1.0
INTRODUCTION
Recently, there has been renewed interest in technologies for industrial
wastewater treatment (Kwon et al., 1999; Di Iaconi et al., 2010; Santos et al.,
2010; Shukla et al., 2010; Garcia-Segura and Brillas 2011) against the backdrop
that many of the components found in industrial discharges are refractory to
conventional biological oxidation (Guedes et al., 2003; Méndez-Arriaga et al.,
2009; Santos et al., 2010). These compounds have been known to have
detrimental effects on the environment with the potential of negative impacts on
human health and the ecosystem (Di Iaconi et al., 2010; Jagadevan et al., 2011;
Pang et al., 2011).
There are several traditional wastewater treatment technologies, namely
coagulation/flocculation, membrane separation (ultrafiltration, reverse osmosis)
and activated carbon adsorption (Lucas and Peres 2006). However, these
approaches remove toxic pollutants from wastewater simply by transferring them
from a liquid to a solid phase, leading to secondary pollution (Lucas and Peres
2006; Sun et al., 2011). Among the conventional treatment methods, biological
systems (BS) are the only exception, where potential degradation of the organics
occurs; these methods are arguably the most cost effective (de Souza et al., 2006).
However, BS are ineffective in the face of recalcitrant wastewaters as most of
these contaminants resist biological treatment. Furthermore, BS suffers from the
inactivation of several microbiologic processes due to the toxicity of these
wastewaters (de Souza et al., 2006; Lucas and Peres 2006; Méndez-Arriaga et al.,
2009). The implication is that the receiving water bodies risk contamination from
these toxic compounds (Méndez-Arriaga et al., 2009).
There are also certain recalcitrant compounds such as non-steroidal antiinflammatory drug (NSAIDs), where the application of oxidants alone, such as
O3, H2O2, permanganate, etc., not only fails to completely degrade the
compounds, but requires large amounts of oxidants (Méndez-Arriaga et al., 2009).
It thus becomes imperative to adopt oxidation technologies with potential for
partial or complete mineralisation of the wastewaters, thereby leading to less toxic
by-product discharges from the wastewater treatment plants. In this direction, the
hydroxyl radical (•OH), a strong oxidising species with an oxidation potential of
2.8 eV, has been deemed most suitable. •OH is regarded as the second strongest
oxidising species after fluorine which has a known oxidation potential of 3.0-3.2
eV (Lucas and Peres 2006; Garcia-Segura and Brillas 2011). However, there is a
technical limitation on the use of fluorine in wastewater treatment (Bigda 1995).
Thus, •OH can be viewed as the most utilisable oxidant which potentially attacks
all organics in a non-selective mode, leading to their overall mineralisation to
carbon dioxide (CO2), water (H2O) and inorganic mineral salts by the reactivity of
•
OH attack (Méndez-Arriaga et al., 2009; Garcia-Segura and Brillas 2011;
Jagadevan et al., 2011; Sun et al., 2011). The use of •OH as the main active
species in the treatment of wastewaters has been proven by many researchers to
be most effective (Ahmadi et al., 2005; Benatti et al., 2006; Alaton and Teksoy
2007; Mahiroglu et al., 2009; Umar et al., 2010; Diya'uddeen et al., 2011). There
are several oxidative processes utilising the in situ generation of •OH which are
generally referred to as advanced oxidation processes (AOPs). The AOPs are
broadly categorised into homogenous and heterogeneous systems, and both
systems can be assisted with ultrasound, electron beams or irradiation (Umar et
al., 2010). However, for the purpose of simplicity and cost concerns, the study
scope of this review is limited to the non-assisted systems.
A very promising AOP wherein wastewater reacts with hydrogen peroxide
(H2O2) in a non-pressurised reactor, at low temperatures, in the presence of a
suitable low-cost transition metal catalyst, is the classical Fenton oxidation (CFO)
process (Guedes et al., 2003). Other forms of the CFO are namely electro-Fenton
(electron beam) and photo-Fenton (irradiation). However, these two oxidation
routes require external sources of reactive energy and thus add to the total
operating cost. Hence, the interest of the research presented here is in the CFO.
The CFO can be run as a homogeneous or heterogeneous system.
However, the homogeneous systems benefits from a lower mass transfer
resistance between phases in comparison to the heterogeneous ones. Thus, rapid
degradation and mineralisation of contaminants is enhanced in the former (Umar
et al., 2010). The enumerated merits of the Fenton oxidation in the literature are
the simplicity of the equipment, safe operation, minimal sludge generation, high
organic destruction efficiency and readily availability reagents (Lopez, et al.,
2005; Lucas and Peres 2006). These factors have promoted the drive for its
utilisation in the treatment of various recalcitrant wastewaters and have made the
CFO one of the most employed AOPs in the management of recalcitrant
contaminants from wastewaters (Santos, Yustos et al., 2010). Included among
these wastewaters are metalworking fluids (Jagadevan et al., 2011), pesticides
(Chen et al., 2007), pharmaceuticals (San Sebastián Martínez et al., 2003; Tekin et
al., 2006; Elmolla and Chaudhuri 2009), chip board (de Souza et al., 2006), textile
effluents (Pérez et al., 2002; Xu et al., 2004; Alaton and Teksoy 2007; Liu, Chiu
et al., 2007; Papadopoulos et al., 2007), azo dyes (Sun et al., 2007; Xu and Li
2010), livestock waste (Lee and Shoda 2008), petroleum refining (Coelho et al.,
2006; Diya'uddeen et al., 2011), cork production (Guedes et al., 2003; Peres et al.,
2004), acid mine drainage (AMD) flotation circuits (Mahiroglu et al., 2009),
cosmetics (Bautista, Mohedano et al., 2007), olive mills (Rivas et al., 2001;
Ahmadi et al., 2005; Kiril Mert et al., 2010) and chemical laboratories (Benatti et
al., 2006).
However, despite the advantages of the CFO, its effectiveness is marred
by sludge generation, the need for pH adjustment before and after oxidation and
residual catalyst in the treated effluent which requires further purification. Thus,
the Fenton oxidation may not be deemed to be attractive for application in the
case of large-scale chemical processes because of the costly oxidant requirements
and the supposed generation of an appreciable amount of sludge. These
drawbacks may pose severe limitations on the success of wider applications in the
face of emerging restrictive legislature on sludge disposal and the attendant cost
of treatment approximated at 1-2 Euro/m3 of wastewater (Di Iaconi et al., 2010),
as well as the overall economic assessment of the wastewater treatment processes
factoring in the oxidant cost.
In light of this, the current review presents a critical discussion on the
limitations of CFO, a brief overview of CFO related •OH producing methods and
provides a justification for the suitability of the CFO oxidation process despite
these limitations. Furthermore, the significant contributions made by various
researchers for an alternate approach for the use of conventional oxidants are
highlighted. In addition, this review draws sound conclusions on sludge
generation, which is poorly addressed in most works in the literature on Fenton
oxidation.
2.
FENTON OXIDATION
2.1
History
A brief overview of the early history of Fenton oxidation starts with the earliest
work of H.J.H. Fenton himself (Fenton 1894). He discovered that H2O2 could be
activated by Fe(II) salts to oxidise tartaric acid. Thus, iron-catalysed H2O2
activation became known as Fenton’s oxidation. Several metals were then
investigated and were reported to have potential in terms of their electron transfer
properties which improve the use of H2O2 and, in the case of certain metals, have
strong catalytic power to generate highly reactive •OH.
A summary of an exhaustive review of the history of Fenton oxidation
covering the century of its existence has been reported by Pignatello et al. (2006),
while a detailed personal history and the chronicle of the research leading to the
discovery of Fenton oxidation by Krzysztof (2009) are summarised and presented
in Table 1.
The laboratory investigations representing the final phase of the trend in
Fenton oxidation (Table 1) promoted interest in the process as the robustness and
economic benefit of the technique was established. This directly accounted for its
widespread application to varied wastewaters; the efficiency of the process was
thus further established. This is evident from the large number of scientific
publications on Fenton oxidation using wastewaters related to the natural
environment, biological chemistry, synthesis, the chemistry of natural waters and
the treatment of hazardous wastes (Pignatello et al., 2006).
Table 1. Summary of the trends in Fenton oxidation evolution (Pignatello et al.,
2006; Krzysztof 2009)
Researcher(s)
Fenton
Fenton
Fenton
Haber and
Weiss
Barb and coworkers
Eisenhauer
Walling and
co-workers
Buxton and
co-workers
Several
workers
Findings/Highlights
Year
Correspondence to the journal editor of Chemical 1876
News describing a violet colour emanating from
the reaction between H2O2, tartaric acid, an Fe(II)
salt and a base as a test for tartaric acid
Identifying the violet colour as arising from a 1881
complex formed between iron and the oxidation
product of tartaric acid
Pioneered the Fenton oxidation by the discovery 1894
that Fe(II) salts could activate H2O2 and
potentially oxidise tartaric acid in a characteristic
way, providing a new and valuable
oxidising agent
Proposed that the active oxidant generated by the 1934
Fenton reaction is •OH
Proposed the chain reaction mechanism in the 1949Fenton oxidation as “classical” or “free radical” 1951
after several years of investigation on organic
compounds
Application of Fenton oxidation in small-scale 1964
industry
Novel work on the free radical pathway of the 1975
Fenton oxidation. The reported results were
central to the understanding and subsequent
proliferation of research in varied branches of the
Fenton oxidation.
Documented over 1700 rate constants for •OH 1988
reactions with organic and inorganic compounds
in aqueous solution
The beginning of extensive research on the 1990
applications of Fenton oxidation for waste
treatment in academic laboratories
Using the keyword “Fenton oxidation”, a search on 15 September 2011
provided approximately 20,000 scientific articles. However, the number of
publications appearing on Fenton oxidation has climbed steadily over the last two
decades, as only over 2,000 articles appeared prior to 1992. The rest were from
1993 to the present, as shown in Fig. 1.
2000
No of Publications
1500
1000
500
0
1992
1994
1996
1998
2000
2002
2004
2006
2008
2010
2012
Year
Fig. 1: Number of publications on Fenton oxidation (Thomson Reuters,
2011)
2.2
Fundamentals of the Fenton and Fenton-type oxidation systems
The principle reactions guiding Fenton oxidation have been extensively discussed
in the literature (Bigda 1995; Guedes et al., 2003; Benatti et al., 2006; Mahiroglu
et al., 2009; Kiril Mert et al., 2010; Santos et al., 2010; Umar et al., 2010). The
process is basically an oxidative cycle with •OH being generated through the
catalytic decomposition of H2O2 by iron (II) ions (Fe2+) as shown in Eq. (1); iron
(III) ions (Fe3+) are also produced. The cycle is completed by the reduction of
ferric iron by H2O2 and the subsequent generation of perhydroxyl radicals and
ferrous iron. The sequences of the reactions resulting in radical generation are
shown in Eqs. (1) – (3) (Lucas et al., 2007).
Fe2+ + H2O2→ Fe3+ +OH− +•OH
(1)
Fe3+ + H2O2→ Fe2+ +HO•2 +H+
(2)
Fe3+ + HO•2 → Fe2+ +O2 +H+
(3)
The catalytic reaction (Eq. (4)) between the radical and organic
components (represented by RH) is very fast (Lee and Shoda 2008). The entity
RH is the organic substrate, which is comprised of carbon chains and/or rings and
other elements, such as oxygen or nitrogen. The mineralisation of the RH
component of the wastewater generates free organics as transient intermediates,
which are further oxidised to more stable products by •OH, Fe3+, Fe2+, H2O2 and
O2, as shown in Eqs. (4) – (7) (Hermosilla et al., 2009; Umar et al., 2010).
RH + •OH → R• + H2O
(4)
R• + Fe3+→ R• + Fe2+
(5)
R• + Fe2+→ R• + Fe2+
(6)
R• + H2O2→ ROH + •OH
(7)
R• + O2→ ROO•
(8)
In the dark, the reaction depicted in Eq. (1) is retarded after the complete
conversion of Fe2+ to Fe3+ (RodrÃ-guez et al., 2005). Fig. 2 shows the photoFenton schematic diagram, where unlike the CFO, the use of irradiation results in
the photo-reduction of the Fe3+ to Fe2+ ions, thereby leading to enhanced ·OH
generation as shown in Eq. 9 and 10 (RodrÃ-guez et al., 2005).
Fe3+ + H2O hv→ Fe2+ + •OH + H+
(9)
H2O2 hv → 2(•OH)
(10)
Fig. 2. Schematic diagram of photo-assisted Fenton oxidation (RodrÃ-guez et al.,
2005)
Furthermore, as suggested from the photo-induced, ligand-to-metal
charge-transfer reactions shown in Eq. 11 & 12, additional Fe2+ and •OH are
produced from the light absorbing and potentially photolabile ferric iron
complexes, namely [Fe3+(OH) -]2+ and [Fe3+(RCO2)-]2+ (Pérez et al., 2002;
RodrÃ-guez et al., 2005). This drives the photo-Fenton reaction, producing
additional •OH and the recovery of Fe(II) needed in the Fenton reaction. The
overall benefit is the relatively higher mineralisation observed in the case of
photo-Fenton oxidation resulting from the decarboxylation of organic-acid
intermediates.
[Fe3+(OH) -]2+ hv → Fe2+ + •OH
(11)
[Fe3+(RCO2)-]2+ hv → Fe2+ + CO2 + R·
(12)
On the other hand, the electro-Fenton (EF) process is basically an
electrically assisted CFO where both Fenton oxidation and electro coagulation are
combined in the same system (Atmaca 2009). The hybrid system and electrical
supply result in increased performance arising from (i) increased oxidising power
of the H2O2, (ii) enhanced generation of •OH and (iii) combined efficiencies of the
hybrid system (Atmaca 2009). In the EF, two options are available for the Fenton
reagents to react: either externally adding both reagents to the reaction mixture
and using as the anode material an inert electrode with high catalytic activity or
by externally adding the oxidant only (H2O2) and providing the catalyst (Fe2+) in
the form of sacrificial cast iron anodes (Atmaca 2009).
However, assisting the CFO system requires energy; the radiation
(electricity) requirement for photocatalytic reactors has been reported by Bandara
et al. (1997) to comprise about 60% of the total operating cost.
2.3
Wastewater mineralisation and degradation by Fenton oxidation
Fenton oxidation, as with most other AOPs, is mainly used for recalcitrant
wastewater not amenable to conventional wastewater treatment. Thus, a general
overview is presented here with regards to recalcitrant wastewater degradation by
Fenton oxidation. Broadly, recalcitrant wastewaters are composed of aromatic
ring organic compounds which account for their resistance to biological
degradation (Pang et al., 2011). The mineralisation of these organics occurs
through non-selective attack of the contaminants by the produced •OH. This attack
follows three generally accepted mechanisms (Neyens and Baeyens 2003;
Navalon et al., 2010), the graphical representation of which is shown in Fig. 3.
i – as oxidising agent; ii – as hydrogen abstraction agent; iii – as electrophile; S - Substrate
Fig. 3. Schematic representation of radical attack mechanisms (Navalon et al.,
2010)
The primary products formed from the degradation of these aromatics
include large quantities of carboxylic acid, methylbenzene, benzene, pmethylphenol and paradioxybenzene (Pérez et al., 2002; Gai 2009). The rapid
attack of organic substrates (RH) by •OH occurs by H-abstraction, electrophilic
reduction and addition to the C-C unsaturated bonds which eventually cause the
chemical decomposition of these compounds (Benitez et al. 2000). However, the
most common route is addition to aromatic, heterocyclic rings and unsaturated
bonds of alkenes or alkynes (Gatrell and Kirk 1992; Neyens and Baeyens 2003),
thus enhancing the degradability of the resulting aliphatic compounds. However,
as the aromatic ring collapses during its hydroxylation, low molecular-weight
carboxylic acids are obtained as intermediates (Yalfani et al., 2009), among which
the most notable fractions are benzoquinone and the carboxylic acid groups
muconic, maleic, succinic, malonic, oxalic, formic and acetic acids (Hermosilla et
al., 2009; Iurascu et al., 2009). Some of these intermediates have been established
as non-oxidisable by the Fenton reagent in conventional Fenton oxidation (Bigda
1995). Generally, the presence of high amounts of oxygen-containing compounds,
e.g., carboxylic acid groups such as oxalic and acetic acid, has been reported to
slow the degradation process of •OH radicals (Lopez et al., 2005; Bedoui et al.,
2009).
A possible explanation for the •OH attack resistance of alicyclic
compounds is that the electrophilic •OH radical attack cannot occur at conjugated
C=C double bonds, in contrast to aromatic compounds, in which ring opening and
further degradation can take place (Neyens and Baeyens 2003). Research findings
by Hermosilla et al., (2009) further support this conclusion, as they observed this
limitation in mineralising acetic acid by a conventional Fenton oxidation. This
leads to coupling the Fenton oxidation to biological treatment or enhancing
radical generation by UV assistance.
In the case of coloured wastewaters, decolourisation is observed to be fast,
first because radicals attack the azo groups and open N=N bonds, thus destroying
the long conjugated π systems, and consequently causing decolourisation. Second,
owing to the fact that N=N bonds are more easily destroyed than aromatic ring
structures, the elimination of adjacent ring structures requires more time (Lucas
and Peres 2006).
3.0
FENTON OXIDATION REAGENTS AND OPERATING MEDIA
The reagents requirements for the CFO are limited to the oxidant and catalyst. For
the oxidant, the optimal performance of the conventional oxidant (H2O2) is
limited by the reaction mixture pH, as well as the cost and composition of the
wastewater ions. This section presents a brief review of these components.
3.1
The oxidant type and source
3.1.1
Alternate oxidants
There are several potential compounds for use as oxidants. This was proposed by
Fenton (1894) where he proposed the use of chlorine water (a solution of chloric
(I) (hypochlorous) acid (HOCl)/hypochlorite (OCl−)) as an alternative to H2O2. Of
the available oxidants, the most common are persulphate (PS), peroxydisulphate
(PDS) and peroxymonosulphate (PMS). However, the traditional and most
commonly employed oxidant is H2O2. The use of H2O2, however, has the
problems of a small window of optimal operating pH and its relatively high cost.
In this regard, several novel approaches have been proposed that
adequately address this problem by using an oxidant with oxidising abilities
comparable to those of •OH by substituting H2O2 with persulphate (PS) and
peroxymonosulphate (PMS), which generate sulphate radicals (SO4•-). Several
authors have reported the use of sulphate-based oxidants such as peroxydisulphate
(PDS, 2KHSO5·KHSO4·K2SO4) and peroxymonosulphate (PMS, K2S2O8) to
provide sulphate radicals (Sun et al., 2011). Reports in the literature indicate that
there is a wide range of pH at which the activation of PS and PMS produces SO4•(1-10.5) (Antoniou et al., 2010). Sulphate radicals, with an oxidation potential
ranging between 2.5 to 3.1 eV, have been shown to result in improved oxidation
of wastewater organic contaminants relative to traditional H2O2 despite a lower
oxidation potential (Sun et al., 2011). Moreover, the activation of these oxidants
can be achieved using different activation processes such as (i) catalyst activation,
(ii) radiation, (iii) thermal energy, (iv) anions and (v) UV irradiation, as shown in
Eqs. 13-20 for PMS and PDS (Antoniou and Dionysiou 2007; Yang et al., 2010).
PMS reaction with heat, metal ions and UV
Mn+ + HSO5- → M(n+1)+ + SO4·- + OH-
(13)
HSO5- + hv → SO4·- + •OH
(14)
HSO5- + heat → SO4·- + •OH
(15)
HSO5- + e → SO4·- + OH-
(16)
PDS
Mn+ + S2O82- → M(n+1)+ + SO4·- + SO42-
(17)
S2O82- + hv → 2SO4·-
(18)
S2O82- + heat → 2SO42-
(19)
S2O82- + e → SO4·- + SO42-
(20)
However, Yang et al. (2010) have shown that although PMS was activated
with some anions, its heat activation failed within a temperature range of 2580°C, while PS showed relatively good degradation efficiency with heat but could
not be anion activated. Although the thermal activation of peroxydisulphate has
been shown to be good, its efficiency increases with increasing temperature and
oxidant concentration (Diaz Kirmser et al., 2010).
Considering the activation of these two emerging oxidants, H2O2-Fenton
oxidation still competes favourably with these oxidants. This is further supported
by comparative studies in which the three oxidants were evaluated for the
degradation of the azo dye Acid Orange 7 (AO7); it was established that this
process could be better activated by irradiation than by PMS, since PS cannot be
anion activated and PMS cannot be heat activated. Moreover, the use of sulphate
radicals (SR) has been shown to be limited in media composed of cyclic organic
compounds which imposes limitations that prohibit the reaction with SR
(Antoniou and Dionysiou 2007). In contrast to •OH, SR are organic-selective
oxidants (Huang et al., 2009); thus, •OH is prone to more aggressive attack in
most organic contaminated wastewaters.
3.1.2
In situ generation
The oxidant can also be entirely generated, thus saving on the cost of transport for
the bulk oxidant and addressing issues related to the cold storage requirement for
H2O2. Moreover, as H2O2 is consumed before decomposing to H2O and O2
through disproportionation as shown in eq. (21), oxidant loss is greatly reduced
(Yalfani et al., 2009). However, literature reports on the utilisation of this
methodology are scarce in terms of its efficacy in organic destruction (Yalfani et
al., 2009).
2H2O2→ H2O + O2
(21)
A method of direct H2O2 generation from a mixture of hydrogen (H2) and
oxygen (O2) over catalytic systems was proposed by several researchers. In this
regard, Yalfani et al. (2009) detailed the hydrogen-oxygen mixture (H2-O2) trend
for the H2O2 generation from many reported articles utilising H2-O2 mixture as the
substrate and metal-supported titanium silicate (TS-1) as the catalyst. These
systems were more focused on hydroxylation and epoxidation purposes, and on an
industrial scale, the risk of explosion associated with utilising the H2-O2 mixture
was further magnified. This necessitates replacing the H2 with a more attractive
compound. In line with this, the use of hydrazine and hydroxylamine were
proposed; however, toxicity and/or explosivity limit their application. A novel
solution addressing both issues was advanced by Yalfani et al. (2009), where
formic acid was found to be a suitable substitute for H2, the former being an
environmentally compatible and safer alternate. They demonstrated the efficiency
of the O2-formic system for phenol treatment where up to 60% mineralisation was
obtained. Moreover, formic acid is a known recalcitrant compound which is
difficult to degrade by the Fenton reaction and has been described to account for
the residual TOC after treatment (Navalon et al., 2010). Thus, its utilisation by
addition into recalcitrant wastewater seems to be unattractive. Furthermore, the
requirement for acid media has not been addressed.
In summary and in comparison to other bulk oxidants, H2O2 is relatively
safe, easily activated and handled, and poses no lasting environmental threat since
it readily decomposes to water and oxygen (Pignatello et al., 2006; Yalfani et al.,
2009). Thus, the potential oxidants considered in the literature as replacements are
not as attractive.
3.1.3
Minimising consumption
The cost of the oxidant arises from the high H2O2 concentration requirements
typical of the Fenton oxidation (Santos et al., 2010). However, many processes
have been reported to occur at H2O2 below the stoichiometric requirements.
The use of modified Fenton catalysts has been reported to result in lower
consumption. For instance, Dantas et al. (2006) evaluated the treatment of textile
wastewater using Fe2O3/carbon composites as a heterogeneous catalyst and
observed a lower consumption of H2O2 than that required by the homogeneous
Fenton process. Likewise, there are studies that have proposed alternate routes to
the application of CFO while minimising the loss of this oxidant due to the
complexation of the catalyst (iron cations) with chelating oxidation intermediates
(Santos et al., 2010).
Santos et al. (2010) showed that the effective total mineralisation of
organic contaminants at even lower oxidant dosages than those theoretically
computed could also be achieved. They demonstrated the feasibility of this by
developing and validating a simplified pseudokinetic model that predicted the
mineralisation evolution of organic pollutants. The novelty of the model lies in its
applicability to different wastewaters. Moreover, use of an excess oxidant dosage
does not improve mineralisation because an asymptotic value of the TOC percent
removal is usually obtained, supporting the use of an oxidant dosage lower than
that theoretically required by stoichiometry which should be more convenient
from an industrial point of view (Santos et al., 2010).
The reduction step of the transition metals is known to enhance the
treatment process. Towards this end, several phenol derivatives have been
reported to potentially achieve this. Others have reduced oxidant consumption by
promoting the reduction of the Fenton-like catalyst. Aguiar and Ferraz (2007)
evaluated the reduction of Fe3+ and Cu2+ in the presence of the five most efficient
reducers of the two metallic ions. The reducers, namely 3,4dihydroxyphenylacetic, 2,5-dihydroxyterephtalic, gallic, chromotropic and 3hydroxyanthranilic acids, were found to comparatively decolourise slightly better
than CFO with 87% for the 3-hydroxyanthranilic acid/Fe3+/H2O2 against 75% for
CFO. However, the same study reported higher efficiency of CFO over other
reducers investigated, where gallic and syringic acids, catechol and vanillin
attained a lower reduction than CFO.
3.1.4
Other sources
Yasar et al. (2007) demonstrated the feasibility of using bleaching wastewater
(BWW) sourced from a finishing mill as the H2O2 source. This approach has the
merit of minimising resource consumption and serves to reduce environmental
pollution by utilising a wastewater stream and doing away with costly H2O2.
Technically, the need to refrigerate H2O2 is negated. Their findings in the
treatment of a reactive dye (Blue CL-BR) showed superior performance than
using analytical grade H2O2. This was expected as there was a dual effect in the
removal from the high solids contents in the BWW by adsorption and coagulation
by the catalyst to result in the 10% relatively higher decolourisation. Interestingly,
the improved efficiency of BWW treatment was attained using only 50% of the
catalyst concentration when using analytical grade H2O2.
3.2
Catalyst limitations
The involvement of the catalysts in •OH generation is crucial, as direct oxidation
using UV (photolysis) is limited by high-power requirements and have been
reported to fail in the treatment of refractory compounds (Sun et al., 2011). The
conventional catalyst used for the CFO is iron (II), which has the benefit of being
more cost effective, safe to use and environmentally friendly (Pignatello et al.,
2006). In addition, the performance of the catalyst is comparatively higher than
other transition metal catalyst. However, transition metals exhibiting at least two
oxidation states can potentially activate H2O2. Such metals include copper,
ruthenium, cerium and manganese.
Xu et al. (2004) investigated the catalytic ability of four different metals,
namely Fe2+, Cu2+, Mn2+ and Ag+. Their findings showed a steady decrease in
performance in the order Fe2+> Cu2+> Mn2+> Ag+. A range of Fe-based catalyst
exists, namely Fe 3+, Fe2+, and Fe0. Khan et al. (2009) compared the different Fe
types on four different organic contaminants, and showed that a wastewater
mineralisation efficiency and biodegradability improvement with Fe2+ in
comparison to Fe3+; Fe0 was superior. In the case of the CFO and Fenton-like
processes, the acidic pH requirement has been established at a range between pH
values of 2 to 4 (Shukla et al., 2010). However, many wastewaters are at a pH of
outside the operational limit of the catalyst. Thus, the modification and
development of a catalyst that allows for oxidation to occur at high pH has
become paramount. One catalyst with comparable efficiency to the CFO and
exhibiting sufficient performance even in alkaline media is the immobilised
Nafion-Fe3+ catalyst (Balanosky et al., 1999). The observed degradation indicated
that the system’s performance was not affected by the pH which allowed for
treatment at a pH of 8. Other non-conventional catalysts include the use of a steel
industry by-product, steel dust, although higher amounts are consumed
(approximately double) with 10% lower removal efficiency. Some drawbacks of
utilising this catalyst, aside from non-availability, are the preparation
requirements necessary for enhancing the specific surface areas to allow for
increased access to catalytic sites. This is achieved by the removal of surface
anchored groups and aggregation reduction through chemical acid enhancement
(Lee et al., 2009).
Other researchers have focused on heterogeneous Fenton oxidation
catalysis to provide a better alternative to homogeneous oxidation. Several novel
approaches have been reported utilising highly ordered structure and non-ordered
mesoporous silica materials, summarised by Shukla et al. (2010). They showed
that data are scarce in the literature in terms of studies exploring the large pore
size and high surface area of these materials. A detailed summary of the materials
studied in this subject area includes hydrothermal co-condensed silica loaded with
Fe, non-ordered, iron-containing on mesostructured SBA-15 catalysts, nonordered mesoporous silica, nanocomposites of the same SBA-15 silica with
crystalline Fe2O3 and CuO and immobilised iron oxide nanoparticles in aluminacoated mesoporous silica (Shukla et al., 2010).
3.3
Wastewater pH value
The CFO is significantly affected by changes in pH. This is associated with the
chemical limitation of the catalyst and the oxidant in the case of H2O2. Based on
several studies, a pH range of between 3 and 3.5 has been established as the
optimum (Guedes et al., 2003). For instance, the decrease in oxidation potential of
•
OH as the pH of the reaction mixture increases (Kiril Mert et al., 2010) and the
catalyst activity is reduced by the formation of Fe(II) complexes such as [Fe(II)
(H2O)6]2+, which react more slowly with H2O2 (Guedes et al., 2003). At pH values
lower than 3, H2O2 is stabilised as H3O2+ (Kwon, Lee et al., 1999) which might
account for the decreased performance in many reported systems. For instance,
Lucas and Peres (2006) observed an increase in RB5 decolourisation from 8% to
51.6% at a pH of 2. Low acidic values favour the conversion of Fe2+ to Fe(OH)2+
and Fe(OH)2+ and this directly limits the •OH generation step. Moreover, more
hydrogen ions (H+) are available within the reaction mixture as a consequence of
the scavenging effect of •OH by H+, thereby inhibiting the all-important reaction
of Fe3+ with H2O2 (Guedes et al., 2003). This leads to low amounts of •OH, which
is needed for sustaining the treatment.
However, between pH values of 4 and 8, the decolourisation was not
better than at pH 3. Under these conditions, the amount of Fe3+ within the system
is reduced, resulting in Fe3+ precipitation as Fe(OH)3 which subsequently retards
the production of •OH as a result of catalytic decomposition of H2O2 to O2 and
H2O this also eliminates the continued reaction of Fe3+ and H2O2 (Lucas and Peres
2006).
In the case of a heterogeneous catalyst, the pH has been shown to be a
significant factor on the reaction rate. Shukla et al. (2010) reported on the
oxidation of 2,4-dichlorophenol using Fe/SBA-15, and observed an increased
reaction rate when the solution acidity was increased from a pH of 6 to 3 reaching
a maximum at pH 3. Thus, pH modulation to a value of between 3-3.5 becomes
necessary for optimal performance and this accounts for the use of this pH limit
by most researchers (Arslan et al., 2000; Rivas et al., 2001; Pérez et al., 2002;
Benatti et al., 2006; Shukla et al., 2010).
Still, recycling dried iron sludge from the homogeneous process negates
the need for pH adjustment, as Martins et al. (2010) have observed the elimination
of this step due to the release of OH-. Moreover, some recalcitrant wastewaters are
at a pH value requiring slight or no adjustment for the Fenton oxidation to occur.
Moreover, some recalcitrant wastewaters are at pH levels in the acidic range, as
shown in Table 2.
Table 2: List of some wastewaters at pH <5
Ref.
Wastewater
(Mahiroglu et al., 2009)
(Guedes et al., 2003)
(de Souza et al., 2006)
(Blumenschein 2011)
(Elsayed and Sab 2009)
acid mine drainage-floatation circuit
cork cooking wastewater
chip board
coal mine water
acidic industrial wastewater
pH
level
4.8
4.3
4.0
2.7
1.6
4.
ISSUES RELATED TO FENTON-TREATED WASTEWATERS
4.1
Residual total iron
The presence of soluble iron at concentrations above those acceptable by
environmental laws is a drawback which implies the possible adoption of further
expensive unit operation to reduce the concentration (de Souza et al., 2006). There
is no fixed limit for total iron; however, the maximum limits prescribed by
various environmental regulations are presented in Table 3.
Although there is wide variation in the allowable soluble limits of between
5-15 mg/L, none of the limits approximate values that may be detrimental to
biological activities. According to Guedes et al. (2003), total soluble iron up to
values of 100 mg/L is not inhibitory to activated sludge systems.
Table 3. Various standards expressing maximum discharge limits for total iron in
treated wastewater
Ref.
Standard
(Benatti et al., 2006)
(Santos et al., 2010)
(Hsueh et al., 2005)
(www.doe.gov.my)
(de Souza et al., 2006)
(Martins et al., 2010)
Brazil
Spain
Taiwan
Malaysia (Standard B)
Brazil (Class 3)
Portugal
Max allowable limits
(mg/L)
15
10
10
5
5
2
Ways of ensuring compliance to the limit include the use of low Fe2+
concentrations, for instance approximately twice the limit for class 3 waters as
reported by de Souza et al., (de Souza et al., 2006), or the maximum concentration
of iron cations permitted for the discharge of industrial wastewater in the region
(Santos et al., 2010). Others have attained discharges with iron contents below the
allowable limits by optimisation of the ratio of the oxidant to the catalyst (Benatti
et al., 2006). This approach ensures that excess Fe2+ is not unnecessarily
introduced into the reaction mixture, leading to the radical scavenging effect
shown in equation (22) (Martins et al., 2010) and contributing to decreased
efficiency and unutilised catalyst in the final treated effluent.
Fe2+ + •OH → Fe3+ + OH −
(22)
Table 4: Total soluble iron content of some CFO treated effluents
Ref.
Wastewater
(Shukla et al., 2010)
(Martins et al., 2010)
(Benatti et al., 2006)
(Guedes et al., 2003)
(Santos et al., 2010)
phenolic contaminants
cheese production wastewaters
chemical laboratory wastewater
cork cooking wastewater
aqueous solution of phenol
Effluent total iron
(mg/L)
undetectable
1.9
4.4
5.5
> 10
Although it appears that despite the use of alternate catalysts to the
conventional ferric iron, different oxidants and metal-chelating agents, the CFO
still shows remarkable performance. A similar observation was made by Benatti
et al. (2006) where they highlighted that these options are stoichiometrically
inefficient in comparison to the CFO operating conditions.
4.2
Sludge generation in Fenton oxidation
One of the identified demerits of the Fenton oxidation is the generation of sludge
(Pignatello et al., 2006; Shukla et al., 2010); approximately 1-2 Euro/m3 of
wastewater is spent on sludge treatment (Di Iaconi et al., 2010). However, few
studies have dealt with the quantification and characterisation of Fenton sludge
(Guedes et al., 2003; Dewil et al., 2005; Liu et al., 2007; Atmaca 2009; Mahiroglu
et al., 2009). Sludge generation in Fenton oxidation is expected as the catalyst
possesses coagulating potentials (Arslan and Balcioglu 1999) and the catalyst in
the form of Fe3+ precipitates during the neutralising step (Kiril Mert et al., 2010).
A consequence of this is sludge formation in the form of iron oxyhydroxides
(Umar et al., 2010). Unlike in biological processes where sludge generation is
high and the sludge treatment reported to cost approximately 35-50% of the total
operating costs of wastewater treatment, in Fenton oxidation, sludge generation is
minimal (Dewil et al., 2005). Hsueh et al. (2005) studied the investigation of
Fenton and Fenton-like reactions at low iron concentrations to oxidise three
commercial azo dyes, namely Red MX-5B, Reactive Black 5 and Orange G. Only
a small amount of sludge was produced which met the local effluent standards for
disposal, and thus required no further treatment.
Moreover, oxidation processes have been known to positively affect
sludge generation. For instance, Di Iaconi et al. (2010) reported saving 60% of the
operating costs of conventional biological treatment by reducing sludge
generation by 25-30 times through recirculating the biologically treated
wastewater. Dewil et al. (2005) studied the Fenton sludge and proposed the
Fenton oxidation as a form of an advanced sludge treatment (AST) process for
improving sludge settling and other mechanical properties (Dewil et al., 2005).
Moreover, literature reports on Fenton sludge characterisation indicate that these
sludges are easily formed and handled, and are of good quality (Mahiroglu et al.,
2009).
Generally, Fenton sludges have low settled sludge volume (SSV) and
sludge settling index (SVI) values, which suggests ease of settleability and
dewaterablity with high density and highest sludge settling velocity (vs) thus no
further conditioning of the sludge is required (Mahiroglu et al., 2009). The
volume of produced sludge ranges between 40 and 180 mL/L of treated
wastewater (Guedes et al., 2003; Benatti et al., 2006; Mahiroglu et al., 2009).
Studies on Fenton sludge treatment indicate favourable treatment of the sludge.
For sludge from biological treatment plants, Fenton oxidation has reportedly been
used to effectively improve sludge properties such cake dewaterability (Dewil et
al., 2005). In another work, Pérez et al. (2002) demonstrated that Fenton treatment
improves sludge dewaterability. Exhaustive studies of the effect of the most
significant oxidation operating parameters were conducted by Neyens et al. (2002,
2003). The results of their findings can be summarised as: (i) considerable
reduction of dry solids content (DS) and organic dry solids content (ODS) in the
filter cake of approximately 20% and (ii) an improved dewaterability with a 30%
reduction in the sludge volume, and a 30% increase in the cake DS content in
comparison to untreated sludge samples. These characteristics allow for the
convenient storage and possible use of these sludges as secondary fuel (Baeyens
and Van Puyvelde 1994).
The fast settling rate exhibited by Fenton sludge negates the need for
assisting the settling process. Besra et al. (2003) reported increasing the settling
rate by 10-fold with the addition of polyacrylamide (PAM-C) as a cationic
flocculant. However, such addition and mode of application in the presence of
surfactants led to complex interactions and may lead to precipitate formation in
the mixture (Besra et al., 2003).
The Fenton sludge has proven to be effective when recycled back into the
treatment system as it does not add to the organic loading of the coagulation step
(Yoo et al., 2001). Thus, the sludge potentially aids in organic treatment since
additional organic load reduction is achieved at the coagulation stage (Umar et al.,
2010). Thus, the Fenton process does not fall under the category of wastewater
treatment technologies that need revaluation in the current drive aiming to replace
technologies generating sludge (Di Iaconi et al., 2010).
5.
CONCLUSIONS
The classical Fenton oxidation process has the advantages of being non-selective
and effective in the treatment of recalcitrant wastewater. However, the method
may not be deemed to be attractive for application in the case of a large-scale
chemical process due to the use of costly oxidants and the supposed generation of
an appreciable amount of sludge. These drawbacks may apparently pose severe
limitations to the success of this process in wider applications in the face of
emerging restrictive legislature on sludge disposal and economic assessment of
wastewater treatment processes. In this review, studies in the literature on these
potential factors that may mitigate and limit the application of classical Fenton
oxidation were analysed.
This review has shown that various feasible alternatives are available for
generating the oxidant, ranging from in situ generation, the use of persulphate
(PS), peroxydisulphate (PDS) and peroxymonosulphate (PMS) and wastewater
containing H2O2. The flexibility of the CFO method to accommodate the various
options of utilising the oxidant further makes the CFO process attractive.
Some streams require no pH modulation as the wastewater is in an acidic
condition and the recycled sludge can be utilised to adjust the pH as it produces
hydroxyl ions (OH-), thereby increasing the pH value of the wastewater. In terms
of the residual iron content in treated wastewater, the reviewed reports suggest
that the effluent meets the local standards of various regional limits imposed by
respective regulatory bodies. Meanwhile, this review has established that the
limitation phenomenon of the resistance of certain compounds to the CFO process
is equally common with the assisted Fenton oxidation processes as well as other
AOPs.
Sludge treatment is reported to cost approximately 35-50% of the total
operating costs of wastewater treatment. From the currently available reported
studies on sludge quantification, the range was found to be between 40 and 180
mL/L of treated wastewater. The sludge was also found to easily settle, easily
dewatered with a high density and settling velocity, which requires no further
conditioning. However, the sludge generation and characterisation is not reported
in the bulk of CFO research. Thus, further researches on CFO would
quantitatively establish the sludge generation status vis-a-vis it constituting a
drawback to the process.
Furthermore, as the CFO sludge is proven to be effective when recycled
back into the treatment system (as it does not add to the organic loading during
the coagulation step), the spent sludge should be evaluated for the possibility of
extracting some energy content by oxidizing them as fuel in boilers and furnaces.
In addition, the sludge analysis should also be extended to hybrid CFO systems as
post-oxidation treatment of biological wastewater has been shown to reduce
sludge generation by 25-30 fold and results in saving up to 60% of the operating
costs.
Acknowledgements
The authors thank the Ministry of Higher Education Malaya (MOHE) and the
University of Malaya (UM), Malaysia for financially supporting this study (Grant
No.: UM.C/HIR/MOHE/ENG/37.
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