Eliminarea Poluantilor Organici Persistenti din Apa Prin Procedee de Oxidare Avansata. Studiu de Caz Eliminarea 2,4 Diclorofenolului Prin Fotocaliza Eterogena

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Eliminarea Poluantilor Organici Persistenti din Apa Prin Procedee de Oxidare Avansata. Studiu de Caz Eliminarea 2,4 Diclorofenolului Prin Fotocaliza Eterogena

Byadmin ianuarie 1, 2024

Eliminarea poluanților organici persistenți din apă prin procedee de oxidare avansată.

Studiu de caz:

Eliminarea 2,4 Diclorofenolului prin fotocataliza eterogenă

Cuprins:

Introducere

I.1. Chlorophenolii în mediul înconjurător

I.2. Proprietățile clorofenolilor

I.3. Procedee de oxidare avansată

I.4. Fotocataliza eterogenă

I.4.1.Principiile de bază ale fotocatalizei

I.4.2. Tipuri de fotocatalizatori

I.4.3. Dioxidul de titan

I.5. Influiența principalilor parametrii

I.5.1. Concentrația de catalizor

I.5.2. Lungimea de undă

I.5.3. Concentrația inițială de poluant

I.5.4. Temperatura

I.5.5. Intensitatea fluxului luminos

I.5.6. Quantum yield

I.5.7. Influence of oxygen pressure

I.5.8. Influiența pH-ului

II.1. Materiale și metode

II.1.1. 2,4 Diclorofenolul

II.1.2. Media fotocatalitică

II.1.3. Procedura experimentală

II.1.4. Materiale folosite pentru prelevarea eșantioanelor

II.2. Metode analitice

II.2.1. Măsurarea pH-ului

II.2.2. Măsurarea intensității luminoase

II.2.3. Determinarea carbonului organic total

II.2.4. Determinarea 2,4 diclorofenolului prin HPLC

III. Rezultate și discuții

III.1. Adsorpția 2,4 diclorofenolului

III.2. Fotoliza directă

III.3. Efectul concentrației inițiale de 2,4-dichlorophenol la flux luminos maxim

III.4. Efectul concentrației de catalizor la flux luminos maxim (306 W/m2)

III.5. Efectul variației fluxului luminos

III.6. Efectul variației pH-ului la flux maxim luminos

IV. Concluzii

Bibliografie

Introduction

In recent decades, the discharge of large quantities of synthetic chemicals, such as solvents, plasticizers, insecticides, herbicides, and fungicides, into the environment through industrial, agricultural, medical, and domestic activities has produced significant ecotoxicological problems with serious consequences for all living organisms. These substances include chlorophenols, which have been classified as priority pollutants by the US Environmental Protection Agency [1].

Chlorophenols are of serious environmental concern because of their widespread occurrence throughout the environment. These pollutants are present in wastewater, sludge products, surface waters, and groundwater [1]. A number of physical, chemical and biological methods have been studied to eliminate chlorophenols from industrial effluents and neither of these methods and their combinations has been used to achieve complete mineralization of chlorophenols [2].

As a response, the development of newer eco-friendly methods of destroying these pollutants became an imperative task. Ultimately, research activities centred on advanced oxidation processes (AOPs) for the destruction of synthetic organic species resistant to conventional methods. AOPs rely on in situ generation of highly reactive radical species, mainly HO• by using solar, chemical or other forms of energy [3].

Among AOPs, heterogeneous photocatalysis has proved to be of real interest as efficient tool for degrading both aquatic and atmospheric organic contaminants [3].

Heterogeneous photocatalysis involve the acceleration of photoreaction in presence of semiconductor photocatalyst. One of the major applications of heterogeneous catalysis is photocatalytic oxidation (PCO) to effect partial or total mineralisation of gas phase or liquid phase contaminants. Even though degradation begins with a partial degradation, the term ‘photocatalytic degradation’ usually refers to complete photocatalytic oxidation or photomineralisation, essentially to CO2, H2O, NO3−, PO43− and halide ions [4].

In this work, we investigated the performance of TiO2 catalysts in the photocatalytic of 2,4- dichlorophenol under UV irradiation. Effects of initial concentration of the pollutant, the radiant flux applied on the catalyst concentration and the degradation of the target compound were studied.

I.I. Chlorophenols in the natural environment

Investigations into the state of the environmental pollution, carried out across the world, confirm the presence of chlorophenols in many ecosystems: surface and ground waters, bottom sediments, atmospheric air and soils [5].

Under some circumstances chloroderivatives of phenol may become substrates for the formation of polychlorinated biphenylenes and dioxins [6].

Results of the study carried out by Wagner and al. (1990) showed an enzymatic oxidation of 2,4,5-trichlorophenol in the presence of hydrogen peroxide and a culture filtrate of the soil fungus Phanerochaete chrysosporium. Studies of Svenson and al. (1989), Oberg and al. (1990) and Oberg and Rappe (1992) also reported an enzymatic conversion of chlorophenols in the presence of horseradish peroxidase enzyme and other peroxidases [5].

In 1994, Vollmuth and al. proposed a photochemical mechanism explaining the formation of dioxins during irradiation of water solutions containing PCP with UV rays. Klan and al. (2001) showed that hydroxychlorobiphenyl and phenol might be formed by irradiating 3-chlorophenol present in ice with UV light [5]. A scheme of the reaction is presented in figure I.4.

Figure I.1. Photodegradation of 3-chlorophenol in ice (Klan and al., 2001).

In the aquatic environment, chlorophenols exist as dissociated, non-dissociated or adsorbed onto suspended matter. The forms of occurrence of chlorophenols depend on pH of the environment, as well as on physical and chemical properties of the particular compounds [5,6].

Table I.1. presents chlorophenols concentrations reported in the environment from a number of studies. While the table is not exhaustive, it provides a cross-section of the range of samples and concentrations reported in the literature [5].

Table I.1. Chlorophenols concentration detected in different natural environments

I.2. Properties of chlorophenols

Chlorophenols are chlorinated aromatic ring structures consisting of the benzene ring, –OH group and atom(s) of chlorine [2].

Jointly with the 19 possible isomers, chloroderivatives of methyl- and ethyl-phenols are also considered as chlorophenols. All chlorophenols are solids at room temperature except 2-chlorophenol (2-CP), which is a liquid. The aqueous solubility of chlorophenols is low, but the sodium or potassium salts of chlorophenols are up to four orders of magnitude more soluble in water than the parent compounds [2].

The acidity of chlorophenols increases as the number of chlorine substitutions increases. The n-octanol/water partition coefficients (Kow) of chlorophenols increase with chlorination, indicating a propensity for the higher chlorophenols to bio-accumulate [6]. The partition of an organic pollutant between the water and organic phases is generally correlated with various properties, such as the water solubility (S) and the octanol/water partition coefficient (Kow) [6]. The environmental fate of organic compounds is also correlated with the organic carbon partition coefficient (Koc) [5].

Chlorophenols are subject to a series of physical, chemical, and biological transformations. Sorption, volatilization, degradation, and leaching are the primary processes governing their fate and transport. The pH in water, soil and sediment is a major factor affecting the fate and transport of chlorophenols in these media, since the degree to which the compounds ionize increases with increasing pH [5,6].

In addition, physicochemical properties of chlorophenols such as water solubility, Henry’s law constant, organic carbon sorption coefficient, volatilization rate, and photolysis rate determine transport processes. Important environmental parameters influencing these processes include organic matter content and clay content in soil, sediment, and water, as chlorophenols are in general preferentially adsorbed to these soil constituents [7].

Generally, as the number of chlorine molecules increase, there is a reduction in vapor pressure, an increase in boiling point, and a reduction in water solubility of the chlorophenols [8]. Therefore, increasing chlorination increases the tendency of these compounds to partition into sediments and lipids and to bioconcentrate [7].

Physical and chemical properties of some selected chlorophenols are shown in Table 1.

Table I.2. Physical- chemical properties of chlorophenol compounds [2,5]

Ocatnol-water partition coefficient (Kow); dissociation constant (pKa); not available (na).

I.3. Advanced oxidation processes

The use of advanced oxidation processes (AOPs) in the treatment of wastewater and contaminated environments has been widely employed for degradation of highly persistent and non-biodegradable pollutants [8].

Generally, the degradation under optimum conditions leads to the mineralization of organic contaminants to CO2, H2O, and ions (chloride, sulfate, and others) [3]. Among the AOPs, is the heterogeneous photocatalytic oxidation using TiO2 as catalyst. It is well established that superoxide radical anions and hydroxyl radicals are the primary species generated after the photogeneration of the electron–hole pair of the semiconductor catalyst in presence of water and air; however, the mechanism associated to this process is complex [8].

Mechanistic studies show that the hydrogen peroxide degradation process is catalyzed by iron ions to produce reactive species such as HO• and HO2• at different stages. The formation of these species is given in Eqs [8].

Initial reactions:

H2O2 → H2O + ½ O2 (1)

H2O2 + Fe2+ → Fe3+ + HO- +HO• (2)

H2O2 + Fe3+ → Fe(OOH)2+ + H+ ↔ Fe2+ + HO•2 + H+ (3)

HO− + Fe3+ → Fe(OH)2+ ↔ Fe2+ + HO• (4)

Propagation:

HO• + H2O2 → HO• 2 + H2O (5)

HO•2 + H2O2 → HO• + H2O + O2 (6)

HO•2 + HO−2 → HO• + HO− + O2 (7)

Termination:

Fe2+ + HO• → Fe3+ + HO− (8)

HO• 2 + Fe3+ → Fe2+ + H+ + O2 (9)

HO• + HO•2 → H2O + O2 (10)

HO• + HO• → H2O2 + O2 (11)

OH· radicals are extraordinarily reactive species, they attack the most part of organic molecules with rate constants usually in the order of 106–109M−1 s−1 [7,8].

They are also characterized by a little selectivity of attack which is a useful attribute for an oxidant used in wastewater treatment and for solving pollution problems [3].

A suitable application of AOP to waste water treatments must consider that they make use of expensive reactants as H2O2, and/or O3 and therefore it is obvious that their application should not replace whenever/possible, the more economic treatments as the biological degradation.

According to Huang and coworkers (1993), the AOP can be classified homogeneous and heterogeneous. Domènech and coworkers (2001) presented this classification in terms of whether light is used in the process (table I.4.) [3].

Table I.4. Types and classification of advanced oxidation processes [3].

I.4. Heterogeneous photocatalysis

I.4.1. Basic principles of photocatalysis

Heterogeneous photocatalysFe2+ + H+ + O2 (9)

HO• + HO•2 → H2O + O2 (10)

HO• + HO• → H2O2 + O2 (11)

OH· radicals are extraordinarily reactive species, they attack the most part of organic molecules with rate constants usually in the order of 106–109M−1 s−1 [7,8].

They are also characterized by a little selectivity of attack which is a useful attribute for an oxidant used in wastewater treatment and for solving pollution problems [3].

A suitable application of AOP to waste water treatments must consider that they make use of expensive reactants as H2O2, and/or O3 and therefore it is obvious that their application should not replace whenever/possible, the more economic treatments as the biological degradation.

According to Huang and coworkers (1993), the AOP can be classified homogeneous and heterogeneous. Domènech and coworkers (2001) presented this classification in terms of whether light is used in the process (table I.4.) [3].

Table I.4. Types and classification of advanced oxidation processes [3].

I.4. Heterogeneous photocatalysis

I.4.1. Basic principles of photocatalysis

Heterogeneous photocatalysis is a discipline which includes a large variety of reactions: organic synthesis, water splitting, photoreduction, hydrogen transfer, O218–O216 and deuterium–alkane isotopic exchange, metal deposition, disinfection and anti-cancer therapy, water detoxification, gaseous pollutant removal, etc. [9,10].

Among these appearances titania-assisted heterogeneous photocatalytic oxidation has received more attention for many years as alternative method for purification of both air and water streams.

This process offers several avantages:

it can be operated easily at room temperatures;

low pressure requirements;

no chemical addition;

complete mineralization of the polluants [10, 11].

Table I.5. Non-exhaustive list of aqueous organic polluants mineralized by photocatalysis [12].

Photocatalytic reaction is initiated when a photoexcited electron is promoted from the filled valence band of semiconductor photocatalyst (SC) to the empty conduction band as the absorbed photon energy, hʋ, equals or exceeds the band gap of the semiconductor photocatalyst leaving behind a hole in the valence band. Thus in concert, electron and hole pair (e− –h+) is generated. The following chain reactions have been widely postulated [10].

Photoexcitation: TiO2/SC + hʋ → e− + h+ (1)

Oxygen ionosorption: (O2)ads + e−→ O2•− (2)

Ionization of water: H2O → OH− + H+ (3)

Protonation of superoxides: O2•− + H+ → HOO• (4)

The hydroperoxyl radical formed in (4) also has scavenging property as O2 thus doubly prolonging the lifetime of photohole [10]:

HOO• + e− → HO2− (5) HOO− + H+ → H2O2 (6)

Both the oxidation and reduction can take place at the surface of the photoexcited semiconductor photocatalyst (Figure I.1.). Recombination between electron and hole occurs unless oxygen is available to scavenge the electrons to form superoxides (O2•−), its protonated form the hydroperoxyl radical (HO•2) and subsequently H2O2 [10].

Figure I.2. Schematic photophysical and photochemical processes over photon activated semiconductor cluster (p) photogeneration of electron/hole pair, (q) surface recombination, (r) recombination in the bulk, (s) diffusion of acceptor and reduction on the surface of SC, and (t) oxidation of donor on the surface of SC particle [10].

As for classical heterogeneous catalysis, the overall process can be decomposed into five independent steps:

1. Transfer of the reactants in the fluid phase to the surface

2. Adsorption of a least one of the reactants

3. Reaction in the adsorbed phase

4. Desorption of the product(s)

5. Removal of the products from the interface region [10].

I.2.2. Types of photocatalysts

Solids that can promote reactions in the presence of light and are not consumed in the overall reaction are referred to as photocatalysts. These are invariably semiconductors. A good photocatalyst should be photoactive, biologically and chemically inert, inexpensive and non-toxic. In order for a semiconductor to be photochemically active as a sensitizer for the above reaction the redox potential of the photogenerated valence band must be sufficiently positive to generate OH• radicals, which can subsequently oxidize the organic pollutant. The redox potential of the photogenerated conductance band electron must be sufficiently negative to be able to reduce adsorbed O2 to superoxide [11].

Si, TiO2, ZnO, WO3, CdS, ZnS, SrTiO3, SnO2, WSe2, Fe2O3, etc can be used as photocatalysts. Table I.6. and figure I.3. give band energies band gap positions of these catalysts. Figure I.2. also gives the redox potentials of the H2O/OH• and O2/HO2• couples [11].

Table I.6. Bandgap energy for various photocatalysts

Oxidation of many pollutants, especially organic species, requires high potentials with result that the valence band location at the semiconductor-electro-lyte interface has to be rather positive, as exemplified by TiO2 and CdS.

Generally, the pollutant is organic and the semiconductor is TiO2. The range of organic pollutants that can be completely photomineralized by oxygen, using TiO2 as the sensitizer, is very wide and includes many aliphatics, aromatics, detergents, dyes, pesticides and herbicides [11].

Figure I.3. Band-edge energies of typical semiconductors, O2/HO2• and H2O/OH• pair

I.2.3. Titanium dioxide

Titanium dioxide is one of the basic materials in everyday life. It has been widely used as white pigment in paints, cosmetics and foodstuffs.

Generally, titanium dioxide is a semiconducting material which can be chemically activated by light. The photoactivity of TiO2 which is known for approximate 60 years is investigated extensively. For a long time there was a considerable problem especially what its application as pigment concerns. Under the influence of light the material tends to decompose organic materials. This effect leads to the well-known phenomenon of “paint chalking”, where the organic components of the paint are decomposed as result of photocatalytic processes [13].

TiO2 exists in three crystalline modifications: rutile, anatase, and brookite (Figure I.4.) [13]. Compared with rutile and brookite, anatase shows the highest photoactivity. In addition, anatase is the most active allotropic form among the various ones available, either natural (rutile and brookite) or artificial (TiO2–B, TiO2–H). Anatase is thermodynamically less stable than rutile, but its formation is kinetically favored at lower temperature (<600 °C). This lower temperature could explain higher surface area, and a higher surface density of active sites for adsorption and for catalysis [13].

Therefore, the TiO2 used in industrial products is almost exclusively from the rutile type.

Figure I.4. structure of the three most common allotropes of TiO2:

anatase (a), rutile (b) and brookite (c).

TiO2 absorbs only approx. 5 % of the solar light reaching the surface of the earth, it is the best investigated semiconductor in the field of chemical conversion and storage of solar energy.

Photocatalysis with TiO2 presents some distinct advantages:

high TiO2 chemical stability in a large pH range,

low cost of catalyst and chemicals,

none or low inhibitions by common ions present in water,

a wide spectrum of organic contaminants can be converted to water and CO2,

relatively mild reaction conditions required (room temperature, atmospheric pressure),

no chemical reactants must be used and no side reactions are produced,

success in the decomposition of several toxic refractory pollutants [13].

I.5. Influence of physical parameters governing the kinetics

I.5.1. Mass of catalyst

Either in static, or in slurry or in dynamic flow photoreactors, the initial rates of reaction were found to be directly proportional to the mass m of catalyst.

It was found equal to 1.3 mg TiO2/cm2 of a fixed bed and to 2.5 mg TiO2/cm3 of suspension. These limits correspond to the maximum amount of TiO2 in which all the particles – i.e., all the surface exposed – are totally illuminated. For higher quantities of catalyst, a screening effect of excess particles occurs, which masks part of the photosensitive surface. For applications, this optimum mass of catalyst has to be chosen in order to avoid excess of catalyst and to ensure a total absorption of efficient photons [12].

I.5.2. Wavelength

The variations of the reaction rate as a function of the wavelength follows the absorption spectrum of the catalyst, with a threshold corresponding to its band gap energy. For TiO2 having EG = 3.02 eV, this requires: λ ≤ 400 nm, i.e., near-UV wavelength (UV-A). In addition, it must be checked that the reactants do not absorb the light to conserve the exclusive photoactivation of the catalyst for a true heterogeneous catalytic regime (no homogeneous nor photochemistry in the adsorbed phase) [12].

I.5.3. Initial concentration in pollutant

Generally, the kinetics follows a Langmuir – Hinshelwood mechanism confirming the heterogeneous catalytic character of the system with the rat r varying proportionally with the coverage Ɵ as:

For diluted solutions (C < 10−3 M), KC becomes « 1 and the reaction is of the apparent first order, whereas for concentrations > 5 x 10−3 M, (KC » 1), the reaction rate is maximum and of the zero order [12].

I.5.4. Temperature

Because of the photonic activation, the photocatalytic systems do not require heating and are operating at room temperature.

The optimum temperature is generally comprised between 20 and 80 °C. This explains why solar devices which use light concentrators require coolers [14]. This absence of heating is attractive for photocatalytic reactions carried out in aqueous media and in particular for environmental purposes (photocatalytic water purification) [12].

I.5.5. Radiant flux

The rate of reaction r is proportional to the radiant flux. This confirms the photo-induced nature of the activation of the catalytic process, with the participation of photo-induced electrical charges (electrons and holes) to the reaction mechanism [12].

I.5.6. Influence of Ph

pH has a higher direct effect on the conversion rate; it can affect either the surface properties of the photocatalyst or the chemical form of the substrate. TiO2 has an amphoteric character with a point of zero charge around pH equal to 6 (5), and the substrate can undergo acid-base equilibria. Consequently, the adsorption of the substrate may be affected, strongly influencing the degradation rate [12].

II. Materials and methods

II.1.1. 2,4-Dichlorophenol

2,4-Dichlorophenol (DCP) is an important intermediate in the manufacture of 2,4- dichloro-phenoxyacetic acid (2,4-D), the well-know industrial commodity herbicide. It is also used in the manufacture of other pesticide products and pharmaceuticals [14].

Table II.1.1. Physical and chemical properties of 2,4-Dichlorophenol

For the heterogeneous photocatalysis experiments 2,4-Dichlorophenol of analytical purity (> 99%) was used. It was provided by Sigma-Aldrich® (France).

Figure II.1. 2,4-Dichlorophenol used for photocatalytic kinetic study

II.1.2. Photocatalytic media

To study the kinetics of catalytic degradation of the target compound was used in the titanium dioxide powder, Millenium PC-500 supplied by Millennium Chemicals. The main features of the catalyst used are shown below (Table II.1.2.).

Table II.1.2. Physical and chemical properties of TiO2 used in this work

II.1.3. Experimental procedures

The pilot

The device used to study the photocatalytic degradation of 2,4 dichlorophenol is a pilot rectangular with the following dimensions: length = 100.2 cm and width = 40.2 cm (Figure II.1.).

This type of pilot is an opaque enclosure is divided into three compartments, as follows:

The central compartment or batch reactor (a glass beaker) 3L capacity lies. TiO2 is suspended in the solution with a metal stirrer with a small diameter provided with four straight blades. The stirring speed of the stirrer is adjustable and may vary from 40 to 1200 rev / min. The selected speed during our experiments was 555 rev / min.

The different samples are taken of the solution without opening the enclosure, using a plastic pipe (diameter 3 mm) which plunges directly into the solution. The volume of the sample is adjusted using a plastic syringe with a volume of 12 ml.

In the two others compartments of the pilot is transmitted ultraviolet radiation (UV), artificially produced by 2 UVA lamps, situated on either side of the central reactor. UV rays irradiate the solution to treat the beaker placed through two openings square. The UV lamps are fastened on a rod whose position relative to the reactor can be adjusted by means of screws and nuts according to the wish of the operator. To prevent evaporation of the solution to be treated during our manipulations in, a Plexiglas cover was used on the reactor.

Figure II.2. The reactor used in this work for the study of the degradation kinetics of 2,4-DCP.

Step 1: Preparation of the solution to be treated

For the preparation of the solution to be treated and the quantities of 2,4 dichlorophenol and TiO2 required for each trial were weighted on a precision balance (Mettler AE 160). In a flask (V = 2000 ml) which contain 1000 ml ultra-pure water is dissolved the mass of 2,4 dichlorophenol weighed and after, the volume is complete with ultra-pure water until to sign.

During the study on photocatalytic degradation kinetics of 2,4-DCP, the ultra-pure water used was produced by a device Elga Option-Q DV 25 (figure II.3.). It has the following characteristics: 18.2MΩ.cm resistivity at 298 K, TOC < 50 μg C/L.

Then the solution thus prepared was placed on a magnetic stirrer for 24 hours for to be completely solubilized because 2,4-DCP has a low solubility in water.

It should be noted that for to avoid the photodegradation of 2,4 dichlorophenol in the presence of light during the step of agitating, the flask containing the solution was wrapped in an aluminum foil (figure II.3.). Also, before to transfer the solution in the pilot, was measured the initial pH, as is described in the paragraph II.2.1.

Figure II.3. Preparation flask containing solution analyzed for magnetic stirring

Step 2: Implementation of adsorption experiments

To start the adsorption process, we transferred the solution thus prepared in the reactor, and then the mechanical stirrer was turned on. A sample was immediately carried out to know the concentration at time t = 0 (beginning of the adsorption kinetics). Then, the catalyst (the mass of each specific test as mentioned in the results section of this report) was added to the reactor. To ensure that the steady state is reached in the reactor, the solution is left in contact with the catalyst in the dark for 60-70 minutes.

Note that this term has been determined experimentally. Then the kinetics of photocatalytic degradation can start.

Step 3: Photocatalytic degradation experiments

At first, the distance of the lamps inside the reactor was adjusted (depending on the value of the radial flux incident desired), then the two square slots ensuring the passage of UV rays to the surface of the reactor were sealed. Lamps were lit and preheated for 30 minutes to reach their maximum intensity. Then the two slots ensuring the passage of light have been opened, that is the time t = 0 of the photocatalytic degradation kinetics.

Samples are taken for 3h or 4h (180 and respectively, 240 minutes) as follows: every 5 minutes for 30 minutes and then every 30 minutes. On these samples several analyzes were performed as follows: the determination of the concentration of 2,4-DCP by HPLC/DAD, the TOC concentration and pH. For each experiment the luminous flux incident provided was measured using a radiometer (VLX 3W, ), as described in section II.3.4.

For each experiment the operating conditions used were as follows: the pH of 5.9, the room temperature of 24 °C and a stirring speed of 555 rev / min.

II.1.4. Materials used for the sampling

Samples with a volume of approximately 2 ml or 20 ml were performed manually as appropriate, using a 12 ml syringe and a sampling tube. The sample was then filtered to separate the TiO2. Various analyzes were performed on the filtrate result, namely the measurement of total organic carbon and concentration of 2,4-DCP by HPLC. For each filtration was used a filter with a porosity of 0.45 µm and the diameter of 25 mm, made of polyester provided by CHROMAFIL® XTRA PET – Germany.

Figure II.4. Materials used for the sampling sample and its filtration

II.2. Analytical methods

II.2.1. pH measurement

To measure the initial pH of the solution to be treated and for determining the change in pH during the degradation kinetics a pH meter was used.

As a reminder the operation of a pH meter is based on the relationship between the concentration of H3O+ (definition of pH) and electrochemical potential difference that is establishe in the pH- meter once immersed in test solution. In the figure below shows the main elements of a pH meter.

1 – Body of the glass electrode

2 – KCl crystals (visible in some cases)

3 – Saturated KCl solution

4 – Calomel electrode (or Ag / AgCl)

5 – Active Glass

6 – Reference calomel electrode (or Ag / AgCl).

Figure II.2.1. Schematic presentation of a pH-meter

Calibration of the pH meter in 3 points

The pH-meter, used in this study, was calibrated with buffer solutions of known pH. For more precise measurements, a three buffer solution calibration was preferred.

The first step in calibration is to put electrode intro pH 7 buffer. This pH 7 is essentially, a "zero point" calibration (akin to zeroing or tarring a scale or balance), calibrating at pH 7 first, calibrating at the pH closest to the point of interest (e.g. either 4 or 10) second and checking the third point will provide a more linear accuracy to what is essentially a non-linear problem. According to the measures that have been performed, the device is being calibrated with a solution of pH 7 and then with a solution of pH 4 to make measurements in an acidic medium and then with a solution of pH 10 for measurements in a basic medium.

It’s important, to rinse the electrode with distilled water from a wash bottle into an empty beaker before immersing it into new solution. It must do it every time when the electrode is moved from one solution to other to minimize contamination. Indicated, to use fresh pH buffer solutions for the most accurate results.

Figure II.2.2. pH meter used in this work

II.2.2. Radiant flux measurement

For the measurement of the energy of the electromagnetic radiation emitted by UV lamps, has been made a radiometer. The detectors used in the radiometry transform radiant energy into an electrical signal whose intensity is function of the power of the incident radiation (Braun et al. 1986).

In our case, the radiometer use is a VLX-3W (Vilbert Lourmat, Marne La Vallée, France), with a cell CX-365, measuring the energy flux emitted by the lamp at 365 nm. The measurement range used by this radiometer is between 355 and 375 nm.

In experimental conditions, it was impossible to directly measure the light intensity of the radiometer, the surface of the catalyst in the reactor because the sensor can not be placed in water and in addition has a thickness of 1 cm.

Figure II.2.3. Radiometer used in our study

II.2.3. Determination of Total Organic Carbon (TOC)

General discussion

The organic carbon in water and wastewater is composed of a variety of organic compounds in various oxidation states. Some of these carbon compounds can be oxidized further by biological or chemical processes, and the biochemical oxygen demand (BOD), assimilable organic carbon (AOC), and chemical oxygen demand (COD) methods may be used to characterize these fractions.

Measurement of TOC is of vital importance to the operation of water treatment and waste treatment plants. Drinking water TOCs range from less than 100 µg/L to more than 25,000 µg/L. Wastewater may contain very high level of organic compounds (TOC > 100 mg/L).

To determine the quantity of organically bound carbon, the organic molecules must be broken down and converted to a single molecular form that can be quantitatively. TOC methods utilize high temperature, catalysts, and oxygen, or lower temperatures (>100°C) with ultraviolet irradiation, chemical oxidants, or combinations of these oxidants to convert organic carbon to carbon dioxide (CO2). The CO2 may be purged from the sample, dried and transferred with a carrier gas to a no dispersive infrared analyzer or colorimetric titration. Alternatively, it may be separated from the sample liquid phase by a membrane selective to CO2 into a high-purity water in which corresponding increase in conductivity is related to the CO2 passing the membrane.

Fractions of total carbon

The methods and instruments used in measuring TOC analyze fractions of total carbon (TC) and measure TOC by two or more determinations. These fractions of total carbon are defined as:

inorganic carbon – the carbonate, bicarbonate and dissolved CO2;

total organic carbon (TOC) – all carbon atoms covalently bonded in organic molecules;

dissolved organic carbon (DOC) – the fraction of TOC that passes through a 0.45 µm pore -diam filter;

Suspended organic carbon-also referred to as particulate organic carbon, the fraction of TOC retained by a 0.45 µm filter;

purgeable organic carbon – also referred to as volatile organic carbon, the fraction of TOC removed from an aqueous solution by gas stripping under specified conditions;

nonpurgeable organic carbon – the fraction of TOC not removed by gas stripping.

Scheme II.2.1. Different forms of the organic carbon present in surface water

Note: Minimum detectable concentration – 1 mg C/L or less, depending on the instrument used. This can be achieved with most high-temperature combustion analyzers although instrument performance varies. The minimum detectable concentration may be concentrating the sample, or by increasing the portion taken for analysis.

II.2.2. Description of the method used in this study for the TOC determination

Throughout of this study, the amount of carbon dioxide produced during illumination of aqueous-solution, was determined by measuring the inorganic carbon content using a Total Organic Carbon Analyser model, TOC-VCPH/CPN, provided with an auto-sampler ASI-V, purchased from Shimadzu Corporation – Japan (figure II.2.4.). For the calibration standard was used potassium phthalate solution (as described in the paragraph below).

It should be noted that the TOC-VCPH/TOC-VCPN (hereinafter referred to as TOC-V) instrument measures the amount of total carbon (TC), inorganic carbon (IC) and total organic carbon (TOC) in water.

Oxidative combustion – infrared analysis is a widely-used TOC measurement method that has been adopted by the international standards. The TOC-V instrument can also measure total water-borne nitrogen (TN) if the optional TNM-1 is installed. TN is measured using the principles of “oxidative combustion-chemiluminescence”.

Figure II.2.4. Apparatus used for the determination of TOC

Oxidative combustion infrared analysis (or high-temperature combustion) is a method used for a wide variety of samples, but its utility is dependent on particle size reduction because it uses small-orifice syringes.

Principle:

The sample is homogenized and diluted as necessary and a microportion is injected into a heated reaction chamber packed with an oxidative catalyst such as cobalt oxide, platinum group metals, or barium chromate. The water is vaporized and the organic carbon is oxidized to CO2 and H2O. The CO2 from oxidation of organic and inorganic carbon is transported in the carrier – gaz streams and is measured by means of nondispersive infrared analyzer, or titrated coulometrically.

Because total organic carbon is measured, inorganic carbon must be removed by acidification and sparging or measured separately and TOC obtained by difference.

Measure inorganic carbon by injecting the sample into a reaction chamber where it is acidified. Under acidic conditions, all inorganic carbon is converted to CO2, which is transferred to the detector and measured. Under these conditions organic carbon is not oxidized and only inorganic carbon is measured.

Alternatively, convert inorganic carbonates to CO2 with acid and remove the CO2 by purging before sample injection. The sample contains only the nonpurgeable organic carbon fraction of total carbon: a purgeable organic carbon determination also is necessary to measure TOC.

Preparation of standard curve for the total carbon analysis by TOC meter

The volume of sample required for analysis is 20 ml.

The duration of an analysis is about 30 minutes. The measurement uncertainty is 0.05 mg C / L and the limit of quantification is 0.2 mg C / L.

Description of the steps used for the preparation of the calibration curve:

Placement of the solid potassium hydrogen phthalate powder in an aluminum cup and then in an oven for to be dry and after, cool the cup in a desiccator (each time before weighing), to remove all traces of water from the product.

Weigh accurately 106.4 mg of anhydrous powder and then dissolve with ultrapure water (UPW) in a volumetric flask of 500 mL. The solution obtained (S100) is 100 mg C / L. One mole of KHC8H4O4 mass 204.22 g contains 8 moles of carbon atoms of 96 g mass, so that 106.4 mg containing 50 mg of compound C in 500 mL.

Take 10 ml (S100), transfer to a 100 mL volumetric flask and make up with UPW to obtain a solution (S10) of 10 mg C / L.

Prepare, in the vials of 20 mL, solutions the calibration range, as shown in the following tables:

The vials for the measurement of COD are washed with sulfuric acid if necessary, rinsed with UPW placed in a stove to dry thoroughly.

A vial labeled S0 UPW filled and the others from S2 to S10 are completed by filtration through 0.45 µm by solution of the same name.

Nine of these vials are labeled from 1 to 9 and completed by filtration through 0.45 µm as shown in the following table:

Enter the values of the standard concentrations in two ranges:

Range 1: 0 – 1 – 5 – 9 and 10 mg COD / L;

Range 2: 0 – 20 – 40 – 60 to 80 and 100 mg COD / L.

Table II.2.1. Preparation of standard solutions used for calibration courve

Note: Storage of standard solutions – the standard solutions undergo concentration changes, particularly when low-concentration solutions are stored even for short periods. As a result, high-concentration standard stock solutions (for example, 1000 mgC/L) should be stored in airtight containers in a cool, dark place. Glass bottles are suitable storage containers. Dilute the stock solution prior to each use.

Storage time – The limitation on storage of standard solutions is about 2 months for 1000 mgC/L standard stock solutions and about 1 week for diluted standard solutions (for example, 100 mgC/L). The limitations are for cold storage in sealed containers.

Figure II.2.5. The vial used for analysis TOC

II.2.4. Determination of 2,4 dichlorphenol by High-Performance Liquid Chromatographic (HPLC)

Principle: HPLC is an analytical technique in which a liquid mobile phase transports a sample through a column containing a liquid stationary phase. The interaction of the sample with the stationary phase selectively retains individual compounds and permits separation of sample components. Detection of the separated sample compounds is achieved mainly through the use of absorbance detectors for organic compounds and through conductivity and electrochemical detectors for metal and inorganic components.

Equipment for high performance liquid chromatography

The equipment used to operate HPLC columns can be divided into two main areas, the delivery of solvent and sample to the column and the detection of molecules emerging from the column.

Close regulation of the flow rate of solvent through the column and of the composition of the solvent in gradient procedures is achieved by microprocessor feedback systems controlling the pumps. Together with the use of a sample injection valve this makes any HPLC method highly reproducible. To identify the appearance of the various molecules being separated as they emerge from the column a detector is often attached "in line" with the column outlet. Usually detectors monitor the absorbance or fluorescence of the emerging molecules, but other methods may sometimes be used. In line monitoring enables an immediate and continuous record of the separation achieved to be recorded on paper by a chart recorder.

2,4-DCP analysis by HPLC/DAD

The degradation of 2,4-DCP was monitored by Waters HPLC system (). This unit was equipped with a pump (model 600), an injector Waters 717 plus Autosampler.

The UV detector was WatersTM 996 Photodiote Array and the stationary phase was a Symetri® C18 column from the same manufacturer, with the following dimensions: 4.6 mm × 250 mm i.d., and particle size of 5 µm.

The mobile phase consisting of methanol, pure water and formic acid (60:40:1, v/v/v) was applied at the flow rate of 1 mL/min. and the injected sample volume was 50 µL. Elution is performed in isocratic mode.

Note: Before using the mobile phase in the HPLC system, it was filtered with a 0.45 µm filter and then was degassed, for 5 min. in a sonic bath (as shown in figure II.2.6).

Figure II.2.6. The mobile phase filter system and sonic bath used

The retention time (tR) was obtained for 2,4 dichlorophenol at 6.3 minutes and the detection was made at 287 nm.

A calibration range was made with external standards, each containing six different concentrations (between 25 and 400 mg/L), how is described in preparation protocol and the equation of the calibration curve which are presented in tables II.2.3. and II.2.4.

Figure II.2.7. Waters HPLC system used for the determination and identification of 2,4 DCP

During the photocatalytic degradation kinetics was observed occurrence of mini peaks (byproducts) on the chromatogram of the compound studied.

Identification of these byproducts was achieved by injection into the same HPLC system, standard solutions of 2-CP, 4-CP and clorohydroquinone. The mobile phase used was the same as for 2,4-DCP. Retention times and wavelengths at which these compounds were identified are given bellow.

Table II.2.2. Retention times and the wavelengths of some byproducts which appear during photocatalytic degradation of 2,4 dichlorophenol.

Preparation of standard curve for determination of 2,4 DCP by HPLC / DAD

Wish to prepare a stock solution of 500 mg / L 2,4 DCP, in methanol;

Choose a volumetric flask of 250 mL, half containing methanol, weigh on analytical balance 125 mg of 2,4 DCP, add volumetric flask and fill up to the mark with methanol. We will call this solution standard SM;

Prepare in volumetric flasks of 50 mL, the range of the calibration solutions as indicated below:

Table II.2.3. Preparation of standards for analysis by HPLC/PDA detector

From each elutions made is filled a vial of 1 mL, for HPLC analysis. The procedure of filling is performed using a micropipette.

Depending on the concentration of each elution and peak area calibration curve can be drawn.

Table II.2.4. Equation of the calibration curve for the determination of 2,4 DCP by HPLC

III. Results and discussion

It is well known that the efficiency of the photocatalytic degradation of organic compounds depends on several factors, such as: the initial concentration of pollutant, photocatalyst concentration, radiation intensity, pH or temperature.

Thus, we performed a series of experiments focused to determine the optimum operating parameters for photocatalytic degradation of 2,4-DCP.

III.1. Adsorption of 2,4 DCP using TiO2, without UV irradiation

All experiments were started with adsorption, before starting the study of the kinetics of degradation of the pollutant considered preliminary tests were conducted to determine the amount of pollutants adsorbed on the catalyst surface at equilibrium qe (expressed in mg 2,4 DCP/gTiO2).

This experiments were carried out in the dark, under the identical conditions as those of the degradation kinetics experimental conditions (concentration of TiO2 and 2,4-DCP, pH, temperature, speed of agitation).

Value of absorbance qe was calculated using the following equation:

qs = (C0 – Cs)*V / CTiO2

In which: C0 and Cs represent the initial concentrations and respectively, the equilibrium of the pollutant in solution (mg/L), V represent volume of solution used, and CTiO2 represent the catalyst amount (in g/L of solution).

It was observed that the adsorbed amount increases with agitation time. It was also demonstrated that for all tests the kinetic equilibrium adsorption of 2,4 DCP is reached after approximately 60 minutes. The percentage of degradation was about 2%.

III.2. Direct photolysis

The UV irradiation of 2,4-DCP in the absence of TiO2 did not show observable evidence of degradation. Therefore, the direct photodegradation could be disregarded under the experimental conditions here employed.

In the literature is found that direct photolysis of 2,4-DCP in the aqueous environment leads to formation of 4-chlorocyclopentadienyl carboxy acid, chlorohydroquinone or 4-chloro-1,2-benzenediol (Figure III.1.).

Pandiyan et al. investigating the photolysis of 2,4-DCP using GC/MS, identified monochlorophenol as the main product [15].

Hirvonen et al. investigating photo-oxidation reactions using UV/H2O2 methods forwavelenght λ = 254, observed the formation of numerous hydroxylated intermediates such as: trichlorohydroxybiphenyl, tetrachlorodihydroxybiphenyls, tetrachlorotrihydroxybiphenyl and dichlorobenzenediols [15].

The authors suggested that the presence of dichlorobenzenediol in solution after photo-oxidation is the result of a hydroxyl group attached to the aromatic ring without dechlorination. The chlorohydroxybiphenyls are formed by dimerization of organic radicals or from their reactions with the parent chlorophenols [15].

Figure III.1. Photo-degradation pathway of 2,4 DCP

III.3. Effect of initial concentration of 2,4-Dichlorophenol

Degradation of 2,4-DCP in aqueous medium by TiO2/UV photocatalytic treatments has been investigated.

For this, different initial 2,4-DCP concentrations was employed to investigate how the initial pollutant concentrations affects the efficiency of 2,4-DCP removal. Was varied in a range between 253.94 and 1.4 mg/L 2,4 DCP and the catalyst concentration was set at 1 g/L of TiO2.

These experiments were made at room temperature, without adjustment of pH solution, at a stirring speed of 555 rev/min. and high UV irradiation intensity.

It should be noted that for all experiments the pH measured was between 5.5 and 6. Was observe that it evolution was increased in the moment of the beginning of photocatalyse degradation and during of the irradiation it decreased until approximately 4 to 4.8.

Each experiment has started with adsorption stage, in the dark and then, was started the photocatalytical degradation, for 180 minutes in the present of the maximum light flux.

The following are some of the results obtained in this study.

The graph from figure III.2. present effect of concentration for 253,94 mg/L 2,4-DCP and 1 g/L of TiO2 in aqueous medium. The reaction was conducted at initial pH of the solution, measured at 5.9.

Figure III.2. Effect of 2,4-DCP concentration during photocatalytic degradation for an initial concentration of 253.94 mg/L of 2,4-DCP, 1g / L of TiO2 and maximum light flux (306 W/m2).

The results showed that for first 60 min. dedicated of adsorption stage, the mass of pollutant adsorbed in the surface of TiO2 was insignificant. It can be seen that in the presence of UV lamps, pollutant concentration begins to fall. During the photocatalytic degradation kinetics, the initial concentration decreased to 224,75 mg/L after 180 minutes.

For this concentration, percentage of degradation was of 11.49 %, after 180 min. of UV irradiation and COD was mineralized in proportion of 9,3% (figure III.3.).

Figure III.3. Photocatalytic degradation and TOC mineralization percentage for 253,49 mg/L of 2,4- DCP, 1 g/L TiO2, and maximum light flux (306 W/m2)

In the same conditions of the work, was tested the rate of degradation for 149,65 mg/L. mg/L 2,4-DCP. The results obtained are shown in figure III.4.

Figure III.4. Effect of 2,4-DCP concentration during photocatalytic degradation for an initial concentration of 149.65 mg/L of 2,4-DCP, 1g / L of TiO2 and maximum light flux (306 W/m2)

During the photocatalytic degradation kinetics, the initial concentration of 149,65 decreased to 129.54 mg / L of 2,4-DCP, after 180 min of reaction.

For this concentration, the percentage degradation was of 15.68 % and the percentage of TOC mineralization was of 9.22 % (figure III.5.)

Figure III.5. Photocatalytic degradation and TOC mineralization percentage for 149.65 mg/L of 2,4- DCP, I g/L TiO2, and maximum light flux (306 W/m2).

In the following table are shown the percentages of degradation and mineralization obtained in the study on effect initial concentration and 1 g/L TiO2 at maximal radiant flux, after 240 min. of UV irradiation.

Table III.1. The percentage of degradation and mineralization after 30 min. and 180 min. of irradiation for the experiments made at different concentrations of 2,4-DCP, for 1g/L TiO2, at maximum light flux (306 W/m2 ).

The best result of degradation and mineralization, after 180 min. of heterogeneous photocatalysis was obtained for the concentration of 3.92 mg/L 2,4-DCP.

The results obtained showed that the degradation of 2,4-DCP is strongly influenced by the initial concentration of pollutant and that this diminishes over time.

These data show that the degradation ratio decreased with the increases of 2,4-DCP concentration. The reason is that the intermediates concentration increased with the enhancement of organic initial concentration, which restrained organic from adsorbtion on TiO2 surface and transferring in solution, so, degradation ration decreased.

Analyzing the curve shown in figure III.5. we can observed that the photocatalytic degradation of 2,4-DCP is done in 2 stages:

1 stage: faster degradation, during the first 30 min.of reaction;

2 stage: slower degradation, marked by the by products appearance.

Figure III.6. Evolution of the degradation rate for the different concentrations of 2,4-DCP at

I g/L TiO2, and maximum light flux (306 W/m2).

For each experiment the initial degradation rate was determined by calculating the slope of the curve C2,4-DCP = f (t) for the first 30 minutes. By definition, the initial degradation rate is defined as shown below:

Table III.2. The initial rate of degradation calculated for each concentration of 2,4-DCP at

1 g/L TiO2 and maximal radiant flux (306 W/m2).

III.4.Effect of the catalyst concentration at maximum light flux (306 W/m2)

In order to determine the optimum concentration of TiO2 under conditions used in this studies, the degradation rate of 2,4-DCP was measured at different concentrations of TiO2 in the range of 0.5 – 3 g/L. For these experiments was imposed a concentration of 3 mg/L 2,4-DCP, approximately.

These tests were conducted under the same conditions of pH, temperature and stirring rate the same as those carried out previously.

For each condition tested the degradation kinetics were followed for 180 minutes of irradiation.

The results show that the initial rate of photodegradation increases rapidly at a concentration of TiO2 – P500 up to 0.7 g/L, approximately.

In the graph form figure III.7. are presented the curves of degradation for each concentration of TiO2 tested.

Figure III.7. The kinetics of photocatalytic degradation of 2,4-DCP in aqueous medium at different initial concentrations of TiO2 at a fixed concentration of 3 mg/L 2,4-DCP and luminous flux of 306 W/m2.

Figure III.8. The curves of degradation and mineralization of 2.4-DCP obtained for different initial concentration of TiO2 at maximum light flux (306 W/m2).

The results obtained from this experiments where we varied the concentration of TiO2 and we kept constant the concentration of pollutant, in our case of 2,4-DCP, made at maximum light flux, are presented in the following table .

Table III.3. The percentage of degradation and mineralization after 180 min. of irradiation for the experiments made at different concentrations of TiO2.

Was observed that for:

A low catalyst concentration the degradation rate is increased, that is expected from the increase in active sites available for the phenomenon of adsorption, and also to the increase of hydroxyl radical generation on the catalyst surface.

At higher amount of catalyst, exist enough active sites present to adsorb the 2,4-DCP molecules and the additional amount of TiO2 tends to stabilize.

For each experiment the initial speed of degradation was determined by calculating the slope of the curve C2,4-DCP = f (t) for the first 30 minutes.

The results obtained are presented in the following table:

Table 4.The initial rate of degradation of 2,4-DCP calculated for each concentration of TiO2, for 3 mg/L of 2,4-DCP and maximum light flux (306 W/m2).

III.5. Effect of radiant flux

In this several study tests were conducted to assess the influence of photon flux on the kinetics of degradation of the compound. In this report we have chosen to study the efficiency of the degradation kinetics performed for an initial 2,4-DCP concentration of 3 mg/L and 1 g/L of TiO2. Was tested 3 luminous flux as follows: 306 W/m2, 135.72 W/m2 and 49.38 W/m2.

Figure III.9. Evolution of 2,4-DCP concentration during the photocatalytic degradation for initial concentration of 3 mg /L 2,4-DCP, 1 g/L TiO2 and for three incident light flows.

Figure III.7. shows the results obtained for 180 minutes of irradiation. It can be seen that increasing of the luminous flux promotes overall degradation of the pollutant. This is explained by an increase in oxidative entities formed.

For the test conducted for a maximum light flux the degradation rate of about 82% was observed after 180 minutes of irradiation.

Table III.5. The percentage of degradation and mineralization after 180 min. of irradiation, for the experiments realized at different incident light flux and constant concentration of TiO2 and 2,4-DCP.

From the results, shown in the table III.5. we can observe that the degradation rate of 2,4-DCP is more efficient when it is used a intensity light at maximum power. If the intensity of UV lamps increased then the percentage of degradation is increased.

We can see that in this case the percentage of mineralization is directly proportional with the intensity of UV lamps.

The same observation we can make for the initial speed of degradation, shown in table III.6.

Table III.6. The initial speed of degradation of the 2,4-DCP calculated for each radiant light tested for an intial concentration of 3 mg/L of 2,4-DCP and 1g/L TiO2.

III.6. Effect of pH on 2,4-DCP degradation

It is known that pH can affect the mechanism and routes of degradation. The TiO2 point of zero charge (pzc) is between pH 5.6 and 6.4. Hence, depending on the pH the catalyst surface will be charged positively (for Ph < pzc), negatively (for pH > pzc) or neutrally (for pH ~ pzc). This characteristic affects significantly on the adsorption and desorption properties of TiO2. Also, the structure of pollutant will change with the pH [23].

This aspect was studied in our system with different initial pHs (2, 4, 8, 11) which was adjusted by using NaOH and H2SO4 1M solution.

We decided to see if the change is better to work at a basic pH or acidic pH.

For these experiments 2,4-DCP concentration of 3 mg/L approximately, was imposed. These tests were conducted under the same conditions of pH, temperature and stirring rate the same as those carried out previously.

For each condition tested the degradation kinetics were followed for 180 minutes of irradiation.

Results obtained from experiments with varying pH from 2,25 to 10,5 are illustrated in figure III.10.

Figure III.10. Final TOC and 2,4-DCP degradation after 180 min. of illumination for different pHs in non-buffered conditions. 2,4-DCP = 3 mg/L, TiO2= 1 g/L at maximal radiant flux.

It was found that pH has a pronounced influence on the mechanism pathway and final degradation of 2,4-DCP. In all the case, the pH decrease along the time, showing that more acidic species were appearing.

Analysing the 2,4-DCP curve, it can be observed how for pHs greater than 5.9, the final 2,4-DCP removal decreased. For lower pH, the final degradation was practically the same.

These results would suggest that the affinity between 2,4-DCP and TiO2 improves in acidic conditions.

On the other hand, for alkaline conditions, 2,4-DCP and TiO2 are mostly charged negatively so can exist a repulsion between both compounds.

In the following table are presented the results obtained for each value of pH tasted.

Table III.7. The percentage of degradation and mineralization after 180 min. of irradiation, for the experiments realized at different pH values and constant concentration of TiO2 and 2,4-DCP, maximum light flux.

Table III. 8. The initial speed of degradation of the 2,4-DCP calculated for the experiments where the initial pH of solution was adjusted at an intial concentration of 3 mg/L of 2,4-DCP, 1g/L TiO2 and maximum light flux.

IV. Conclusions

The main objective of this work was to study the photocatalytic degradation of a herbicide compound in an aqueous medium, in the presence of a semiconductor, titanium dioxide PC 500. 2,4-Dichlorophenol is an organic compound known as refractory to the conventional methods that use routinely in the treatment of industrial effluents or in the case of wastewater.

The influence of various process parameters on the degradation kinetics of this compound was evaluated. The main parameters studied were: the effect of concentration of 2,4-DCP, the catalyst concentration, the effect of UV radiation, the effect of pH.

The analysis of experimental results obtained in this work reveals the following conclusions:

UV radiation alone slightly degrades DCP in aqueous solution (less than 10 %).

The degradation of 2,4-DCP is strongly influenced by the initial concentration of pollutant and that this diminishes over time.

The effect of adsorption also has been analyzed and it has been concluded that the substances which are readily adsorbed are degraded at a faster rate.

The degradation rate of 2,4-DCP is more efficient when it is used a intensity light at maximum power. If the intensity of UV lamps increased then the percentage of degradation is increased.

The percentage of mineralization is also, directly proportional with the intensity of UV lamps.

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