Fenton treatment

Accepted Manuscript II Removal of phosphonates from industrial wastewater with UV/Fe , Fenton and UV/ Fenton treatment Eduard Rott, Ralf Minke, Ulusoy Bali, Heidrun Steinmetz PII:

S0043-1354(17)30483-9

DOI:

10.1016/j.watres.2017.06.009

Reference:

WR 12963

To appear in:

Water Research

Received Date: 8 February 2017 Revised Date:

2 June 2017

Accepted Date: 4 June 2017

Please cite this article as: Rott, E., Minke, R., Bali, U., Steinmetz, H., Removal of phosphonates from II industrial wastewater with UV/Fe , Fenton and UV/Fenton treatment, Water Research (2017), doi: 10.1016/j.watres.2017.06.009. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

ACCEPTED MANUSCRIPT 1

Removal of phosphonates from industrial wastewater with

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UV/FeII, Fenton and UV/Fenton treatment Eduard Rotta,*, Ralf Minkea, Ulusoy Balib, Heidrun Steinmetzc

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a

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of Stuttgart, Bandtäle 2, 70569 Stuttgart, Germany

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b

Environmental Engineering Department, Cumhuriyet University, 58140 Sivas, Turkey

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c

Chair of Resource Efficient Wastewater Technology, University of Kaiserslautern, Paul-

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Ehrlich-Str. 14, 67663 Kaiserslautern, Germany

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* Corresponding author. Tel.: +49 711 68560497; fax: +49 711 68563729; E-mail address:

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[email protected] (E. Rott)

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Graphical abstract

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Industry:

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Institute for Sanitary Engineering, Water Quality and Solid Waste Management, University

Catalyst: H2SO4/ FeSO4 NaOH UV lamp

UV-FSR: UV free surface reactor ST: Sedimentation tank NT: Neutralization tank

Catalyzed photolysis: H2SO4/NaOH

(Photo-)Fenton:

Phosphonate reduced wastewater ST

NT

UV-FSR

Fenton reagent: FeSO4/H2O2

Receiving water

Sludge Ca(OH)2/NaOH

UV lamp

H2SO4

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Paper/ textile production Rinsing processes in industry Other industries

Concentrate

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Cooling system

Organically polluted wastewater

ST

ST NT

UV-FSR Sludge (acidic)

Phosphonate reduced wastewater

Wastewater treatment plant

Sludge (neutral)

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Abstract

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Phosphonates are an important group of phosphorus-containing compounds due to their

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increasing industrial use and possible eutrophication potential. This study involves

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investigations into the methods UV/FeII, Fenton and UV/Fenton for their removal from a pure 1

ACCEPTED MANUSCRIPT water matrix and industrial wastewaters. It could be shown that the degradability of

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phosphonates by UV/FeII (6 kWh/m³) in pure water crucially depended on the pH and was

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higher the less phosphonate groups a phosphonate contains. The UV/FeII method is

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recommended in particular for the treatment of concentrates with nitrogen-free phosphonates,

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only little turbidity and a low content of organic compounds. Using Fenton reagent, the

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degradation of polyphosphonates was relatively weak in a pure water matrix (<20%

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transformation to o-PO43–). By means of the Photo-Fenton method (6 kWh/m³), those

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phosphonates with the smallest numbers of phosphonate groups were easier degraded as well

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at pH 3.5 in a pure water matrix (o-PO43– formation rates of up to 80%). Despite an

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incomplete transformation of organically bound phosphorus to o-PO43– with Fenton reagent in

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an organically highly polluted wastewater (max. 15%), an almost total removal of the total P

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occurred. The most efficient total P elimination rates were achieved in accordance with the

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following Fenton implementation: reaction → sludge separation (acidic) → neutralization of

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the supernatant → sludge separation (neutral). Accordingly, a neutralization directly after the

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reaction phase led to a lower total P removal extent.

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Keywords: Phosphonates, metal-catalyzed photolysis, Photo-Fenton, wastewater treatment

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1 Introduction

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1.1 Motivation

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Phosphonates are versatile complexing agents due to their excellent complex stability and

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their substoichiometric effectiveness as ‟thresholders” (Gledhill and Feijtel, 1992). They are

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used in the paper and textile industries, are part of household and industrial detergents as well

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as cosmetics, are used as antiscalants in membrane technology, as cooling stabilizers in

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cooling systems and are also found in oil production, cement, electroplating, and medical

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applications. In 1998, global consumption of phosphonates was 56,000 tons (Davenport et al.,

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ACCEPTED MANUSCRIPT 2000). A total consumption of 94,000 tons in 2012 shows a significant increase in the

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consumption of these compounds (EPA, 2013).

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Fig. 1 gives an overview of the chemical structures of the quantitatively most important

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phosphonates PBTC, HEDP, NTMP, EDTMP and DTPMP. The covalent C-P bond in the

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phosphonate groups (C-PO(OH)2) of the phosphonates ensures high molecular stability

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(Gledhill and Feijtel, 1992). Phosphonates can be divided into the group of nitrogen-free

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phosphonates, also characterized by carboxyl and hydroxyl groups, and into the group of

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aminophosphonates. Phosphonates with more than one phosphonate group are also referred to

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as polyphosphonates. Phosphonates have a high water-solubility, a low solubility in organic

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solvents and a very low volatility (Nowack, 2003; Gledhill and Feijtel, 1992).

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O

HO

O

O

H3C OH

O

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P O OH OH

PBTC 2-Phosphonobutane-1,2,4tricarboxylic acid

O HO

P OH

OH OH P OH O

HEDP 1-Hydroxyethylidene(1,1-diphosphonic acid)

HO P HO

N HO

OH

HO P HO

O OH

HO

O

OH OH

N

HO

P

NTMP Nitrilotrimethylphosphonic acid

P

O

N

P

OH P OH O

O

EDTMP Ethylenediaminetetra(methylene phosphonic acid)

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HO

P

O

O

OH

P

O HO P HO

N

OH HO OH HO

O P

O N

N P O

P OH OH

OH OH

DTPMP Diethylenetriaminepenta(methylene phosphonic acid)

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Fig. 1:

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Phosphonates are subject to natural elimination mechanisms (Nowack and Stone, 2000;

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Matthijs et al., 1989; Schowanek and Verstraete, 1990), which speak for long-term release of

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bioavailable ortho-phosphate (o-PO43–) in water, thus a contribution of phosphonates to the

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eutrophication of water bodies cannot be excluded (Studnik et al., 2015; Grohmann and

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Horstmann, 1989). Direct discharge of phosphonate-containing cooling water or membrane

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concentrate, as well as insufficiently purified municipal wastewater, can also lead to increased

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phosphorus concentrations in water bodies. Furthermore, phosphonates are associated with

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the remobilization of toxic heavy metals bound in sediments (Gledhill and Feijtel, 1992;

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Bordas and Bourg, 1998; Nowack, 2003). Industrial effluents are partly directed to central

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Chemical structures of phosphonates (based on ACS, 2016).

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ACCEPTED MANUSCRIPT wastewater treatment plants, either unpurified or pre-cleaned (indirect discharge). There,

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phosphonates can interfere with phosphate precipitation by masking the precipitant

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(Horstmann and Grohmann, 1984). In biological wastewater treatment plants, phosphonates

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are neither degraded aerobically (Horstmann and Grohmann, 1988) nor anaerobically

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(Nowack, 1998). However, phosphonates are removed to degrees of 50 to at least 95%

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(Metzner, 1990; Müller et al., 1984; Nowack, 1998 and 2002). Their elimination in

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wastewater treatment plants is therefore primarily ascribed to adsorption onto activated sludge

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and metal precipitates (Nowack, 2002). Accordingly, these compounds are brought into the

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environment when the contaminated sewage sludge is applied in agriculture. But still low

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concentrations of phosphonates may leave wastewater treatment plants via the water path

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through the effluent. Removal of phosphonates from industrial wastewater before its direct or

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indirect discharge is therefore particularly desirable.

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Phosphonate-contaminated industrial wastewaters can have very different compositions: on

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the one hand, usually clear, organically only slightly polluted concentrates with typically high

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water hardness and anion concentrations (e.g. cooling water, membrane concentrate), and

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organically polluted wastewaters, e.g. wastewaters from industrial rinsing processes and

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wastewaters from the textile and paper industries.

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Studies about the elimination of phosphonates (elimination by flocculation, filtration, O3,

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MnII/O2, UV/FeII) (Fischer, 1992; Klinger et al., 1998; Nowack and Stone, 2000; Boels et al.,

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2010; Zhou et al., 2014; Lesueur et al., 2005) generally comprise experiments with pure water

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spiked with phosphonates or a very small phosphonate spectrum, so that the influence of the

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wastewater matrix effects on the elimination processes is largely unknown. So far, there is

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still no established procedure for the selective removal of phosphonates from wastewater.

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Phosphonates are oxidized to a greater extent, especially in the presence of metal ions

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(MnII/O2, UV/FeII, FeII/H2O2) (Nowack and Stone, 2000; Matthijs et al., 1989; Pirkanniemi et

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ACCEPTED MANUSCRIPT al., 2003). This study therefore investigates the removal of phosphonates from industrial

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wastewaters by the promising UV/FeII and (UV/)FeII/H2O2 methods exploiting these

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properties.

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1.2 Metal-catalyzed photolysis (UV/FeII)

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Matthijs et al. (1989), Gledhill and Feijtel (1992) and Fischer (1993) have shown that

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phosphonates are degraded to o-PO43– in the presence of metals and UV radiation or sunlight.

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Of all the metals studied (FeIII, CrIII, ZnII, CuII, CaII), iron proved to be the most efficient

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catalyst. In the experiments of Lesueur et al. (2005), after 1.5 h at least 80% of the organically

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bound phosphorus of the aminophosphonates NTMP, EDTMP, DTPMP and HDTMP (1

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mg/L) dissolved in pure water in a UV immersion lamp reactor (150 W, medium pressure) at

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pH 3 and 5–6 in the presence of 0.2 mg/L FeII was converted to o-PO43– (due to the complex

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analysis of phosphonates, only the total P or their oxidation to o-PO43– are determined in most

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studies, not the concentrations of the phosphonates themselves). Without iron, the reaction

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proceeded more slowly, but resulted in a pronounced o-PO43– formation of between 70 and

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90% at the same pH values. The reaction proceeded significantly slower at pH 10, so that the

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o-PO43– formation extent did not exceed the 75% mark even after 1.5 h for all phosphonates.

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The experiments showed that aminophosphonates can be degraded even without a catalyst at a

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sufficiently high UV power, the use of a catalyst, however, enhances the degradation of

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phosphonates significantly and may also decrease the required UV energy consumption.

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Matthijs et al. (1989) suggested that the degradation mechanism is a photoinduced ligand-to-

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metal charge transfer, where FeIII-complexes are subject to photolysis. Lesueur et al. (2005),

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however, did not rule out that FeII-phosphonate complexes can also be photolytically

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degraded.

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ACCEPTED MANUSCRIPT So far, there are no current studies about the removal of phosphonates with the UV/FeII

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method. All the studies mentioned have in common that the phosphonates have only been

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spiked in pure water, so that no conclusions can be drawn about the technical feasibility of the

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UV/FeII method for selective phosphonate removal from wastewater.

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1.3 Fenton (FeII/H2O2) and Photo-Fenton (UV/FeII/H2O2)

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Phosphonates have a high stability against decomposition by H2O2, whereby they are able to

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function as bleach stabilizers under strongly alkaline conditions and at very high

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temperatures, such as those present in bleaching liquors of the paper and textile industries. In

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contact with H2O2, aminophosphonates are readily oxidized to N-oxides which retain their

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ability of complexation and peroxide stabilization (Croft et al., 1992; Carter et al., 1967). One

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way to oxidize H2O2 stable compounds is to convert the oxidizing agent into a much more

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reactive form by means of catalysts as in the Fenton reaction. The complex reaction

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mechanism of the Fenton reaction was described extensively by Ya Sychev and Isak (1995)

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and can be summarized as follows:

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Fe2+ + H2O2 → Fe3+ + OH– + •OH Fe3+ + H2O2 → Fe2+ + •O2H + H+

(1) (2)

In the Fenton reaction, ferrous iron is oxidized to ferric iron by H2O2 to produce a highly

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reactive hydroxyl radical (•OH) (Equation 1). Ferric iron can be reduced to ferrous iron again

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in the Fenton-like reaction releasing a hydroperoxyl radical (•O2H) (Equation 2). The

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effectiveness of Fenton reagent can be enhanced by UV radiation. This is essentially due to

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the reaction mechanisms in Equations 3 and 4 (Faust and Hoigné, 1990; Legrini et al., 1993).

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ACCEPTED MANUSCRIPT Fe(OH)2+ + hv → Fe2+ + •OH

(3)

H2O2 + hv → 2 •OH

(4)

UV light stimulates the reduction of ferric iron resulting in the release of further hydroxyl

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radicals to form ferrous iron. Thus, the accumulation of ferric iron and the premature

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termination of the Fenton reaction are counteracted. Furthermore, hydrogen peroxide is

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homolytically cleaved into two hydroxyl radicals under exposure to UV light.

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The higher the reaction pH is, the more unreactive iron hydroxides are formed, which is why

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the Fenton reagent is generally dosed at low pH values. The precipitated sludge acts as an

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adsorbent for reaction products as well as for stable compounds.

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Croft

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aminophosphonates in detail. They found that the peroxide stabilizing effect is not due to the

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retardation of the Fenton reaction (Equation 1). On the contrary, they found that the

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complexation of FeII by aminophosphonates contributed to an increase in the rate constant of

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the Fenton reaction (75 L/mol/s without and 1·104–2·105 L/mol/s in the presence of

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aminophosphonates). The peroxide stabilizing effect is much more due to the fact that

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aminophosphonates stabilize the higher valence state of iron, hinder the effective reduction of

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FeIII to FeII (Equation 2) and thus interfere with the Fenton cycle. Experiments on the removal

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of phosphonates spiked in pure water by means of Fenton reagent were only carried out by

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Grohmann and Horstmann (1989) and by Fürhacker et al. (2005). Grohmann and Horstmann

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(1989) described an experiment in which only 1% of a Fe3+-PBTC complex (100 µmol/L, 27

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mg/L phosphonate) was converted to o-PO43– by equimolar FeII and H2O2. However, the pH

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value underlying the experiment was not mentioned, so that these test results have only

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limited validity. Fürhacker et al. (2005) showed that 1 mg/L HDTMP at pH 3 and 5.8 were

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converted much more effectively to o-PO43– by means of Photo-Fenton reagent

al.

(1992)

investigated

the

mechanism

of

peroxide

stabilization

by

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ACCEPTED MANUSCRIPT (UV/FeII/H2O2) than by means of UV/FeII (same concentrations, only without H2O2). Thus

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far, there have neither been conclusive studies about the removal of phosphonates with Fenton

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reagent with pure water, nor with industrial wastewater, which means that the Fenton and the

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Photo-Fenton method for phosphonate removal still requires considerable research.

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1.4 H2O2 dosage

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The stoichiometrically required H2O2 concentration (cH₂O₂, 100%) can be calculated from the

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COD concentration (chemical oxygen demand for the oxidation of organic constituents in

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water) of the raw sample. The COD is determined by adding the much more reactive

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potassium dichromate to the sample, the oxidation takes place under optimal conditions (high

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temperature, acidic pH, catalyst, etc.) and the amount of potassium dichromate consumed is

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converted into oxygen equivalents. Such a conversion can also be carried out for H2O2. For

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the reduction of 1 mol H2O2 (34 g/mol), 2 mol electrons are required (Equation 5). This is half

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the amount of the substance which is required during the reduction of 1 mol O2 (32 g/mol)

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(Equation 6). Taking this ratio into consideration, a factor of 2.125 results, which must be

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multiplied by the COD of the raw sample in order to determine the stoichiometrically required

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H2O2 concentration (Equation 7). Stoichiometrically under-dosed and over-dosed H2O2

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concentrations can be set in the percentage ratio (VH₂O₂) to the stoichiometrically required

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concentration. By means of Equation 7, this ratio can also be represented as a function of the

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COD of the raw sample (Equation 8).

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H2O2 + H+ + 2 e– → H2O + OH–

(5)

O2 + 4 H+ + 4 e– → 2 H2O

(6)

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ACCEPTED MANUSCRIPT g 4e– 34 mol H2 O2 mg mg cH₂O₂, 100% ቂ ቃ = – · ·COD ቂ ቃ =2.125·COD ቂ ቃ g L 2e L L 32 O mol 2

(7)

mg ቃ mg L -1 -1 mg VH₂O₂ ሾ%ሿ= ·100%=c ቂ ቃ ·100%·2.125 ·COD ቂ ቃ H₂O₂ mg L L cH₂O₂, 100% ቂ ቃ L

(8)

mg

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cH₂O₂ ቂ

If we assume that mineralization of organic matter converts C, H, P, and N into inorganic

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carbonate, water, ortho-phosphate, and nitrate, then the COD for the total oxidation of 3 mg/L

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phosphonate (experiments of this work with a pure water matrix were always carried out with

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this concentration) can be calculated as follows: PBTC: 2.310 mg/L, HEDP: 1.398 mg/L,

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NTMP: 2.088 mg/L, EDTMP: 2.751 mg/L, DTPMP: 3.098 mg/L (Supplementary data,

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Section A).

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2 Materials and methods

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2.1 Examined wastewaters

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2.1.1

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For experiments with wastewaters, cooling tower effluent (concentrate) and wastewater from

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phosphonate production (organically polluted wastewater) were obtained. All samples were

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taken randomly during operation. After sampling, all samples were stored at approx. 4°C

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without further processing and were generally used within one week.

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High alkaline earth metal and COD concentrations can result in severe problems regarding the

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detection of phosphonates in wastewater (Nowack, 1997; Klinger et al., 1997; Knepper,

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2003). Only recently, Schmidt et al. (2014) described a method using the coupled system IC-

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ICP-MS, which combines the advantages of low determination limits for polyphosphonates in

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the range of 0.1 µg/L with the applicability to environmental samples. Whether this method

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ACCEPTED MANUSCRIPT can also be applied to heavily polluted industrial wastewater and concentrates with high water

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hardness still needs further investigation, which is why no phosphonate detection was applied

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in this work. However, In all wastewater samples the total P of the raw samples (TP0) was

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predominantly composed of the dissolved organic phosphorus fraction to which phosphonates

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are attributed. Therefore, the determination of TP and o-PO43–-P was considered sufficient for

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all experiments.

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2.1.2

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The wastewater sample from a coal-fired power plant was taken directly from the cooling

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tower basin. The cooling tower is fed with flocculated river water. According to the operator,

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the hardness stabilizer PBTC is used. The taken sample had only concentrations of 0.15 mg/L

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TP0 and 0.03 mg/L o-PO43–-P due to prolonged rainfall prior to sampling. Gartiser and Urich

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(2002), however, assume typical phosphonate use concentrations in cooling systems of 1.5 to

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20 mg/L. Accordingly, and to allow better comparability of the results with pure water

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experiments in which phosphonate concentrations of 3 mg/L were used, the sample was

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spiked with 3 mg/L PBTC (0.35 mg/L PBTC-P) resulting in a TP0 concentration of 0.5 mg/L.

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The cooling tower effluent (pH 7.4) was very clear (3 NTU) and colorless with a very low

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COD of 20 mg/L, a general water hardness of 35–40 dGH (160–180 mg/L Ca, 50–60 mg/L

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Mg), an electrical conductivity of 2.1 mS/cm and an acid capacity to pH 4.3 of 1.8 mmol/L.

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Further parameters were 380 mg/L chloride, 390 mg/L sulfate, 0.14–0.16 mg/L Fe, 1.3–1.5

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mg/L Al and <45 µg/L Mn.

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2.1.3

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The phosphonate producer synthesizes technical solutions and solids of polyphosphonates,

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primarily HEDP, NTMP, EDTMP, and DTPMP. Wastewater is mainly produced when the

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reactors are rinsed. The wastewater sample taken had a very high TP0 concentration of

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350 mg/L with an o-PO43–-P concentration of only approx. 10 mg/L. It must therefore be

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Cooling tower effluent of a power plant

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Wastewater from phosphonate production

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ACCEPTED MANUSCRIPT assumed that phosphonates formed the largest proportion of the phosphorus compounds

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present. The pH of the sample was 9 and the COD concentration was 4.8 g/L. Furthermore,

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the following composition was typical for this kind of wastewater (minimum and maximum

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values in three randomly taken samples): 14–28 mS/cm electrical conductivity, 110–750 NTU

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turbidity, 5.4–18.7 mmol/L acid capacity to pH 4.3, 30–42 mg/L Ca, 2.8–3.4 mg/L Mg, 1.9–

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2.8 mg/L Fe, 1.7–2.3 mg/L Al, 47–61 µg/L Mn, 2.7–7.9 g/L chloride and 1.9–4.0 g/L sulfate.

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2.2 Reagents and chemicals

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For all stock solutions and dilutions, pure water was used produced on-site by means of an ion

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exchanger (Seradest SD 2000) and a downstream filter unit (Seralpur PRO 90 CN).

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FeSO4·7H2O (analytical grade) was purchased from Sigma-Aldrich (St. Louis, Missouri,

223

U.S.A.), H2O2 solution (30%, pure) from AppliChem (Darmstadt, Germany), H2SO4 solution

224

(95–97%) from Merck (Darmstadt, Germany) and solid NaOH from VWR International

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(Radnor, Pennsylvania, U.S.A.). PBTC was obtained from Zschimmer & Schwarz Mohsdorf

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(Burgstädt, Germany) as a technical solution (50%, CUBLEN P 50). HEDP·H2O (≥95%) and

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NTMP (≥97%) were purchased as solids from Sigma-Aldrich. EDTMP (approx. 5.3% water

228

of crystallization) and DTPMP (approx. 16% water of crystallization) were synthesized by

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Zschimmer & Schwarz Mohsdorf. All phosphonate samples had no significant phosphate

230

impurity (ratio of phosphate-P to total P: <1%, from own measurements).

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2.3 Experimental procedure

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2.3.1

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In Experiments 1 and 2, the degradability of the phosphonates PBTC, HEDP, NTMP,

234

EDTMP and DTPMP dosed to pure water in a concentration of 3 mg/L (0.35 mg/L PBTC-P,

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0.90 mg/L HEDP-P, 0.93 mg/L NTMP-P, 0.85 mg/L EDTMP-P, 0.81 mg/L DTPMP-P) was

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investigated by means of Fenton, UV/Fenton, UV/FeII, UV/H2O2 and UV only by varying the

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Overview of the experiments

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ACCEPTED MANUSCRIPT FeSO4 and H2O2 concentrations as well as the pH. In Experiments 3 and 4, the UV/FeII and

238

the (Photo-)Fenton method were investigated with phosphonate-containing industrial

239

wastewater regarding the elimination of phosphorus and the required dosage concentrations

240

(variation of FeII-H2O2-ratio and H2O2 concentration). Depending on the experiment, several

241

samples were prepared in duplicate or triplicate and analyzed with single determinations. The

242

results of duplicate or triplicate samples were averaged and the standard deviation was

243

calculated (graphical representation with error bars).

244

Phosphonates were spiked using 3 g/L phosphonate stock solutions prepared in advance. All

245

test series were carried out in 100 mL bottles on magnetic stirrers (250 rpm) filled with either

246

50 mL (samples with pure water) or 100 mL sample (samples with wastewater since due to

247

the higher amount of analyzed parameters more volume was required). The stoichiometric

248

H2O2 concentration was calculated using the chemical oxygen demand (COD) of the raw

249

sample (see Section 1.4). In the case of low dosage concentrations (<0.2–0.3 g/L Fe),

250

FeSO4·7H2O was dosed from a solution prepared with pure water; otherwise, FeSO4·7H2O

251

was directly dosed into the sample bottles. The irradiation of samples with UV light was

252

carried out as described in Section 2.3.2. Subsequent to a one-hour reaction phase, a partial

253

volume was taken from all samples in the still homogeneous state, which was analyzed for the

254

o-PO43– concentration. In order to determine the sum of dissolved, precipitated and adsorbed

255

o-PO43–, the analysis was conducted without membrane filtration (WMF) (marked as ‟o-PO43–

256

-PWMF”, for further descriptions see Section 2.4). After neutralization, the supernatants were

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analyzed for the TP, o-PO43–-P or COD. For detailed descriptions of the experiments, see

258

Sections 2.3.3 to 2.3.6.

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ACCEPTED MANUSCRIPT 2.3.2

UV lamp

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Four sample bottles, each equipped with magnetic rods, were randomly placed in a square

261

arrangement under the Sterisol NN 30/89 UV lamp (30 W total power, approx. 300 W/m²,

262

low-pressure, 11 W UV-C power at 254 nm) on four magnetic stirrers. The distance between

263

the sample surface and the UV lamp was generally about 10 cm. The irradiated sample

264

surface was about 10 cm2. For a sample volume of 100 mL, the specific energy consumption

265

per sample was therefore about 3 kWh/m³ (=300 W/m²·10 cm²·1 h/100 mL). Of this, the

266

active UV-C range accounted for 1.1 kWh/m³ (=3 kWh/m³·11 W/30 W). With a sample

267

volume of 50 mL, the specific energy consumption corresponded to 6 kWh/m³, of which 2.2

268

kWh/m³ accounted for the UV-C range.

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2.3.3

270

1 liter of pure water was spiked with a phosphonate concentration of 3 mg/L, brought to pH

271

3.5 by means of H2SO4, and then intensively stirred and analyzed for the TP0 concentration.

272

FeSO4 was added at various concentrations to 50 mL samples of this solution (molar FeII-

273

H2O2-ratios: 1:30–5:1 and one ratio in the range between 1:100 and 1:400 depending on the

274

phosphonate to take into account an equimolar FeII-phosphonate-ratio as well; the highest

275

molar FeII-phosphonate-ratio tested was 2,000:1). Then, H2O2 was added to the samples in a

276

stoichiometrically eightfold excess (at this H2O2 concentration a maximum of o-PO43–

277

formation could be observed in preliminary experiments; Supplementary data, Section B)

278

resulting in a maximum pH decrease down to pH 2.75 depending on the FeII concentration.

279

After stirring for 1 h in the open bottle (an additional pH drop of no more than 0.2 was

280

observed here; the pH range of this experiment was therefore 2.5–3.5), the o-PO43–-PWMF

281

concentration was determined in each sample. Each sample was prepared in duplicate.

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ACCEPTED MANUSCRIPT UV/FeII, UV/H2O2, UV/FeII/H2O2 in a pure water matrix (Experiment 2)

2.3.4

283

1 liter of pure water was provided with a phosphonate concentration of 3 mg/L and FeSO4

284

several times in an equimolar ratio to the phosphonate. Batches without FeSO4

285

(c(FeII):c(phosphonate)=0) were also prepared. The initial pH (3.5, 5.0, 6.5, 8.0, 9.5) was

286

adjusted by means of H2SO4 or NaOH, the solution was then stirred intensively and analyzed

287

for the TP0 concentration. Optionally, H2O2 was added in the stoichiometrically sixteen-fold

288

concentration to 50 mL samples of this solution (at this H2O2 concentration a maximum of

289

o-PO43– formation could be observed in preliminary experiments; Supplementary data,

290

Section B). After stirring for 1 h under the UV lamp, the o-PO43–-PWMF concentration was

291

determined in each sample. Due to the experimental setup (the UV lamp was mounted directly

292

above the reaction vessel), a pH correction was not possible during this stirring phase.

293

However, the pH after 1 h was measured and depending on the tested variant (UV, UV/FeII,

294

UV/H2O2, UV/FeII/H2O2), different, mostly only small pH changes were observed. Thus, the

295

following pH ranges were tested in this experiment: 3.4–3.6, 4.6–5.0, 6.2–6.9, 7.0–8.0, 8.0–

296

9.5. Each sample was prepared in triplicate.

297

2.3.5

298

The cooling tower effluent spiked with 3 mg/L PBTC was brought to pH 3.5 using H2SO4.

299

Optionally, 4.6 mg/L FeII and optionally 30 mg/L H2O2 were added to 100 mL samples of this

300

sample in order to test the following four variants: UV, UV/FeII, UV/H2O2, UV/FeII/H2O2

301

(because 100 mL samples were used, the specific UV energy consumption was half as high

302

[3 kWh/m³] as in the experiments with pure water [6 kWh/m³]). Each of these variants was

303

prepared fourfold (four bottles with the same contents). After stirring for 1 h under the UV

304

lamp, the o-PO43–-PWMF concentration (homogeneous) was determined for each variant in all

305

four samples. The addition of chemicals and the contact time had not resulted in any

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ACCEPTED MANUSCRIPT significant pH change (maximum drop down to pH 3.4) due to the strong buffer capacity of

307

the wastewater. Two of the initial four equally prepared samples of each variant were

308

immediately membrane-filtered (0.45 µm pore size, nylon filter) and the o-PO43– as well as

309

the TP concentrations of the filtrate (pH 3.5) were determined. The two other samples were

310

first neutralized by means of NaOH (pH 7.0), then membrane-filtered and now the o-PO43– as

311

well as the TP concentrations of the filtrate were determined.

312

2.3.6

313

The raw sample of phosphonate production wastewater was brought to pH 2.5 by means of

314

H2SO4. After intensive mixing, it was analyzed for the TP0, o-PO43– and COD concentrations.

315

FeSO4 and H2O2 were added to several 100 mL samples of this sample resulting in a pH

316

decreased to 2.2–2.5. Each sample was prepared in replicate. Each 100 mL sample was then

317

stirred in the open bottle for 1 h. Here, an insignificant additional pH drop of no more than

318

0.15 was observed. Immediately thereafter, the o-PO43–-PWMF concentration (homogeneous)

319

was determined in each sample. The further procedure involved two variants. One of the

320

replicated samples was treated according to the neutralization variant 1. Here, the sample was

321

neutralized directly with NaOH and then sedimented for 15–24 h. The resulting supernatant

322

(‟supernatant N1 (neutral)”) was then analyzed for parameters such as o-PO43–-PWMF and TP.

323

The other sample was treated according to the neutralization variant 2. Here, the sample was

324

first sedimented for 15–24 h without prior pH adjustment. 60 mL of the resulting supernatant

325

(‟supernatant N2a (acidic)”) were transferred into an empty bottle. At that point, the o-PO43–

326

-PWMF and TP concentrations of this supernatant were determined. The transferred sample was

327

first neutralized by means of NaOH, and then sedimented for 15–24 h. The resulting

328

supernatant (‟supernatant N2b (neutral)”) was then analyzed for the o-PO43–-PWMF and TP

329

concentrations.

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ACCEPTED MANUSCRIPT 2.4 Analytical methods

331

All glass materials which came into contact with the sample were rinsed with hydrochloric

332

acid and pure water in advance. The total P (TP) determination was carried out according to

333

ISO 6878 (molybdenum blue method) by means of a one-hour peroxodisulfate digestion. ISO

334

6878 provides for the separation of the suspended sample components by membrane filtration

335

in advance to the o-PO43– analysis. In order to determine the extent of organically bound

336

phosphorus oxidized to o-PO43– during the UV/FeII and Fenton reactions, such membrane

337

filtration was not carried out (marked as ‟o-PO43–-PWMF”) (see also Supplementary data,

338

Section C). Precipitated phosphate or adsorbed o-PO43–, which are separated from the sample

339

by a filtration, were thus measured together with dissolved o-PO43–. This is important, since

340

during the reaction sludge (predominantly Fe(OH)3) precipitates, so that o-PO43– formed in

341

the reaction phase either adsorbs or reacts to non-soluble FePO4. The ascorbic acid solution

342

(reducing agent) required according to ISO 6878 was always added directly to the partial

343

volumes for the o-PO43–-PWMF determination in order to interrupt the oxidation processes.

344

Turbid and colored samples were additionally treated with a compensation solution according

345

to ISO 6878. The extinctions were measured with the UV/VIS spectrophotometer JASCO V-

346

550.

347

The pH was determined using the WTW pH electrode SenTix 81 in combination with the

348

WTW pH91 instrument. The chemical oxygen demand (COD) was determined using the

349

Hach Lange cuvette rapid tests LCK 414 and LCK 514. In these tests, a certain quantity of

350

sample is added to a predefined mixture of sulfuric acid, potassium dichromate, silver sulfate

351

and mercury salt, heated up to 148°C for two hours in a thermostat (Hach Lange HT200S)

352

and then analyzed in a photometer (Hach Lange DR2800). H2O2-containing samples were

353

pre-treated with Aspergillus niger catalase (Sigma-Aldrich) (max. 20 mg/L) before the

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ACCEPTED MANUSCRIPT phosphorus and COD analysis as hydrogen peroxide in the sample was found to be interfering

355

with the analysis (Talinli and Anderson, 1992).

356

3 Results and discussion

357

3.1 FeII/H2O2 in a pure water matrix (Experiment 1)

358

In order to clarify the question of to what extent phosphonate-phosphorus is oxidized to

359

o-PO43–-P by exposure to Fenton reagent for one hour without influence of UV radiation,

360

H2O2 was added to each phosphonate at pH 2.5–3.5 in pure water in stoichiometrically

361

eightfold excess, while the FeII-H2O2-ratio was subject to a variation.

362

No significant reaction to o-PO43– was observed for all phosphonates when no FeII was dosed

363

(sole H2O2 dosage: c(o-PO43–-PWMF)/c(TP0)<2%, not shown in Fig. 2 due to logarithmic

364

representation), which is plausible with regard to their use as bleach stabilizers in bleaching

365

liquors.

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366 367

Fig. 2:

100 90 80 70 60 50 40 30 20 10 0 0.001

PBTC HEDP NTMP EDTMP DTPMP

0.010 0.100 1.000 II c(Fe )/c(H2O2) (mol/mol)

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c(o-PO43–-PWMF)/c(Total P0) (%)

ACCEPTED MANUSCRIPT

10.000

Oxidation of phosphonates (initial concentration of 3 mg/L) to o-PO43– at pH 2.5–3.5 (pH 3.5 was adjusted using H2SO4; the pH could decrease during the reaction phase to a

369

minimum of 2.5) in water as a result of a one-hour treatment with Fenton reagent at a

370

stoichiometrically eightfold H2O2 dosage concentration with varying FeII-H2O2-ratios

371

(cH₂O₂=8·2.125·COD0, COD0 according to Section 1.4). For circled symbols:

372

c(FeII):c(phosphonate)=1:1.

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In Fig. 2, those points indicated by a circle correspond to an equimolar Fe-phosphonate-ratio

374

for each phosphonate. At the equimolar Fe-phosphonate-ratio, only a slight o-PO43– formation

375

was observed for nitrogen-free phosphonates. At higher FeII-H2O2-ratios – thus also at higher

376

FeII-phosphonate-ratios – some of the iron was present in the excess to the phosphonate and

377

therefore not necessarily completely complexed. For polyphosphonates, this resulted in a

378

maximum 20% conversion to o-PO43–. PBTC stuck out with the highest conversion rates of up

379

to 50%. Accordingly, in the presence of Fenton reagent, a preferred oxidation of the nitrogen-

380

free phosphonate PBTC to o-PO43– occurred, whereas in the presence of MnII/O2 (Nowack

381

and Stone, 2000) and O3 (Klinger et al., 1998) aminophosphonates were preferably oxidized.

382

The very reactive hydroxyl radical which is active in the Fenton reaction is very short-lived

383

and therefore does not react selectively with sites of increased electron density such as O2 and

384

O3 (Lee and von Gunten, 2010). Furthermore, Croft et al. (1992) had shown that N-oxides of

385

the aminophosphonates react with hydroxyl radicals to form long-lived nitroxides in which

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ACCEPTED MANUSCRIPT the C-P bonds remain intact, what could explain the low formation extent of o-PO43– in the

387

case of aminophosphonates. It is thus probable that the C-PO(OH)2 groups in the

388

phosphonates represent the most stable component in the molecule to hydroxyl radicals, and

389

thus are the limiting factor in the degradation rate of phosphonates by Fenton reagent.

390

Especially the nitrogen-free phosphonates with three carboxyl groups in the case of PBTC and

391

an alcohol group in the case of HEDP have functional groups which are not present in

392

aminophosphonates. In order to better understand the role of these functional groups with

393

regard to the stability of these phosphonates, more detailed investigations on the degradation

394

pathways of the phosphonates would be required.

395

Molar FeII-H2O2-ratios >1 were associated with lower o-PO43– formation rates. A strong

396

formation of hydroxyl radicals can occur by means of excessively high excess of iron, so that

397

the hydroxyl radicals react essentially only among themselves (recombination). Furthermore,

398

excess Fe2+ ions can act as free-radical scavengers (Barbusiński and Filipek, 2001). As a

399

result, there are less radicals which can oxidize the phosphonate.

400

3.2 UV/FeII, UV/H2O2, UV/FeII/H2O2 in a pure water matrix (Experiment 2)

401

Fig. 3 summarizes the conversion rates of all phosphonates to o-PO43– at different pH ranges

402

in a pure water matrix as a result of a one-hour treatment with various combinations of UV,

403

FeII (equimolar to phosphonate) and H2O2 dosed in a sixteenfold stoichiometric excess (this

404

dosage concentration was chosen to obtain maximum conversion rates to o-PO43–). For this

405

experiment, the pH of the phosphonate-containing sample was adjusted after the possible FeII

406

dosing. Due to the subsequent H2O2 dosing, the pH value decreased only by a maximum of

407

0.3. Only after the one-hour reaction phase partially a stronger pH change was observed

408

predominantly in the alkaline range. The areas grayed out in Fig. 3 thus illustrate the

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ACCEPTED MANUSCRIPT minimum and maximum pH value that could be measured in the corresponding samples

410

during the experiment. 100 90 80 70 60 50 40 30 20 10 1000 90 80 70 60 50 40 30 20 10 100 0 90 80 70 60 50 40 30 20 10 1000 90 80 70 60 50 40 30 20 10 0

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(a) UV

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(d) UV/FeII/H2O2

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c(o-PO43–-PWMF)/c(Total P0) (%)

409

3

PBTC

411 412

Fig. 3:

4

5 HEDP

6

pH

NTMP

7

8 EDTMP

9

10 DTPMP

Oxidation of phosphonates (initial concentration of 3 mg/L) to o-PO43– as a result of a

413

one-hour UV treatment (6 kWh/m³) with combinations of FeII (equimolar to phosphonate)

414

and H2O2 (cH₂O₂=16·2.125·COD0, COD0 according to Section 1.4) at the pH ranges 3.4–

20

ACCEPTED MANUSCRIPT 415

3.6, 4.6–5.0, 6.2–6.9, 7.0–8.0, 8.0–9.5 in water. The initial pH (3.5, 5.0, 6.5, 8.0, 9.5) was

416

adjusted using H2SO4 and NaOH in the phosphonate and FeSO4 containing sample before

417

addition of H2O2. The pH ranges are therefore due to pH change after the H2O2 dosing

418

and during the reaction phase.

None of the phosphonates could be degraded to o-PO43– by UV irradiation alone (Fig. 3a),

420

which underlines their UV resistance in the noncomplexed state. Lesueur et al. (2005) had

421

been able to detect an oxidation of aminophosphonates to o-PO43– even without a catalyst

422

using a medium-pressure UV lamp at pH values between 3 and 10. In their experiments, a

423

significantly higher specific UV energy consumption was used (112 kWh/m³ compared to the

424

experiment described here of 6 kWh/m³), which could explain the oxidation despite the

425

absence of a catalyst.

426

In the presence of FeII, PBTC was very well degraded by UV radiation (up to 80%) while

427

polyphosphonates showed only low degradation rates below 30% (Fig. 3b). For PBTC,

428

highest degradation occurred in the acidic milieu (pH 3.4–5.0), while polyphosphonates at pH

429

3.4–3.6 were not significantly degraded to o-PO43– at all. With increasing pH, the

430

degradability of PBTC decreased slightly and fell sharply in the alkaline range. At this pH

431

range, a precipitation of metal hydroxides was observed for both PBTC and HEDP. However,

432

due to the o-PO43– analysis without previous membrane filtration (o-PO43–WMF), o-PO43– that

433

adsorbed onto Fe(OH)2 or Fe(OH)3 was also determined. The precipitation of iron hydroxide

434

suggests that PBTC and HEDP were partly present in the noncomplexed state. In this state,

435

phosphonates are not degradable solely by UV radiation of 6 kWh/m³ (Fig. 3a). Approx. 10%

436

of NTMP was converted to o-PO43– almost constantly over the entire pH range. The

437

degradation rate of these three phosphonates between pH 5 and 8 decreased according to their

438

number of phosphonate groups (in parentheses): PBTC (1) >> HEDP (2) > NTMP (3).

439

EDTMP (4) and DTPMP (5) proved to be stable against UV radiation of 6 kWh/m³, even in

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ACCEPTED MANUSCRIPT the presence of FeII. Lesueur et al. (2005) had shown a significantly stronger oxidation of the

441

aminophosphonates NTMP, EDTMP and DTPMP to o-PO43– with higher UV performance

442

(112 kWh/m³) and lower pH values (3 and 5–6) than was the case at pH 10. The results of

443

Experiment 2 show that the stable C-PO(OH)2 groups of the phosphonates may also be the

444

limiting factor with respect to the conversion rate of the phosphonates to o-PO43– by UV/FeII

445

as in Experiment 1.

446

All phosphonates were converted to o-PO43– by treatment with UV/H2O2 in the neutral and

447

alkaline pH range between 30 and 50% (Fig. 3c). With increasing pH, the aminophosphonates

448

also showed a slightly increasing oxidation to o-PO43–, which is due either to an improved

449

efficiency of the UV/H2O2 process at higher pH values (Benjamin and Lawler, 2013) or

450

possibly to a better degradability of phosphonate species with a larger negative charge. It is

451

noticeable that EDTMP and DTPMP could be degraded better by UV/H2O2 over the entire

452

tested pH range than by UV/FeII. This means that EDTMP and DTPMP do not necessarily

453

have to be present as complexes in order to be significantly converted to o-PO43– by hydroxyl

454

radicals. The nitrogen-free phosphonates PBTC and HEDP showed almost no pH dependence

455

with respect to their conversion to o-PO43– by UV/H2O2 over the entire investigated range.

456

Interestingly, a small peak with respect to the oxidation to o-PO43– could be observed for both

457

nitrogen-free phosphonates around pH 5. At pH 5, the doubly negatively charged species is

458

predominant in both phosphonates (pKa values of PBTC: 3.74, 4.23, 5.14, 6.80, 9.05 (Liu et

459

al., 2000), pKa values of HEDP: <1, 2.54, 6.97, 11.41 (Carroll and Irani, 1967)). It is possible

460

that this reacts preferably to o-PO43–.

461

While PBTC was the only phosphonate reacting more weakly to o-PO43– by UV/FeII/H2O2

462

with increasing pH, the polyphosphonates showed a tendency towards more pronounced o-

463

PO43– formation, especially with increasing pH (Fig. 3d). For the phosphonates HEDP and

464

EDTMP, the maximum c(o-PO43–WMF)/c(TP0)-ratio was already established above a neutral

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22

ACCEPTED MANUSCRIPT pH value, whereas for the phosphonates NTMP and DTPMP even with the highest pH values

466

(pH 8.0–9.5) the conversion to o-PO43– still increased further. PBTC was by far the most

467

easily degradable phosphonate (70–80% conversion to o-PO43–) in the acidic pH range, while

468

the conversion rate to o-PO43– of all other phosphonates at pH 3.4–3.6 was, at most, 35%. At

469

this pH, the oxidation of the phosphonates to o-PO43– correlated with their phosphonate group

470

number (in parentheses): PBTC (1) >> HEDP (2) > NTMP (3) ≈ EDTMP (4) > DTPMP (5).

471

In the neutral pH range, the extent of the conversion of the phosphonates to o-PO43– was as

472

follows: HEDP (2) ≈ PBTC (1) > EDTMP (4) ≈ NTMP (3) > DTPMP (5). At pH >6, HEDP

473

had the highest degradability to o-PO43–, while here the maximum conversion of all other

474

phosphonates to o-PO43– was between 50 and 60%.

475

The decreasing degradation of PBTC to o-PO43– with increasing pH can be explained by the

476

precipitation of iron hydroxide. This precipitation occurs rather for nitrogen-free

477

phosphonates, since these form less stable complexes with metals compared to

478

aminophosphonates (Knepper, 2003). Thus, at higher pH values, free PBTC is present, which

479

is significantly less degradable by UV/H2O2 compared to the FeII-PBTC complex (Fig. 3c). In

480

the case of polyphosphonates, degradation by UV/FeII/H2O2 predominantly occurred in the

481

alkaline range. In this pH milieu, polyphosphonates preferably complex the dosed FeII, so that

482

it can be assumed that the improved degradation with alkaline pH values is essentially due to

483

the fact that the phosphonates are more easily attacked in a complexed form by hydroxyl

484

radicals (at least for HEDP and NTMP). Furthermore, this experiment clearly showed that

485

phosphonates are converted less to o-PO43– by means of UV/FeII and UV/FeII/H2O2 the larger

486

their number of phosphonate groups is. As already explained in Section 3.1, the C-PO(OH)2

487

groups in the phosphonates are considered to be the most stable component in the molecule,

488

so that a high number of phosphonate groups also limits the degradation rate of the

489

phosphonates. In conclusion, the treatment with UV/FeII/H2O2 (Photo-Fenton) in water

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ACCEPTED MANUSCRIPT proved to be a good means of oxidizing all phosphonates. The next experiment (Exp. 3)

491

should show whether this good oxidisability is also present in wastewater where UV light

492

absorbing turbidity may be present or oxidation of other readily degradable organic

493

components as well as free-radical scavengers can occur.

494

3.3 UV/FeII, UV/H2O2, UV/FeII/H2O2 in cooling tower effluent (Experiment 3)

495

Fig. 4 summarizes the results of an experiment in which cooling tower effluent spiked with

496

3 mg/L PBTC was treated with UV, UV/FeII, UV/H2O2 and UV/FeII/H2O2 at pH 3.5. The

497

results of the oxidation extent of organically bound phosphorus to o-PO43– (o-PO43–WMF) after

498

the one-hour reaction phase, as well as the phosphorus fractions immediately following

499

membrane filtration (‟MF”) at pH 3.5, and the same for the sample neutralized immediately

500

after the reaction phase and also membrane-filtered, are shown.

M AN U

0.50 0.45 0.40 0.35 0.30 0.25 0.20 0.15 0.10 0.05 0.00

Total P0

o-PO43–-P

TE D

mg/L

SC

RI PT

490

503 504 505 506

Raw sample

Fig. 4:

UV

o-PO43–-P (MF) (pH 3.5)

UV/FeII

Total P (MF) (pH 3.5) UV/H2O2

o-PO43–-P (MF) (pH 7.0)

Total P (MF) (pH 7.0)

UV/FeII/H2O2

Formation and concentration of o-PO43– and TP in cooling tower effluent as a result of a

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Total P0 and o-PO43–-P o-PO43–-PWMF (raw sample) (homogeneous) (pH 3.5)

one-hour UV radiation (3 kWh/m³) at pH 3.5 (adjusted with H2SO4) with 4.6 mg/L FeII and 30 mg/L H2O2 of a membrane-filtered (MF) sample before (pH 3.5) and after neutralization using NaOH (pH 7.0). No significant pH drop could be observed during the contact time.

507

The oxidation extent of organically bound phosphorus to o-PO43– was similar to the results

508

found in Experiment 2 with PBTC dissolved in pure water (UV/FeII ≥ UV/FeII/H2O2 >

509

UV/H2O2 > UV). An oxidation of PBTC to o-PO43– already occurred during treatment with 24

ACCEPTED MANUSCRIPT UV light without chemical dosage, whereas no oxidation of PBTC to o-PO43– had been

511

observed by sole UV radiation in the pure water test, even at a higher UV energy consumption

512

(6 kWh/m³). The UV oxidation of PBTC can be explained by metals present in the cooling

513

tower effluent (e.g., Fe and Al, Section 2.1.2). Thus, a photolysis catalyzed by metals could

514

also occur without FeII dosage, although in this case the o-PO43– formation was only half as

515

high as with FeII dosage.

516

In the sample treated with UV/FeII/H2O2 could a reduced concentration of o-PO43– and TP

517

after membrane filtration at pH 3.5 be seen. In this batch, an oxidation of ferrous iron to ferric

518

iron was evident by H2O2. In contrast to FeII, FeIII precipitates at pH 3.5 (Barrera-Díaz et al.,

519

2003), resulting in adsorbent for o-PO43– and organically bound phosphorus. Furthermore,

520

non-soluble FePO4 could form.

521

The neutralization immediately after the one-hour reaction phase resulted in an almost

522

complete removal of the o-PO43– formed in the samples treated with UV/FeII and

523

UV/FeII/H2O2 and a total of 80–85% TP decrease. FeII precipitates above pH 8.7 at the dosed

524

concentration of 4.6 mg/L FeII (Barrera-Díaz et al., 2003). Thus, in both samples, FeIII (mainly

525

the adsorbent iron hydroxide and non-soluble FePO4) had to be formed. In the sample treated

526

with UV/FeII/H2O2, this was done predominantly by the oxidation of FeII to FeIII by H2O2, as

527

already described. In the sample treated with UV/FeII, this oxidation must have taken place

528

during the neutralization by O2 dissolved in the sample.

529

The samples without addition of FeII (UV, UV/H2O2) retained the concentration of dissolved

530

o-PO43– largely after neutralization. The TP could only be slightly removed by neutralization

531

since there was no iron as precipitant.

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ACCEPTED MANUSCRIPT 3.4 FeII/H2O2 in phosphonate production wastewater (Experiment 4)

533

During the reaction phase (c(FeII)/c(H2O2)=0.08 [g/g]) at pH 2.1–2.5, only about 15% of the

534

organically bound phosphorus of phosphonate production wastewater were oxidized to o-

535

PO43– (50 out of 350 mg/L P) (Fig. 5 left). When the wastewater was treated with Fenton

536

reagent in the course of the neutralization variant 1 (direct neutralization following the

537

reaction phase), an almost 100% TP (supernatant) and o-PO43–-PWMF (supernatant) decrease

538

occurred at an H2O2 concentration of 10.2 g/L (corresponding to the stoichiometrically

539

required oxidant concentration, thus 100% H2O2) and 0.82 g/L FeII. With the neutralization

540

variant 2 (reaction, then sedimentation (acidic), then neutralization of the supernatant and

541

sedimentation (neutral)), an almost 100% TP decrease was already achieved at 5.1 g/L H2O2

542

(50%) and 0.41 g/L FeII.

350 300 250 200 150 100 50 0

547 548

0.8

1.0

0

25

50

75

100

0.4

FeII (g/L) 0.6 0.8

125

1.0

0

0.2

o-PO43–-PWMF (supernatant)

0

25 50 75 100 125 cH₂₂O₂₂ · 100 % · COD0–1 · 2.125–1 (%) Supernatant N2a (acidic)

0

0.4

0.6

0.8

1.0

Total P (supernatant)

25

50

75

100

125

150

Supernatant N2b (neutral)

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Fig. 5:

0.2

Formation and concentration of o-PO43– and TP after treatment of phosphonate production wastewater with Fenton reagent at a constant c(FeII)/c(H2O2) ratio of 0.08

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0.6

Supernatant N1 (neutral)

543

545

0.4

o-PO43–-PWMF (homogeneous)

0

544

0.2

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mg/L

0

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[g/g] (proportional increase of FeII and H2O2) and pH 2.1–2.5 (pH 2.5 adjusted with H2SO4; the pH could decrease during the reaction phase to a minimum of 2.1) according to the neutralization variants 1 and 2. Raw sample: 350 mg/L TP, 4.8 g/L COD0.

549

The low formation extent of o-PO43– was in accordance with the o-PO43– formation found in

550

Experiment 1 (pure water) when no UV radiation was applied. The low o-PO43– formation

551

indicates that complexing agents may have remained in the sample after the reaction phase, 26

ACCEPTED MANUSCRIPT probably predominantly in the form of phosphonates and their degradation products capable

553

of complex formation. Since the reaction rate of the reduction of FeIII to FeII (Equation 2) is

554

slower than the reaction rate of the oxidation of FeII to FeIII (Equation 1), over time an

555

accumulation of FeIII and thus the precipitate Fe(OH)3 in the samples with Fenton reagent

556

occurs. Although, as mentioned, there were probably still complexing agents in the sample

557

after the reaction phase, at pH 2.1–2.5, iron hydroxide sludge could precipitate due to the

558

decreased effectiveness of these complexing agents because of the low pH. Furthermore, o-

559

PO43–, phosphonates, as well as their degradation products have the best adsorption affinity

560

for iron hydroxides at low pH values (Nowack and Stone, 1999). When the sample was

561

neutralized directly after the reaction phase (neutralization variant 1), the residual complexing

562

agents could have interfered with the precipitation of iron hydroxide, and thus, with the

563

formation of adsorbent for the TP, because the increase in pH led to a more pronounced

564

effectiveness of the residual complexing agents. The complete absence of a TP elimination in

565

a small dosage concentration range (below 30 %) and mere part removal at values between 50

566

and 100 % (stoichiometric ratio) when performing the neutralization variant 1 indicated that

567

clearly. Thus, the neutralization variant 2 performed better than neutralization variant 1 since

568

most of the complexing agents were separated from the sample when a sludge separation

569

occurred at an acidic pH.

570

The similarly high TP concentrations in both of the supernatants (N2a and N2b) resulting

571

from the neutralization variant 2 show that the phosphorus compounds were removed mainly

572

in the first sedimentation stage (‟supernatant N2a (acidic)”), thus at acidic pH. Phosphorus

573

compounds which had not been removed in the first sedimentation stage were separated by

574

iron hydroxide flocks at the latest produced during neutralization. The slightly higher o-PO43–-

575

PWMF concentration in the supernatant of neutralization variant 1 (‟supernatant N1 (neutral)”)

576

compared to the concentration prior to the sedimentation (‟o-PO43–-PWMF (homogeneous)”)

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27

ACCEPTED MANUSCRIPT can be explained by residual H2O2 in the sample, which could further oxidize organic

578

phosphorus compounds during the sedimentation phase.

579

4 Conclusions

580

The degradability of the phosphonates PBTC, HEDP, NTMP, EDTMP and DTPMP by means

581

of metal-catalyzed photolysis (UV/FeII), Fenton method (FeII/H2O2) and Photo-Fenton method

582

(UV/FeII/H2O2) was investigated.

583

This work has shown that polyphosphonates can be converted to phosphate by a maximum of

584

20% by means of Fenton reagent at pH 2.5–3.5 in a pure water matrix even at a very high

585

excess of H2O2 (also high Fe excess, thus, influence of complex formation is marginal).

586

Therefore, in order to oxidize phosphonates to a higher extent, the oxidation process has to be

587

supported by UV radiation (whether in the UV/FeII or in the UV/FeII/H2O2 process).

588

Interestingly, the degradation of polyphosphonates in the UV/FeII/H2O2 process increased

589

with increasing pH. This is in contrast to the general assumption that the Fenton process

590

works best at low pH values. Further tests with wastewater would therefore have to

591

investigate whether, in the special case of the desired phosphonate oxidation, higher reaction

592

pH values than usual lead to better results on the TP removal.

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CATALYZED PHOTOLYSIS: catalyst: FeSO4

UV-FSR: UV free surface reactor NT: neutralization tank ST: sludge separation tank RT: reaction tank

H 2 SO 4/ NaOH

UV

H 2SO 4 /NaOH

concentrate

ST UV-FSR

NT

phosphonate reduced wastewater

FENTON: Ca(OH) 2/NaOH

H 2SO 4

raw wastewater

593

sludge

Fenton reagent: FeSO4 /H 2 O 2

ST

ST NT

RT

sludge (acidic)

sludge (neutral)

28

phosphonate reduced wastewater

ACCEPTED MANUSCRIPT 594

Fig. 6:

Proposal for the implementation of a continuously operated wastewater treatment plant

595

according to the UV/FeII and Fenton principle for industrial wastewater containing

596

phosphonates.

An experiment with clear, Ca2+-rich, organically little loaded and PBTC-containing cooling

598

water could confirm the results from the pure water experiments with PBTC. Accordingly, the

599

organically bound phosphorus was successfully degraded to phosphate in this wastewater by

600

the UV/FeII and the UV/FeII/H2O2 method. Furthermore, the TP removal performance

601

between the UV/FeII and the UV/FeII/H2O2 method was hardly different. It was found that the

602

dosed iron precipitates even when no H2O2 is dosed in the UV/FeII process through pH

603

neutralization carried out following the reaction phase at pH 3.5 if, during the reaction phase,

604

the phosphonate is decomposed completely or decomposed into reaction products which are

605

no longer capable of complexation. Thus, neutralization can take place immediately following

606

the reaction phase (neutralization variant 1) (Fig. 6). The o-PO43– formed during the reaction

607

phase is thereby discharged from the wastewater via filtration or sedimentation. The UV/FeII

608

process is very selective and is recommended only if the organically bound phosphorus

609

fraction in the wastewater is mainly composed of compounds susceptible to metal-catalyzed

610

photolysis such as phosphonates. Furthermore, only a very small sludge amount is to be

611

expected due to the fact that only very low amounts of iron are sufficient for successful

612

oxidation of the phosphonate. Accordingly, the UV/FeII method is particularly suitable for

613

concentrates with nitrogen-free phosphonates, a low turbidity and COD load and no

614

competing complexing agents.

615

The low oxidation extent of polyphosphonates to phosphate by means of sole Fenton reagent

616

(no UV radiation) in pure water could also be found for a phosphonate-containing, organically

617

highly polluted wastewater matrix (phosphonate production wastewater). However, it was

618

discovered that complete conversion of the phosphonate to phosphate was not absolutely

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29

ACCEPTED MANUSCRIPT necessary for removal of TP in this kind of wastewater. Since phosphonates adsorb on the

620

Fenton sludge, they are also removed via the sludge separation (sedimentation or filtration).

621

The experiments showed that the following procedure should be applied for a chemical-

622

saving application of the Fenton method when applied to organically polluted wastewater

623

without support of UV radiation (neutralization variant 2, Fig. 6): reaction → sludge

624

separation (acidic) → neutralization of the supernatant → sludge separation (neutral; in the

625

case of full precipitation of added FeII in the acidic sludge separation no sludge separation

626

could be necessary here). If, on the other hand, neutralization takes place immediately

627

following the reaction phase, phosphonates or other complexing agents remaining in the

628

sample after the reaction phase can interfere with the precipitation of iron hydroxide due to

629

their higher effectiveness as complexing agents at higher pH. Thereby, less adsorbent is

630

present for the removal of phosphonates and their degradation products, which, moreover,

631

exhibit a weaker adsorption affinity for iron hydroxide in the neutral and alkaline pH range

632

than in the acidic pH range (Nowack and Stone, 1999).

633

The UV lamp used in our experiments was too weak for very turbid, heavily organically

634

polluted wastewaters, which is why further investigations with stronger UV lamps have to be

635

carried out to find out whether there is a more pronounced phosphonate degradation by the

636

Photo-Fenton method compared to the Fenton method in these kind of wastewaters.

637

Acknowledgements

638

The authors are grateful for the financial support by the Willy-Hager-Stiftung, Stuttgart. We

639

would also like to thank the employees of Zschimmer & Schwarz Mohsdorf GmbH & Co. KG

640

for providing phosphonate samples.

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ACCEPTED MANUSCRIPT Dr. Eduard Rott INSTITUTE FOR SANITARY ENGINEERING, WATER QUALITY AND SOLID WASTE MANAGEMENT

UNIVERSITÄT STUTTGART

Institute for Sanitary Engineering, Water Quality and Solid Waste Management ● Bandtäle 2 ● 70569 Stuttgart

D-70569 Stuttgart (Büsnau) Bandtäle 2

Editorial Board Water Research To whom it may concern

Phone Fax

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+49 711 685 – 60497 +49 711 685 – 63729

E-Mail: [email protected]

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Date th February 7 2017

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Subject: Paper highlights

Phosphonates could be degraded well using UV/FeII and UV/Fenton.

•

The stability of phosphonates depends on their number of phosphonate groups.

•

UV/FeII treatment worked well on clear, less organically polluted wastewater.

•

For a better Fenton performance, the acidic supernatant should be neutralized.

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•