those wastes
produced by shipyards, where TBT concentrations can be as high as
1 mg L-1
).
Schafran et al. (2001) reported the removal of TBT from shipyard waters
including adsorption on to GAC at both laboratory and full scale treatment. The
GAC chosen was Calgon F400. Schafran concluded that GAC adsorption
removed almost as much TBT as clarification and filtration and throughout the
entire treatment train, as much as 99.8% removal was observed, the capacity of
the Calgon F400 was not reported. Schafran did note in this study that there
maybe a small amount of TBT not being removed from the effluent due to the
presence of fine particulate TBT. The experimental conditions used in this study
were an adsorption time of 24 hours, a solution pH of 7.7, the GAC used was
Calgon F400 and at a quantity of between 0-4 g with volumes of water treated
of 0.75 L g-1
, the initial concentration of TBT used was 4.08 mg L-1
in sonar
dome water
Schafran (2003) also reported the removal of TBT by GAC at laboratory and full
scale treatment processes in industrial wastewaters. This study discovered that
TBT was significantly adsorbed on to GAC where it could be degraded by
biological activity to dibutyltin, monobutyltin and eventually inorganic tin. The
biological activity contributed to the de-sorption of the tin species from the GAC
column as TBT was converted to mono or dibutyltin which desorbed from the
GAC. The GAC used was Calgon F400, a solution pH ranging between 5.5-8.5
to study pH effects on adsorption, TBT concentrations ranged between nothing
(to measure the amount of desorption) and 2mg L-1
. No other experimental
conditions were available
36. Page 24 of 109
Prasad & Schafran in 2006 using a combination of coagulation-flocculation-
clarification and adsorption, managed to achieve 99.9% removal of TBT on a
full scale treatment basis, over 75% of the time, during a three year study. The
GAC contactor provided the best performance in the removal of TBT with 99%
of TBT entering the contactor being removed, however this removal was not
sufficient to bring concentrations to below regulatory levels of 50 ng L-1
. The
experimental conditions used in these experiments were a flow of shipyard
wastewater at 190 L min-1
which had been adjusted to a pH of 7 and an initial
concentration of TBT ranging between 5.5 and 6260 g L-1
. The dose and type
of GAC was not given
The removal of TBT using powdered activated carbon does not appear to have
been widely studied (although the study by Prasad & Schafran in 2006
mentions PAC, it is not expanded upon). This is also the case for permethrin
using adsorption processes where it does not appear to have been widely
studied.
Table 2.6: Experimental conditions of adsorption case studies.
2.3 Current study
From this literature review it can be concluded that:
Study Target
compound
Initial
concentration
GAC GAC
concentration
% removal
Schafran
et al. (2001)
TBT 4.08mg L
-1
Calgon F400
0-4g
(0.75 L g
-1
)
99.9
Schafran
(2003)
TBT 0-2mg L
-1
Calgon F400 Not available
Not
available
Prasad &
Schafran
(2006)
TBT
0.0055 - 6.26
mg L
-1 Not available Not available 99.8
37. Page 25 of 109
• Studies have shown the toxicological effect of TBT, especially in
reference to its impact on oyster populations (Alzieu, 1991) and in
reference to its ability to cause imposex in the dog whelk (Yebra et al,
2004). Studies have also shown the toxicity of permethrin to the aquatic
environment especially its toxicity towards insect and invertebrate
populations (Kamrin, 1997) and also to fish populations (Sanchez-Fortun
& Barahona, 2005)
• The toxicity of TBT and permethrin to the aquatic environment, and their
presence within waste water treatment plant effluents, in high enough
concentration to cause environmental damage makes the study of their
removal necessary.
• Currently there is no active removal of either TBT or permethrin from
waste water treatment plants. Their removal is coincidental as they are
readily adsorbed onto the sewage sludge (Fent, 1996 ; Fent & Muller,
1991; Plagellat, 2004) and then the sewage sludge disposed of to land.
Despite this, concentrations at the effluent of wastewater treatment
plants are in high enough concentrations (Fent, 1996) to warrant the
need for further removal.
• Advanced oxidation processes and adsorption onto GAC have been
shown to be affective in the degradation of TBT (Schafran et al, 2001).
There appears to have been no studies to date on the removal of
permethrin using advanced oxidation processes or by adsorption.
As a result, this study will look at the removal of TBT and permethrin from
spiked de-ionised and wastewater treatment plant effluent using advanced
oxidation techniques and adsorption onto GAC.
38. Page 26 of 109
3.0 Objectives
The main objectives of this project were to:
• Evaluate the effectiveness of UV-photolysis, UV/H2O2 and Fenton’s
reagent to degrade permethrin and TBT in a deionised water and tertiary
wastewater matrix.
• Compare the removal efficiency in comparison to GAC.
• Evaluate the cost of each of the processes in treating a tertiary
wastewater matrix.
40. Page 28 of 109
THE REMOVAL OF PERMETHRIN AND TBT FROM WASTEWATER USING
ADVANCED OXIDATION PROCESS AND ADSORPTION
O.J.Grievson & S.A.Parsons.
Centre for Water Sciences, Cranfield University, Cranfield, Bedfordshire,
MK43 0AL, United Kingdom.
Abstract
Permethrin and TBT are a persistent problem in the aquatic environment with
annual environmental quality standard failures reported in the waterways and
coastal waters of the United Kingdom. Their presence has been reported in
wastewater treatment plant effluents, methods of their removal are required.
This paper examines the degradation of permethrin and TBT with three
advanced oxidation processes; UV-photolysis, UV/H2O2 and Fenton’s reagent.
The effectiveness and cost of each of the three processes is compared to
adsorption onto granular activated carbon (GAC).
UV/H2O2 was able to reduce permethrin and TBT simultaneously in water with
permethrin being removed to below the limit of detection and TBT removed to
89% of its initial concentration. UV-photolysis was able to remove both
permethrin in spiked deionised water but wasn’t as effective for TBT removal.
Fenton’s reagent wasn’t as effective as either of the photolysis methods for the
removal of the target compounds and adsorption, experimentally, wasn’t as
41. Page 29 of 109
effective as it should have been. This will be expanded upon in the results
section.
The final conclusion is that the cost of treatment was significant and for the
AOP’s predictions were £8.14 - £28.95 per m3
in comparison to treatment by
adsorption to GAC ranging between £0.02 – £0.07 per m3
.
Keywords: Permethrin, TBT, UV-photolysis, wastewater, Fenton’s reagent.
4.1 Introduction
TBT and permethrin can be classed as two of many micro-pollutants present in
UK waters that cause problems in the natural environment. The toxicity of TBT
has been well documented especially its properties as an endocrine disruptor
and its previous impacts on the oyster industry (Alzieu, 1991). Permethrin is as
well documented and its effects on aquatic eco-systems can include
invertebrate and fish deaths (Kamrin, 1997).
The prevalence of both of these compounds meant that between 2003-2005
there were a total of 53 failures of environmental quality standards (EQS) in
freshwaters from TBT and 17 failures in freshwaters from a combination of
cyfluthrin and permethrin (Environment Agency, 2007).
42. Page 30 of 109
TBT has been legislated since 1982 as a result of damage to the oyster
industry, with a standard set at 20 ng L-1
(Alzieu, 1991) and later in the United
Kingdom in 1986 (Clark et al, 1998). Current proposed legislation as part of the
Water Framework directive in the proposed directive on environmental quality
standards in the field of water policy is to see the maximum allowable
concentration of TBT set to fall to 1.5 ng/L and an average annual concentration
of 0.2 ng L-1
(Commission of the European Communities, 2006).
Environmental quality standards for permethrin are set at a local level by the UK
Environment Agency and permethrin is classed as a dangerous substance
under the Dangerous Substance Regulations (Environment Agency, 2007).
Permethrin and TBT have been measured in wastewater treatment plants at
concentrations up to 81 g L-1
for permethrin (Kupper et al, 2006) and
0.22 g L-1
for TBT (Fent & Muller, 1991) their removal from the wastewater
treatment process is imperative in order to comply with the proposed UK and
EU legislation.
Whilst a number of major studies have reported that the major route of both
permethrin and TBT within the wastewater environment is to be adsorbed to the
solids and get captured within the sludge (Fent, 1996; Plagellat, 2004) however
this will not remove either compound in the liquid phased to below the proposed
European limit of 1.5 ng L-1
(Schafran, 2003).
43. Page 31 of 109
As a result of this, it is necessary to look at treatment methods for the removal
of both permethrin and TBT from wastewater. Due to both compounds high
affinity to be adsorbed the use of granular activated carbon (GAC) is a
possibility and has been studied previously (Schafran, 2003) and found to be
effective but not enough to remove TBT to below the regulated concentration of
50 ng L-1
, in the shipyard waters that were studied. Due to the GAC’s inability to
reduce TBT to below the legislative limit Schafran also investigated the use of
UV/H2O2 advanced oxidation process and found it to be an effective process in
the removal of TBT from shipyard waters.
The aim of most advanced oxidation processes (AOP’s) is to produce the
hydroxyl radical in order to mineralise organic compounds to less toxic or ideally
harmless inorganic molecules (Parsons, 2004). In this study the advanced
oxidation process’s that are being used are UV photolysis in the UV-C
waveband at 254nm, a combination of UV photolysis and hydrogen peroxide
and finally Fenton’s reagent (a mixture of ferrous iron and hydrogen peroxide).
Their removal from tertiary wastewater streams either by degradation or by
adsorption onto GAC will be examined in terms of the effectiveness of the
techniques and their applicability to the wastewater industry.
44. Page 32 of 109
4.2 Materials & Methods
4.2.1 Reagents.
Permethrin (a mixture of cis & trans), TBT chloride, hydrogen peroxide (35%) &
sodium tetraethylborate were purchased from Sigma Aldrich. Ferrous sulphate
heptahydrate and n-hexane (Suprasolv grade) were purchased from VWR
International. Glacial acetic acid, sodium sulphate anhydrous, methanol, sodium
hydroxide & hydrochloric acid were purchased from Fisher Scientific. Apart from
the n-hexane which was of Suprasolv grade all chemicals purchased were of
analytical grade or better.
Mixed stock solutions of permethrin and TBT were prepared and stored in
amber glass bottles and stored at 4 ± 2ºC. Working solutions of permethrin and
TBT were prepared freshly from stock solutions. Sodium tetraethylborate
solutions (2% w/v) was prepared freshly as needed in a 25mL volumetric flask.
A 10% solution of glacial acetic acid was prepared in order to preserve all
sample/standard solutions (Readman & Mee, 1991).
45. Page 33 of 109
4.2.2 Samples
For the spiked deionised water degradation experiments (phase 1), working
solutions of permethrin and TBT were prepared freshly from stock solutions
using deionised water in a 5L volumetric flask to ensure sample
homogenisation.
For the spiked wastewater degradation experiments (phase 2), wastewater
effluent was taken from Cranfield University wastewater treatment works and
was spiked to the same concentration (0.51 mol L-1
permethrin and TBT) as
used in the spiked deionised water degradation experiments.
4.2.3 Degradation experiments
4.2.3.1 Conditions for UV degradation experiments
The UV degradation experiments were conducted using a collimated beam
apparatus. The collimated beam apparatus delivered UV-C from 4 x 30 W low
pressure Wedeco NLR 2036 UV-lamps providing a total UV dose of
19.18 W m-2
at 254nm, UV dose was determined by Uridine actinometry.
The samples were degraded in an open petri dish under an automated shutter
beneath the UV source. Gentle stirring was provided by a magnetic stirrer so as
to mix the sample but not cause undue surface disturbance.
46. Page 34 of 109
The degradation experiments can be split into two phases:
4.2.3.2 Spiked deionised water degradation experiments
The spiked deionised water degradation experimental conditions can be found
in table 4.1 below:
1 litre of spiked deionised water was degraded for two hours with stirring, the
exception to this are the UV experiments where a 250mL volume of sample was
used and degraded separately for each time period. 100mL aliquots were taken
at 0, 20, 60 & 120 minutes. Once taken each sample was quenched with 10mL
of methanol and preserved with 0.1mL of 10% acetic acid (Readman & Mee,
1991).
The adsorption experiments used 250mL of sample and 0.25, 0.5, 1.25 & 2.5 g
of Norit 1240 GAC, and were shaken on a horizontal shaker for 24 hours. After
24 hours 100mL aliquots were taken and quenched with 10mL methanol and
preserved with 0.1mL 10% acetic acid (Readman & Mee, 1991). Samples were
then refrigerated at 4±2 ºC prior to ethylation (see section 4.2.4). Due to the fact
that the Norit 1240 GAC was not ground before the experiments Freundlich
constants of k and 1/n are unable to be calculated
4.2.3.3 Spiked wastewater degradation experiments
The experimental conditions for the spiked wastewater degradation experiments
can be seen in table 4.2:
47. Page 35 of 109
All experiments were conducted as per the spiked deionised water experiments
including sample quenching and preservation.
4.2.4 Ethylation/Extraction
All samples, standards and analytical quality control samples followed the same
ethylation procedure. Each 100mL aliquot was simultaneously ethylated with
0.5mL of 2% sodium tetraethylborate (Huang, 2004) and extracted into hexane
by shaking for ½ hour. The ethylated sample was then transferred to a
separatory funnel to allow the two phases to separate, at this point the water
fraction was discarded The hexane fraction was dried under a stream of
nitrogen gas and re-dissolved using 2 mL of hexane and analysed by GC-MS
(section 4.2.5). The drying of the hexane fraction was omitted for the spiked
deionised water experiments.
4.2.5 Instrumental
A 2mL fraction of each standard/samples was transferred to a GC-MS vial and
added to the GC-MS auto-sampler (Agilent model 7683B) it was then analysed
by GC-MS (Agilent model 6890N, MSD 5973).
The GC-MS was setup with the following conditions:
48. Page 36 of 109
Column: ZV-5HT Inferno (30m long, internal diameter of
0.25mm and a 5% phenyl- 95% polysiloxane film
thickness of 0.25 m). (Gomez et al, 2007)
Injection volume: 2 L.
Injection mode: Splitless.
Carrier gas: Helium.
Scan range: 40-550 amu.
Solvent delay: 4 minutes.
Injector temperature: 270ºC
Transfer line
Temperature: 300 ºC
The oven programme was set at 80ºC for the first two minutes and then ramped
at 30 ºC min-1
to a temperature of 210 ºC, and then ramped at 3 ºC min-1
to a
temperature of 270 ºC. The oven temperature was then held for 4 minutes at
270 ºC. The total programme lasted 30 minutes and 20 seconds (adapted from
Esteve-Turillias et al, 2006).
An example of the GC-MS chromatogram as produced by the GCMS is shown
below (figure 4.1)
49. Page 37 of 109
4.3. Results
This section will reveal the result of the method development of the GC-MS
method used for this study and discuss the results of the degradation
experiments including a comparison of the quantum yield, rate constants and
electrical energy per unit order (EEO) costs.
4.3.1 Method development
The simultaneous analysis of TBT and permethrin by GC-MS was developed
solely for this study, rather that the more typical analysis by a combination of
GC-ECD (for permethrin) (Lee et al, 2002) and GC-FPD (for TBT) (Michel &
Averty, 1991). As a result of this method development was undertaken to
establish limits of detection and reproducibility of the method. These were
Limit of detection
Permethrin 0.05 g L-1
TBT 0.13 g L-1
Reproducibility
Permethrin 8 g L-1
TBT 13 g L-1
The limit of detection was determined by multiplying by six times the standard
deviation of the measurement of three blank samples. The reproducibility was
50. Page 38 of 109
determined by multiplying three times the standard deviation of the
measurement of three blanks and three samples (spiked to 0.51 M permethrin
and TBT). It is noted that the reproducibility of the analysis of TBT could have
been improved by the addition of an internal standard of another organotin
compound, such as tripropyltin.
4.3.2 Control degradation experiments
In the de-ionised water and wastewater degradation experiments two types of
control experiments were undertaken using (1) Ferrous sulphate and (2)
hydrogen peroxide
Ferrous sulphate addition (concentration 0.3 mM and 3 mM at pH 3 & 5) gave a
maximum removal of 37 % removal of TBT and 28% permethrin at the lower
concentration of 0.3mM Fe(II) and up to 74% degradation at the higher
concentration of 3mM Fe(II). This was due to the high Kow of both compounds
and the formation of ferrous sulphate flocs. When flocs were formed both
compounds naturally adsorbed to the flocculated particles. Ferrous sulphate
degradation was not undertaken on the wastewater matrix.
Hydrogen peroxide in de-ionised water hydrogen peroxide gave removals of
permethrin and TBT ranging from 0% up to 50%. In wastewater, the hydrogen
peroxide degradation of permethrin and TBT was similar to that of de-ionised
51. Page 39 of 109
water with maximum removal rates of 28% for permethrin and 46% for TBT. All
degradation experiments with hydrogen peroxide followed first order kinetics.
The control experiments, as was to be expected, showed generally lower
removal rates for both permethrin and TBT.
4.3.3 Direct & indirect photolysis
As both compounds have to be able to absorb UV light in order to be degraded
by UV-photolysis, experiments were undertake to establish the molar absorption
co-efficient of each compounds. This found that permethrin was highly
photolabile at a wavelength of 254nm and had a molar absorption co-efficient of
1332 M-1
cm-1
and TBT was not, with a molar absorption co-efficient of
3.3 M-1
cm-1
, this compares to 500 M-1
cm-1
for 2-4 di-chlorophenol, 1300 M-1
cm-1
for NDMA and 4000 M-1
cm-1
for diazinon and atrazine (Sabhi & Kiwi, 2001;
(Sharlpless & Linden, 2003 ; Shemer & Linden, 2006; Stefan & Bolton, 2005).
Given the molar adsorption co-efficient for permethrin, it is not surprising that
degradation of permethrin, using just UV-photolysis, gave almost complete
removal with an 85% reduction with a dose of 2300 mJ cm-2
. The addition of
H2O2 at this dose gave little improvement, but at the higher dose of 6900 mJ
cm-2
the concentration of permethrin fell to below the limit of detection (0.05 g
L-1
). The lack of improvement due to the addition of hydrogen peroxide, and the
generation of the hydroxyl radical, was due to the already high photolability of
52. Page 40 of 109
permethrin. All of the degradation experiments with UV and UV/H2O2 follow
initial first order kinetics. This can be seen in figure 4.2
However for TBT the low molar absorption co-efficient means that the
degradation of TBT with UV-photolysis alone is poor with removal increasing
from 25% at a dose of 2300 mJ cm-2
to a maximum of 45% degradation with a
UV dose of 13800 mJ cm-2
(Figure 4.2). Once H2O2 is added the degradation of
TBT is markedly increased with a degradation of 58% at 2300 mJ cm-2
(0.3mM
H2O2) and 89% (3mM H2O2) at a dose of 13800 mJ cm-2
. Degradation of TBT in
deionised water also follows initial first order kinetics
Permethrin degradation in wastewater is initially affected by the wastewater
matrix and the degradation after a dose of 2300 mJ cm-1
is greatly reduced to
that in deionised water especially with the higher concentration of 3 mM H2O2,
(Figure 4.3). However at a UV dose of 13800 mJ cm-2
both UV and UV/H2O2
reduce the concentration to below the limit of detection (0.05 g L-1
)
The degradation of TBT in wastewater exceeded the degradation in deionised
water. The mechanism that is involved was unknown but the quantum yield of
the reaction increased dramatically (ΦTBT in deionised water equalled 9 x 10-3
and in wastewater 2.4 x 10-2
an increase of approximately 20 times) as did the
rate constant, from that observed in deionised water (KTBT in deionised water
1.0 x 10-4
to KTBT in wastewater 1.1 x 10-3
, this is an increase of over an order of
magnitude). The degradation by UV in wastewater, was greater than that by
53. Page 41 of 109
UV/H2O2 at the lower concentration of 0.3mM H2O2 and equalled that of the
higher concentration of 3mM H2O2, with degradation by UV at a dose of 13800
mJ cm-2
equalling 99%, and with hydrogen peroxide addition 89% (0.3mM
H2O2) and 100% (below the limit of detection at 3mM H2O2). This can be seen
in figure 4.3.
4.3.4 Fenton’s reagent degradation experiments
The degradation of permethrin and TBT by Fenton’s reagent was not as
successful a treatment as UV & UV/H2O2. After a 120 minute degradation
period the most effective concentration of Fenton’s reagent (0.51:0.3mM
Fe(II):0.3mM H2O2) achieved a degradation of 73% of TBT at pH 3 (compared
to 37% removal for ferrous sulphate and 15% for H2O2) and 65% of permethrin
at pH 5 and the lower concentration of Fenton’s reagent (compared to 24%
removal with the ferrous sulphate and 28% removal with H2O2). Fenton’s
reagent at pH 5 was more successful at removing both compounds
simultaneously with a 54 % reduction in TBT and a 65 % reduction in
permethrin concentrations (compared with removal rates of 22% TBT and 24%
permethrin with ferrous sulphate and 46% TBT and 28% permethrin with H2O2.
54. Page 42 of 109
4.3.5 Adsorption experiments
The Kow value of permethrin of log 6.1 and the Kow value of TBT of log 4.1
indicate that both permethrin and TBT should readily adsorb onto GAC. The
experimental results indicate that at a GAC concentration of 10g L-1
54% and
60% of permethrin is adsorbed to the GAC and 89 & 90% TBT was removed
from deionised and wastewater respectively (figure 4.5).
Permethrin adsorption in deionised water is more dependent on the dose of
GAC than TBT with removal ranging between 7% at a dose of 1g L-1
to 54% at
10 g L-1
. In wastewater the removal of permethrin was 49% at a dose of 1 g L-1
and 66% at 10 g L-1
.
TBT in wastewater removed 15% at 1 g L-1
and 90% at 10 g L-1
. This can be
seen in figure 4.5
4.3.6 Rate constants, quantum yields & energy consumption
The processes evaluated here can be compared in terms of rate constants,
quantum yields and electrical energy required per order of degradation (EEO).
The rate constant has been calculated using the gradient of the degradation
and either calculated as a factor of time (s-1
) or for the UV degradation in terms
of UV dose (cm2
mJ-1
) and time (s-1
).
55. Page 43 of 109
The quantum yield has been calculated by comparing it to Uridine as a chemical
actinometer using the method as described in von Sonntag & Schumann
(1992).
The electrical energy required per order of degradation (EEO) has been
calculated using the method described by Bolton & Stefan (2002) and using the
following equation for a batch operation:
EEO (kWh m-3
per order) =
)log(
1000
f
i
C
C
V
Pt
where:
P = power of lamp in kW (0.12) (personal communication, Wedeco)
t = time (hours)
V = volume of reactor (L)
Ci = initial concentration (mol L-1
)
Cf = final concentration (mol L-1
)
The values for the quantum yield, rate constants and EEO are shown below
(tables 4.3 & 4.4).
56. Page 44 of 109
4.4. Discussion
4.4.1 Mechanisms of degradation
In this study there were two mechanism for the degradation of permethrin and
TBT, namely the direct photolysis of our compounds with UV light at 254nm and
by the generation of the hydroxyl radical either by the photolysis of hydrogen
peroxide (indirect photolysis) or by Fenton’s reagent.
The degradation of any compound by direct and indirect photolysis is governed
by two parameters, the molar absorption co-efficient and the quantum yield of
the degradation.
The molar absorption coefficient is part of the Beer-Lambert law and states that
a compound must be able to absorb light in order for it to degrade. In this study,
permethrin had a high molar absorption coefficient (1332 M-1
cm-1
) and was
significantly degraded by UV-photolysis. This is comparable to compounds such
as NDMA, simazine, atrazine, and diazinon which have molar absorption
coefficients higher than permethrin (1800, 3330, 3860, 4000 M-1
cm-1
. Sharpless & Linden, 2003; Nick et al, 1992; Shemer & Linden, 2006) and
thus have quantum yields and high rate constants for degradation with UV-
photolysis (Table 4.5). Permethrin’s photolability means that it is readily
degraded by UV-photolysis.
57. Page 45 of 109
TBT degradation also demonstrated the Beer-Lambert law in deionised water.
The molar absorption co-efficient was very low, and as a result its degradation
was very poor, the quantum yield was also low, and thus there was very little
degradation. The generation of the hydroxyl radical was required by the
photolysis of hydrogen peroxide to degrade TBT. This can be seen for
compounds such as azo dyes (table 4.5) which are not significantly degraded
by UV light but are degraded when the hydroxyl radical is generated.
In wastewater the situation changed, permethrin degradation decreased due to
the matrix effects of the wastewater. Transitivity of UV decreased and thus the
direct photolysis of permethrin decreased, only when hydrogen peroxide was
added and the hydroxyl radical generated where the matrix affects counteracted
and degradation of permethrin occurred.
TBT degradation was quite different and it seemed that degradation by direct
photolysis was more efficient in wastewater than deionised water despite the
low molar absorption coefficient. The quantum yield of the reaction increased by
20 times and the rate constant by over an order, the mechanism of this reaction
was unknown. However, studies by Mailhot et al. (1999) using the photolysis of
the ferric iron to generate the hydroxyl radical is a reasonable mechanism for
the observed degradation of TBT in the wastewater matrix. In Mailhot’s study
ferric iron generated the hydroxyl radical via the following reaction:
58. Page 46 of 109
Fe3+
+ hv → Fe2+
+ ●
OH + H+
Although not measured it is reasonable to suggest that the tertiary wastewater
effluent from Cranfield University’s wastewater treatment plant would have iron
concentrations in the range that Mailhot used in his study (≈1.7 mg L-1
). Thus,
what appeared to be the direct photolysis of TBT in wastewater was in fact the
photolysis of the ferric iron to generate the hydroxyl radical and degrade TBT by
indirect photolysis. The initial lag in the degradation was due to generation of
the hydroxyl radical by the ferric iron being slower than the generation of the
hydroxyl radical by hydrogen peroxide. The higher quantum yield and rate
constant of the reaction show that the UV/Ferric degradation was more efficient
at removing TBT that UV/H2O2. If the rate constant’s for UV/Fe(III) degradation
are compared for TBT only to the rate constants for Fenton’s reagent, it can be
seen that UV/Fe(III) is also more efficient that Fenton’s reagent, this suggests
that photo-Fenton’s or photo-Fenton like degradation processes maybe efficient
at removing TBT.
59. Page 47 of 109
4.4.2 Costs of treatment
The performance of each treatment process has been evaluated in terms of the
quantum yields, rate constants and also the electrical energy required per order
(EEo) of degradation. The calculated EEO for permethrin, 200 – 579 kWh per m3
and for TBT between 163 – 480 kWh per m3
(based on a wastewater matrix and
a UV dose of 2300 mJ cm-2
). Based on a cost of £ 0.05 per kWh of electricity
this means that a 90% reduction in permethrin and TBT in the wastewater
matrix would cost approximately £8.14 - £28.95 per m3
of wastewater using
either UV or UV/H2O2.
This compares to £0.15 for compounds such as atrazine (Muller & Jekel, 2001)
and £1.50 for NDMA (Stefan & Bolton, 2002) in tap waters to approximately
£100 for some azo dyes in laboratory experiments with distilled water
(Muruganandham et al, 2007).
In comparison to this the price of using GAC to adsorb TBT and permethrin
would be approximately £0.02 - £0.07 per m3
of wastewater, this is based on an
initial cost per m3
of £1200 for Norit GAC 1240 and a regeneration cost
including the removal of GAC from site of £600 per m3
(Norit, 2007).
60. Page 48 of 109
4.5. Conclusion
From the results of this study it can be concluded that:
• Amongst the AOP’s used in this study the most effective process used
was UV/H2O2 for the simultaneous removal of both permethrin and TBT
from wastewater based on the percentage removal of the compounds.
• Adsorption proved to be the most economical process with costs of
approximately £0.02 - £0.07 per m3
compared to £8.14 - £28.95 per m3
for removal using UV or UV/H2O2.
It can also be concluded that areas that require further study include
• Degradation products of the advanced oxidation processes needs to be
examined so as to ensure toxic by-products are not produced and that
the degradation to non-toxic by-products is complete.
• The effectiveness and economics of other AOP processes including, for
example, photo-Fenton’s reagent, especially considering the
effectiveness of UV/Fe(III) in the degradation of TBT.
61. Page 49 of 109
• The use of other types of adsorbents to remove TBT and permethrin
from wastewater and an examination of the economics of this removal
process.
• The use of advanced oxidation processes and adsorption on un-spiked
wastewater samples and how effective these samples would be on a
wastewater treatment plant scale
4.6 Acknowledgements
I would like to acknowledge the support of United Utilities in the funding of this
project as part of the STAMP scheme.
62. Page 50 of 109
4.7. References
Alzieu C, (1991), Environmental Problems caused by TBT in France:
Assessment, Regulations, Prospects, Marine Environmental Research, 32,
7-17.
Bhatkhande D, Kamble S, Sawant S, Pangarkar V, (2004) Photocatalytic and
photochemical degradation of nitrobenzene using artificial ultraviolet light,
Chemical Engineering Journal, 102, 283 – 290.
Bolton J, & Stefan M, (2002), Fundamental photochemical approach to the
concepts of fluence (UV dose) and electrical energy efficiency in photochemical
degradation reactions, Research on Chemical Intermediates, 28, 7-9.
Clark EA, Sterrit RM, Lester JN, (1988), The fate of TBT in the aquatic
environment: A look at the data, Environmental Science and Technology, 22(6),
Commission of the European Communities, (2006), Proposal for a directive of
the European parliament and of the council on environmental quality standards
in the field of water policy and amending directive 2000/60/EC, European
Parliament.
de Latt J, Gallard H, Ancelin S, Legube B, (1999), Comparative Study of the
oxidation of atrazine and acetone by H2O2/UV, Fe(III)/ H2O2/UV and Fe(III)/
H2O2, Chemosphere, 39 (15), 2693 – 2706.
63. Page 51 of 109
Environment Agency, Web reference http://www.environment-
agency.gov.uk/yorenv/eff/1190084/business_industry/agri/pests/917555,
accessed 15/08/07.
Esteve-Turrillas F, Pastor A, de la Guardia A, (2006), Comprison of different
mass spectrometric detection techniques in the gas chromatographic analysis
of pyrethroid unsecticide residues in soil after microwave-assisted extraction,
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Fent K, (1996), Organo-tin compounds in municipal wastewater and sewage
sludge: contamination, fate in treatment process and eco-toxicological
consequences, The Science of the Total Environment, 185, 151-59.
Fent K, Müller MD, (1991), Occurrence of organotins in municipal wastewater
and sewage sludge behaviour in a treatment plant, Environmental Science and
Technology, 25, 489-93.
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Pilot survey monitoring pharmaceuticals and related compounds in a sewage
treatment plant located on the Mediterranean coast, Chemosphere, 66, 993-
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compounds and their adsorption to different materials, International Journal of
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37-40, Lewis Publishers.
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Kupper T, Plagellat C, Brändli RC, de Allencastro LF, Grandjean D, Tarradellas
J, (2006), Fate and removal of polycyclic musks, UV filters and biocides during
wastewater treatment, Water Research, 40, 2603-12,
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samples during storage and extraction, Journal of Agricultural and Food
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photoinduced by iron (III), Applied organometallic chemistry, 13, 53-61.
Michel P & Averty B, (1991), TBT analysis in seawater by GC FPD after direct
aqueous-phase ethylation using sodium tetraethylborate, Applied
organometallic chemistry, 5, 393-397.
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–through pilot plants (Part 1), Water Science & Technology, 44 (5),
303 – 309.
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quantum yield and electrical energy per order (EEO) for advanced oxidative
decolourisation of reactive azo dyes by UV light, Journal of hazardous
materials, 144, 316-322.
Nick K, Scholer H, Mark G, Soylemez T, Akhlaq M, Schuchmannn H, von
Sonntag C, (1992), Degradation of some triazine herbicides by UV radation
such as used in the UV disinfection of drinking water, Journal of Water Supply
Research & Technology – Aqua, 41, 82-87.
Norit, (2007), Personal communication.
65. Page 53 of 109
Parsons S, (2004), Advanced Oxidation Process for water and wastewater
treatment, IWA.
Plagellat C, (2004), Origines et flux de biocides et de filters UV dans les stations
dépuration des eaux , biblion.epfl.ch/EPFL/theses/2004/3053/EPFL_TH3053
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sea water and its implication on water quality criteria, Marine Environmental
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catalysts, Water Research, 35 (8), 1994-2002.
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Shemer H, & Linden K, (2006), Degradation and by-product formation of
diazinon in water during UV and UV/H2O2 treatment, Journal of hazardous
materials, B136, 553 – 559.
Sharpless C, & Linden K, (2003), Experimental and Model comparisons of Low-
and medium- pressure Hg lams for the direct and H2O2 assisted UV
photodegradation of N-Nitrosodimethylamine in simulated drinking water,
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Stefan M & Bolton, (2005), Fundamental approach to the fluence-based kinetic
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67. Page 55 of 109
4.8 Tables
Table 4.1: Experimental conditions for spiked deionised degradation
experiments.
Degradation Conditions
Blank pH 3, 5, & 7
H2O2 pH 3, 5, 7 & 0.3 mM H2O2
pH 7 & 3 mM H2O2
Ferrous Sulphate pH 3, 5 & 0.3 mM Fe(II)
pH 3, 5 & 3 mM Fe(II)
UV photolysis pH 7 & 19.18 W m-2
UV/H2O2
pH 7, 19.18 W m-2
, 0.3mM H2O2
pH 7, 19.18 W m-2
, 3mM H2O2
Fenton’s Reagent pH 3, 5 & 0.3 mM Fe(II) & 0.3 mM H2O2
pH 3, 5 & 3 mM Fe(II) & 0.3 mM H2O2
Adsorption pH 7 & 0,1,2,5,10 g L-1
Norit 1240 GAC
Table 4.2: Experimental conditions for spiked wastewater degradation
experiments.
Degradation Conditions
Blank pH 7
H2O2
pH 7 & 0.3mM H2O2
pH 7 & 3mM H2O2
UV photolysis pH 7 & 19.18 W m-2
UV/H2O2
pH 7, 19.18 W m-2
, 0.3mM H2O2
pH 7, 19.18 W m-2
, 3mM H2O2
Adsorption pH 7 & 0,1,2,5,10 g L-1
Norit 1240 GAC
76. Page 64 of 109
(a)
Deionised water Wastewater
0
10
20
30
40
50
60
70
80
90
100
%Reduction
(b)
WastewaterDeionised water
0
10
20
30
40
50
60
70
80
90
100
%Reduction
Figure 4.5: Permethrin (a) and TBT (b) reduction by adsorption onto Norit GAC
1240 in spiked wastewater. ( 1 g L-1, 2 g L-1, 5 g L-1,
10 g L-1).
77. Page 65 of 109
5.0 References
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Assessment, Regulations, Prospects, Marine Environmental Research, 32,
7-17.
Alzieu C, (1998), TBT: case study of a chronic contaminant in the costal
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permethrin on guppies Poecilla reticulate, Chemosphere, 51, 469-74.
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Chiron S, Fernandez-Alba A, Rodriguez A, Garcia-Calvo E, (2000), Pesticide
chemical oxidation: State of the Art, Water Research, 34 (2), 366-77.
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environment: A look at the data, Environmental Science and Technology, 22(6),
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Parliament.
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partition coefficients (Kd) for diazinon, propetamphos and cis-permethrin:
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Fent K, Hunn J, Rennggli D, Siegrist H, (1991), Fate of TBT in sewage sludge
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sludge: contamination, fate in treatment process and eco-toxicological
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Fent K, Müller MD, (1991), Occurrence of organotins in municipal wastewater
and sewage sludge behaviour in a treatment plant, Environmental Science and
Technology, 25, 489-93.
Fukui S, Ogawa S, Motozuka T, Hanasaki Y, (1991), Removal of 3-chloro-4-
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Gallard H, De Latt J, (2001), Kinetics of oxidation of chlorobenzenes and
phenyl-ureas by Fe(II)/H2O2 and Fe(III)/H2O2. Evidence of reduction and
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405-13.
Gogate PR & Pandit AB, (2004), A review of imperative technologies for
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Hoch M, Schwesig D, (2004), Parameters controlling the partitioning of TBT
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80. Page 68 of 109
Kamrin M, (1997), Pesticide profiles: Toxicity , environmental impact and fate,
37-40, Lewis Publishers.
Kirk P, Rogers H, Lester J, (1989), The fate of chlorobenezenes and
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(9/10),1771-84.
Kupper T, Plagellat C, Brändli RC, de Allencastro LF, Grandjean D, Tarradellas
J, (2006), Fate and removal of polycyclic musks, UV filters and biocides during
wastewater treatment, Water Research, 40, 2603-12.
Lide D, (2005), CRC Handbook of Chemistry & Physics, 86th
Edition, CRC
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A 1.1 Performance characteristics of the method
Substance determined Degradation of TBT and permethrin.
Type of sample Spiked deionised water and wastewater.
Basis of method Samples are degraded by a number of
methods, ethylated, extracted with hexane
and concentrations are measured by GC-MS.
Calibration curve Linear to within the scope of this method.
Limit of detection 0.13 g L-1
TBT.
0.05 g L-1
total permethrin.
Reproducibility 13 g L-1
TBT.
8 g L-1
total permethrin.
Bias High iron concentrations negatively interfere
with the determination of both TBT.
