Formation and Fate of Point-Source Nonextractable DDT-Related

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Formation and fate of point-source non-extractable DDT-related compounds on their environmental aquatic-terrestrial pathway Xiaojing Zhu, Larissa Dsikowitzky, Sebastian Kucher, Mathias Ricking, and Jan Schwarzbauer Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.8b06018 • Publication Date (Web): 04 Jan 2019 Downloaded from http://pubs.acs.org on January 8, 2019

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Environmental Science & Technology

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Formation and fate of point-source non-extractable DDT-related compounds on their

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environmental aquatic-terrestrial pathway

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Xiaojing Zhu1, Larissa Dsikowitzky1, Sebastian Kucher1, Mathias Ricking2, Jan Schwarzbauer1*

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1 Institute

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20, 52064 Aachen, Germany

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2 Department of Earth Sciences, Freie Universitäte Berlin, Malteser Str. 74-100, 12249 Berlin, Germany

of Geology and Geochemistry of Petroleum and Coal, RWTH Aachen University, Lochnerstr. 4-

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ABSTRACT: Non-extractable residues (NER) are pollutants incorporated into the matrix of natural solid

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matter via different binding mechanisms. They can become bioavailable or remobilize during physical-

* Corresponding author. E-mail addresses: [email protected] (J. Schwarzbauer), [email protected] (X. Zhu), [email protected] (L. Dsikowitzky), [email protected] (S. Kucher), [email protected] (M. Ricking) 1

ACS Paragon Plus Environment


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chemical changes of the surrounding conditions and should thus not be neglected in environmental risk

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assessment.

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dichlorodiphenyltrichloroethane; and its metabolites) were treated with solvent extraction, sequential

13

chemical degradation and thermochemolysis to study the fate of NER-DDX along different environmental

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aquatic-terrestrial pathways. The results showed that DDT and its first degradation products, DDD

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(dichlorodiphenyldichloroethane) and DDE (dichlorodiphenyldichloroethylene), were dominant in the free

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extractable fraction, whereas DDM (dichlorodiphenylmethane), DBP (dichlorobenzophenone) and DDA

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(dichlorodiphenylacetic acid) were observed primarily after chemical degradation. The detection of DDA,

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DDMUBr (bis(p-chlorophenyl)-bromoethylene), DDPU (bis(p-chlorophenyl)-propene) and DDPS (bis(p-

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chlorophenyl)-propane) after chemical treatments evidenced the covalent bindings between these DDXs and

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the organic matrix. The identified NER-DDXs were categorized into three groups according to the three-

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step degradation process of DDT. Their distribution along the different pathways demonstrated significant

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specificity. Based on the obtained results, a conceptual model of the fate of NER-DDXs on their different

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environmental aquatic-terrestrial pathways is proposed. This model provides basic knowledge for risk

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assessment and remediation of both extractable and non-extractable DDT-related contaminations.

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Keywords: DDT-related compounds; non-extractable residues; sequential chemical degradation;

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environmental pathway

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1. Introduction

Sediments,

soils

and

groundwater

sludge

contaminated

with

DDXs

(DDT,

28

Anthropogenic organic contaminants from industrial, municipal and agricultural emissions are ubiquitous

29

in the aqueous phase, the atmosphere, in natural solid substances as well as in (micro)organisms. They can

30

further be sequestrated as non-extractable residues (NER), which are immobilized in the natural solid phase1.

31

NER, also known as bound residues, a portion of organic substances not freely available, either temporarily

32

or permanently, because of their strong association with the environmental particulate matter (e.g. soil,

33

sediment, aquatic suspended particular matter), were first recognized in the 1960s2,3. NER cannot be

34

released from solid matrix after mild solvent extractions that do not cause physicochemical changes to the 2


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compounds or the underlying matrix4,5. Non-extractable contaminants and their metabolites are not

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detectable with normal analytical methods because of their incorporation into solid matrix via physical and

37

chemical interactions, but will get partially bioavailable and can be remobilized by changes in surrounding

38

conditions (e.g. low-weight organic acids6 and biological activities7 may result in release). Thus, the NER-

39

portion of anthropogenic pollutants should not be neglected in environmental risk assessment. The

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environmental fate and transport of NER compounds are quite different regarding their aquatic or terrestrial

41

pathways. In aquatic systems, sediment, suspended particular matter and colloidal organic matter are the

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major targets for NER formation. In terrestrial systems, soil is the principal solid sink for NER. Therefore,

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the effects of these different environmental pathways on the fate of non-extractable contaminants need to

44

be elucidated.

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Dichlorodiphenyltrichloroethane (DDT) was widely used as pesticide since the 1940s and was banned in

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Europe in the 1970s and 1980s because of its toxicity and environmental persistence. DDT is associated

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with serious risks to the environment and human health, including carcinogenesis, endocrine disruption and

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estrogenic

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dichlorodiphenyldichloroethane (DDD) and dichlorodiphenyldichloroethylene (DDE), which could be the

50

congeneric impurities of technical DDT at very low amounts10–12, are still frequently worldwide detectable

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in considerable amounts even decades after the DDT prohibition8. Those metabolites are also reported to be

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environmentally and biologically harmful13–15. The ongoing usage of the DDT related pesticide dicofol, as

53

well as specialty applications of DDT in antifouling paints and limited disease vector control efforts, result

54

in substantial emission of DDT/DDXs that are cause for continued environmental concern16,17.

55

Consequently, knowledge on non-extractable DDXs is still necessary for both risk assessment and

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remediation actions.

57

action8,9.

However,

DDT

and

its

metabolites

(DDXs),

particularly

The formation potential and bioavailability of non-extractable DDT in soil have been investigated for

58

decades, and the

59

NER-DDT and NER-DDE are formed in soil under different matrix conditions, can be released to a low

14C-labelling

was the mainstream approach18–22. With this approach it was proved that

3



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extent and are potentially bioavailable for plants. Sequential chemical degradation has been applied since

61

2003 to release sedimentary bound DDXs, and a variety of DDT metabolites have been detectable as NER23–

62

27.

63

Covalent linkage of DDA (dichlorodiphenylacetic acid) to soil and sediment, which forms NER-DDA was

64

proved by Kalathoor et al. in 201529. Current knowledge on the fate of NER-DDXs (data on their formation

65

are restricted to field samples), especially those beside DDT, DDD and DDE, in natural solid ambient is

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quite limited, and there is even less research with respect to the aquatic-terrestrial pathway since most of the

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aforementioned studies solely considered sediment and/or soil. For that reason, further efforts must be

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carried out to span this knowledge gap.

More recently, pyrolysis and thermochemolysis have been conducted to obtain further DDT derivatives28.

69

So far, previous investigations (although restricted on sediments) revealed that DDXs with higher

70

degradation degree would be more abundant as NER with some molecular structural selectiveness. Ester,

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ether and carbon-carbon bonds were presumed as the possible covalent linkages of DDXs to natural organic

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macromolecules. For a better understanding of the fate of NER-DDXs, soil, sediment and groundwater

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sludge samples from two highly DDT-contaminated areas were therefore collected for this study to

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investigate the distribution variation of NER-DDXs. Solvent extraction and sequential chemical degradation,

75

which have been well developed23,25,27,28, were applied on these samples to obtain extractable and non-

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extractable DDXs, respectively. TMAH (tetramethylammonium hydroxide) thermochemolysis was also

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conducted to obtain further NER-DDT derivatives. This work provides important insights on the fate of

78

DDXs and their different environmental pathways. Such basic knowledge is a prerequisite for risk

79

assessment and remediation.

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2. Material and methods

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2.1. Samples

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Six subaquatic sediment (TC-S1 to TC-S6) and four groundwater sludge (TC-G1 to TC-G4) samples

83

were collected from the Teltow Canal in Berlin, Germany, a highly DDT-contaminated site. A chemical 4


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production plant, Berlin Chemie, which produced extraordinarily high amounts of DDT and other pesticides

85

was once located at the beginning of Teltow Canal30. The groundwater was intensively contaminated by the

86

point source because of a bank filtration area of a former drinking water production plant next to it23.

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Two soils (BFW-SOIL1 (5 – 20 cm topsoil) and BFW-SOIL2 (20 – 40 cm subsoil), from a heavily loaded

88

site) and two subaquatic sediment (BFW-S1 and BFW-S2) samples were obtained from the river system of

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northeastern Bitterfeld-Wolfen, Germany, an industrial megacity, where a wide range of chemical products

90

including pesticides were manufactured31. For further information on sampling areas and methods see our

91

previous works23,32,33. Sampling locations and information on the samples are shown in Fig. 1.

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Fig. 1. Sampling locations at the Teltow Canal and in Bitterfeld-Wolfen. Total organic matter (foc, dry weight) was determined by

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loss on ignition.

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2.2. Extraction

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Prior to chemical degradation, 10 g of pre-air dried sample aliquots were pre-extracted sequentially with

97

acetone, mixtures of acetone and n-hexane (1: 1) and n-hexane by accelerated solvent extraction (Dionex 5



98

ASE 150, Thermo Fisher Scientific, Waltham, MA, USA) to obtain the free extractable fraction (EF). The

99

extracts from the aforementioned three solvents were combined and rotary evaporated at room temperature

100

to reduce the volume. After drying by anhydrous sodium sulfate (Na2SO4) and the removal of elemental

101

sulfur with activated cooper powder, each extract was separated into six factions by micro silica column

102

chromatography using mixtures of n-pentane, dichloromethane and methanol as eluents, as formerly

103

described in detail25. A surrogate standard containing 5.82 ng μL-1 4’-fluoroacetophenone, 6.28 ng μL-1 d10-

104

benzophenone and 6.03 ng μL-1 d34-hexadecane was added to each fraction before gas chromatography-

105

mass spectrometry (GC-MS) analyses.

106

2.3. Chemical degradation

107

Chemical degradation was carried out according to Schwarzbauer et al.27,34 and Kronimus et al.24,28 to

108

release the non-extractable residues. Pre-extracted samples were sequentially treated with KOH (potassium

109

hydroxide)/methanol, boron tribromide (BBr3) and ruthenium tetroxide (RuO4). TMAH thermochemolysis

110

was also applied on the pre-extracted samples.

111

2.3.1. Alkaline hydrolysis and TMAH thermochemolysis

112

5.0 g of each pre-extracted sample was placed in a closeable centrifuge glass tube with 2.5 g KOH, 2 mL

113

ultrapure water and 20 mL methanol. After 15 min treatment in an ultrasonic bath, the closed tube was

114

heated at 105 ºC for 24 h. After cooling and decanting the sample solution into a separating funnel, 50 mL

115

ultrapure water was added, and the mixture was acidified to pH 4 to 5. Thereafter, the solution was extracted

116

three times with 30 mL dichloromethane. The combined organic solution was dried and desulfurized.

117

An aliquot of 150 mg pre-extracted sample was transferred into a DURAN glass tube (8 mm i.d. × 200

118

mm length) with 200 μL TMAH solution (0.1 mol L-1 in methanol). After cautious hand-shaking of the tube,

119

the methanol was evaporated under a gentle nitrogen stream. The tube was sealed and heated at 250 ºC for

120

40 min and hereafter cooled and opened. Then, 1 mL n-hexane was added, and the mixture was sonicated

121

for 3 min. After decanting the extract into a glass flask, the residue was extracted again with 1 mL n-hexane. 6


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A second round of extraction was repeated with acetone. Then, 5 mL n-hexane was added to the combined

123

extract and the solution was rotary evaporated to remove the acetone. The condensed extract was dried and

124

desulfurized.

125 126

After fractionation with micro silica column chromatography, the surrogate standard solution (Section 2.2) was added before GC-MS analysis.

127

2.3.2. BBr3 treatment

128

A BBr3 solution (10 mL, 1.0 mol L-1 in dichloromethane) was placed with a 1.0 g aliquot of an alkaline

129

hydrolyzed sample into a glass centrifuge tube. This mixture was stirred for 30 min at room temperature,

130

sonicated for 15 min and stirred again for 24 h. This procedure was repeated, and subsequently the mixture

131

was sonicated for 15 min. Diethyl ether (10 mL) and 5 mL ultrapure water were added, and the tube was

132

centrifuged for 10 min at 4000 g. The supernatant was decanted into a separating funnel and washed twice

133

with 5 mL of ultrapure water. The organic layer was then evaporated, dried, desulfurized, fractionated and

134

the surrogate standard was added.

135

2.3.3. RuO4 oxidation

136

A 500 mg aliquot of BBr3-treated sample was added into a glass centrifuge tube together with 5 mg

137

sodium periodate, 5 mg RuO4, 8 mL acetonitrile and 8 mL carbon tetrachloride. The mixture was stirred for

138

4 h at room temperature. Afterwards, 50 μL methanol and two drops concentrated sulfuric acid were added

139

to stop the reaction. The tube was centrifuged for 10 min at 4000 g and the supernatant was decanted into a

140

separating funnel. The remaining solid residue was washed again with 3 mL carbon tetrachloride. Ultrapure

141

water (5 mL) was added, and the combined supernatant was washed five times with 10 mL diethyl ether.

142

After evaporation to 0.5 mL, the organic extract was washed again with 0.5 mL of saturated sodium

143

thiosulfate pentahydrate solution. The obtained extract was then dried, desulfurized, fractionated and spiked

144

with the surrogate standard.

7



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2.4. GC-MS analysis

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GC-MS analyses were performed on a Finnigan PolarisQ ion trap mass spectrometer linked to a trace gas

147

chromatograph (Thermo Finnigan) equipped with a 30 m × 0.25 mm i.d. × 0.25 μm film zebron ZB-5 fused

148

silica capillary column (Phenomenex, Aschaffenburg). The GC oven was programmed as follows: 3 min

149

isothermal time at 60 ºC, followed by heating at 3 ºC min-1 to 310 ºC and held for 20 min. The injection

150

volume was 1 μL in splitless mode at 270 ºC injector temperature. Carrier gas was helium set to a velocity

151

of 30 cm s-1. The mass spectrometer was operated in an EI+ full scan mode scanning from 50 to 650 m/z

152

with a scan time of 0.58 s. Ion source temperature was set at 200 ºC.

153

DDX identification was conducted by comparison of the detected mass spectra with mass spectral libraries

154

(NIST/EPA/NIH Mass Spectral Library NIST14, Wiley/NBS Registry of Mass Spectral Data, 7th ed.) and

155

was verified by reference compounds. Quantification was carried out by integrating the peak areas of

156

selected ion chromatograms from the total ion current of a measured sample under consideration of the

157

individual response factors of the GC-MS device. Response factors were determined from linear regression

158

functions based on calibration measurements with different concentrations of authentic reference materials

159

(5 points, concentrations ranged within the linear detection range23,35). The surrogate standard was used for

160

the correction of injection volume and sample volume inaccuracies. The characteristic ions used for DDX

161

quantification are listed in Table S1.

162

3. Results and discussion

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3.1. Occurrence and concentrations of extractable/non-extractable DDXs in subaquatic sediments, soils

164

and groundwater sludge

165

The DDX concentrations in the sample extracts and the sample residues after each chemical degradation

166

step are shown in Fig. 2. All concentration values are given in detail in the Supporting Information.

167

Interestingly, DDT and its first two anaerobic degradation products, DDD and DDE, were most prominent

168

only in the free extractable fraction (EF) of all sample types. A huge amount of DDT (26000 nmol kg-1) was 8


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detected in the EF of sample TC-S1, at the chemical plant outlet to the Teltow Canal. In contrast to samples

170

TC-S2 and TC-S3 from the same location, its foc (total organic matter) was quite low so that DDT occupied

171

even one-third of it. The production and discharge of DDT at this site was stopped decades ago, and thus

172

the sampling site TC-S1 might be an occasional point where DDT largely accumulated and microbial

173

degradation was strongly inhibited.

174

Alkaline hydrolysis releases compounds incorporated by ester bonds. Further compounds bound by ester

175

and ether bonds to the solid matrix can be obtained by BBr3 treatment. RuO4 oxidation, finally, attacks

176

aromatic structures and activated carbon-carbon bonds24. Consistent with previous studies27,34,

177

dichlorodiphenylmethane (DDM), dichlorobenzophenone (DBP) and DDA were the main products released

178

after sequential chemical degradation (see Fig. 2). DDA was only released after alkaline hydrolysis and

179

BBr3 treatment (Fig. 2 b,c), illustrating that the compound is covalently bound to the solid matrix via ester

180

bonds. A new DDT-related compound, whose mass spectrum (Fig. 5) did not match to the mass spectra of

181

any database or of any previous study, was detected in samples TC-S2 and TC-S3 after BBr3 treatment. This

182

compound was tentatively identified as bis(p-chlorophenyl)-bromoethylene (DDMUBr) after comparison

183

with the mass spectrum of DDMU (bis(p-chlorophenyl)-chloroethylene). This compound originates very

184

likely from the cleavage of an ether bond between DDNU and the organic matrix by BBr3 (proposed original

185

structure and mechanism see Fig. 5). Moreover, it was quite interesting that DBP was dominantly detected

186

after RuO4 oxidation. Its possible origins could be the oxidation of other DDXs (e.g. DDNU, DDM and

187

DBH, referring to Memeo et al.36 and Schouten et al.37), a covalently bound DDX obtained by bond cleavage

188

(depicted in Fig. 5, referring to Memeo et al.36, Rup et al.38 and Schouten et al.37, and this origin is more

189

reasonable since former chemical processes have already released a considerable amount of humic moieties)

190

and/or the entrapped DBP released after the structural decomposition of the organic matrix.

191

It should be noted that the distribution of DDXs in the free extractable and hydrolysis released fractions

192

from the groundwater sludge was quite unique as compared to those in soils and sediments (Fig. 2a). Bis(p-

193

chlorophenyl)-acetonitrile (DDCN) was detected at a considerably high amount in the free extractable 9



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fraction of groundwater sludge. Its specific position in the DDT degradation pathway is to date unknown,

195

but it is surely produced under anaerobic conditions30,39,40. Those sludge samples were collected only after

196

physical filtration of groundwater without any biological or chemical treatments. A relatively low DDCN

197

concentration was also detected as a free extractable component in sediments from both sampling areas. The

198

polarity of DDCN41 might be a reason, such that it is mobile in the aqueous phase and could be freely

199

transported into groundwater. DDCN was also reported in the EF of submarine sediments25 and riverine

200

sediments24. Relatively high amounts of DDT and DDD were released from groundwater sludge after

201

hydrolysis rather than from the other two solid substances. This could be attributed to the high foc level of

202

groundwater sludge (see Fig. 1), in which DDT and DDD could temporarily be adsorbed and entrapped.

203

And they could get released after the destruction of solid/colloidal organic matter by the hydrolysis

204

procedure. Groundwater samples from the same polluted area were analyzed by Frische et al in 201035, and

205

DDT, DDD, DDE, DDA, DDCN, DDMU and DBP were detected.

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Fig. 2. DDX concentrations in sample material from two highly DDT-contaminated sites in Germany. All concentrations are given

208

in nmol g-1 dry weight. Figure 2a) shows the freely extractable pollutants and Figures 2b), c) and d) the pollutants released after

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different chemical degradation steps in sequence. The dominant DDXs in each step are marked by yellow bars and the compounds

210

with considerable amount observed exclusively in groundwater sludge are marked by green dash lines.

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Non-extractable DDXs released after TMAH thermochemolysis are shown in Fig. 3a. The most abundant

212

products were DDM and DBP, concentrations of which were higher than of those from sequential chemical

213

degradation. Yet total molar concentrations of DDXs obtained by the two different degradation procedures

214

were basically equal. This indicates that several DDXs are transformed to DDM and/or DBP (both could be

215

formed under natural conditions) during thermochemolysis and/or pydrolysis, which was also proved by

216

Kronimus et al.28.

217

Apart from those well-known DDT metabolites, small amounts of bis(p-chlorophenyl)-propene (DDPU)

218

and bis(p-chlorophenyl)-propane (DDPS) (mass spectra see Fig. 3b) were detectable at the chemical plant

219

outlet to the Teltow Canal. These two compounds were firstly and only reported by Kronimus et al. in 200628

220

in sediments from the Teltow Canal (at the emission point) after TMAH thermochemolysis. DDPU was also

221

detectable after pyrolysis of the same samples. They suggested that DDPU and DDPS were incorporated

222

into the matrix by carbon-carbon bonds, and that the additional carbon atoms might belong to the linked

223

macromolecules. However, in our study these two compounds were only detectable at TC-S2 and TC-S3,

224

and at significantly lower concentrations than other dominant DDXs. There are two reasonable assumptions

225

to explain the absence of DDPU and DDPS in the other samples. The first one is that they were solely

226

formed at the emission point with specific ambient conditions such as microbial activity and natural organic

227

matter makeup/concentration. Another possible reason is that the concentrations of DDPU and DDPS in the

228

rest of the samples were too low to be detected.

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229 230

Fig. 3. Quantitative results of DDXs released by TMAH thermochemolysis (a). The dominant DDXs are marked by yellow bars.

231

Fig. 3b shows the mass spectra of DDPU and DDPS, two extra DDXs obtained after TMAH thermochemolysis.

232

3.2. Distribution of non-extractable DDXs in the three different types of solid samples

233

The degradation pathway of DDT in subaquatic sediments, groundwater and subsoil can be considered

234

as anaerobic. Based on previous studies15,23,42,43 and with some minor modifications, the anaerobic

235

transformation pathway of DDT (although mass balance data for the whole process of direct metabolism

236

are not available) is described in Fig. 4. For a better understanding of the effects of environmental factors

237

on the behavior and fate of non-extractable DDXs, and according to some previous studies23,43, the

238

sequential chemical degradation released NER-DDXs were re-assorted into three groups which are also

239

illustrated in Fig. 4. First, DDT is dehalogenated and the number of chlorine substituents are reduced (step

240

1). Then, a functional group is formed on the dechlorinated carbon atom (step 2). Finally, the newly formed

241

functional group is removed together with the related carbon atom (step 3). For the topsoil BFW-SOIL1, an

242

aerobic and an anaerobic degradation pathway have to be considered. According to the classic aerobic DDT

243

degradation pathway, DDE, DDMU and DBP should be the main metabolites. Research on DDXs in several

244

surface soils44 proposed that DDT is degraded to DDD, DDE, DDMU, DDNU and DBP in surface soils

245

during both, aerobic and anaerobic processes. Hence, the three-group classification can also be applied to

246

BFW-SOIL1.

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The total concentrations of step 1, step 2 and step 3 NER-DDXs in the different solid sample types are

248

also displayed in Fig. 4. These results showed significant distinctions of the NER-DDXs distribution trends

249

between the different types of solid substances. An increase and a decrease of NER-DDXs content along

250

with the degradation process were observed in sediments and soils, respectively (Fig. 4b and d). In the

251

groundwater sludge, an extremely low amount of step 2 NER-DDXs were found, with a clearly higher level

252

of step-1 and step-3 NER-DDXs. The possible reason for this discrimination could be the incorporating

253

mechanisms.

254 255

Fig. 4. Comparison of the contents of non-extractable DDT-metabolites from three steps of natural degradation (a) in different types

256

of solid sample (b sediments, c groundwater sludge, d soils).

257

Natural solid particles consist mostly of minerals and organic matter, and the organic matter covers the

258

mineral surfaces or are present as organo-mineral complexes. The incorporation of DDXs into natural solid

259

particles as non-extractable residues can be i) strong adsorption on solid surfaces as well as on the structural

260

pores (e.g., van-der Waals forces and ionic interactions), ii) physical entrapment in the solid matrix

261

structures, or iii) bondage to the organic macromolecules via covalent linkages with functional groups (e.g.

262

carboxylic, hydroxyl, amino and carbonyl groups)45,46. These incorporation mechanisms and potential 13



263

releasing pathways of covalently bound DDXs as well as the detection of corresponding products are

264

illustrated in Fig. 5. The formation of NER-DDA with sediment organic matter has already been proved to

265

be via ester bonds29. As mentioned above, DDMUBr was likely released when an ether bond between

266

DDNU and the organic matrix was broken by BBr3, DBP might have been formed after the attack of carbon-

267

carbon double bonds by RuO4, and DDPU and DDPS were suggested to be released after carbon-carbon

268

bond cleavage by TMAH thermochemolysis28. The structures of these likely covalently bound precursors

269

(illustrated in Fig. 5) were all in the step 2-DDX-scope. DDMUBr, DDPU and DDPS were detected only in

270

TC-S2 and TC-S3, and could not be quantified. The amounts of DDA released after alkaline hydrolysis in

271

both, sediments and soils, were much more than those released after BBr3 treatment (Fig. 5). In contrast, in

272

groundwater sludge samples, DDA was not detected after hydrolysis but after BBr3 treatment. This released

273

portion of NER-DDA was therefore likely non-covalently bound. In this study, the total DDXs level was

274

significantly higher in sediments than in groundwater sludge, whereas the distribution of DBP after RuO4

275

oxidation was relatively equal (Fig. 5). A potential explanation of this is that DBP might preferably tend to

276

accumulate in solid particles of aquatic mobile phases because of the aqueous-affinity of its keto group. A

277

rough quantitative estimation of covalently and non-covalently bound DDXs was made according to the

278

aforementioned hypothesis (the concentrations of alkaline hydrolysis released DDA and DDOH and RuO4

279

oxidation released DBP were summed up as the covalently bound compounds). As a result, 14 to 70 %, 1.7

280

to 26 %, 25 % and 40 % of NER- DDXs were calculated as covalently bound in sediments, groundwater

281

sludge, BFW-SOIL1 and BFW-SOIL2, respectively.

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Fig. 5. Possible forming mechanisms of NER-DDXs and the potential releasing pathways of some covalently bound DDXs.

284

In general, the DDT metabolites that were detected at the different sampling sites may originate from two

285

sources. The source 1-portion derives from the transport and accumulation of already degraded DDXs from

286

upstream areas. The source 2-portion originates from the in situ degradation of DDXs at this site. In

287

groundwater sludge, the DDT metabolites likely predominantly derived from source 1. The initial

288

metabolites (step 1 NER-DDXs) and the more degraded DDXs (step 3 NER-DDXs) were mostly transported

289

adsorbed on or entrapped in suspended solid particles/colloidal organic matter. Research on the

290

transformation of phenanthrene into NER revealed that microbial activity is essential for the NER formation

291

process with covalent bonds (Wang et al.47). An active microbial community was also believed to promote

292

irreversible sorption of pesticides to soil 4,45,48. This could explain the low proportion of covalently bound 15



293

DDXs in groundwater sludge, where in general lower and divergent microbial effects occur than in surface

294

water49–51, as a result of land filtration and the comparatively lower temperature.

295

The step 2-DDXs, with a relatively higher polarity, have a higher water solubility so that they could have

296

been transported within the aqueous phase over longer distances. During the past DDT production, fine

297

sediment particles with extremely high DDXs loads might have been carried downstream along with the

298

river flow. A considerable portion of those non-strongly bound DDXs then probably partitioned into the

299

aqueous phase and was transported further within this phase. This could explain the low step 1-NER-DDXs

300

content in sediment samples as compared to groundwater sludge and soils, since their incorporation into

301

sediments should be principally non-covalent and thus more reversible. As monitored by Schwarzbauer and

302

Ricking52, DDD, DDE, DDMS, DDNU, DDA, DDOH, DBP were formerly detected in surface water of the

303

Teltow Canal. DDT metabolites in sediments could be from both sources but source 2 might be more

304

dominant corresponding to their high step 3-NER-DDXs level.

305

Besides, from the soil incubation experiments of the herbicide 4-chloro-2-methylphenoxyacetic acid into

306

soil by Riefer et al.53, more significant effects of microbial activity were observed on highly loaded samples.

307

It is reasonable to believe that higher DDXs concentration would promote their incorporation into sediment

308

matrix via covalent bonds. Without frequent rinse of surface water, step 1-DDXs could have been

309

accumulated and have a high residence time in soils. The relatively higher proportion of step 2-DDXs in

310

subsoil than in topsoil indicates the possibility of their vertical transport via soil porous water since they are

311

more hydrophilic than step 1-DDXs. This portion of DDXs might leach to groundwater eventually and the

312

loss of step 2-DDXs would lead to a lower generation of step 3-DDXs in soils. Soil profiles at the same

313

location were formerly studied and DDD was predominantly detected at the depth interval 0 to 6 cm,

314

whereas a lot lower amount of it was observed below 6 cm (even lower than that in downstream sediments)54.

315

In that case, DDD was suggested to be not vertically mobile. To the best of our knowledge, there is no

316

literature so far dealing with the vertical distribution/transport behaviors of the step 2-DDXs in soils.

16


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317

Therefore further efforts have to be done to verify the aforementioned leaching potential of step 2-DDXs in

318

soil profiles.

319

3.3. DDXs remobilization potential on their environmental pathway

320

The temporarily stored or sequestrated DDXs in the solid phase are possibly remobilized under

321

appropriate conditions. As Li et al.55 investigated, the lateral remobilized DDXs from soil contributed to 20

322

to 42 % of the total DDXs fluxes. Consequently, it is quite critical to understand the remobilization potential

323

of NER-DDXs for risk assessments, as it is important for the bioavailability in aquatic and terrestrial systems

324

and the (eco)toxic potential56. The risk of DDXs would be underestimated without considering NER-DDXs.

325

Concentrations of DDA, its precursors as well as of DDCN in EF (defined as EF-risk) plus concentrations

326

of DDA, its metabolites and DDCN in the non-extractable fractions (defined as NER-risk) were calculated

327

as the quantified total DDXs remobilization risk according to Frische, Ricking and Schwarzbauer23,43.

328

Considering the extraordinarily high EF-DDT level in TC-S1, EF-risk (concentrations of DDA, its

329

precursors and DDCN in the extractable fraction) at this location was quantified as the average value of TC-

330

S2 and TC-S3. The results are illustrated in Fig. 6. For sediments, besides the high level at the emission

331

point, the samples downstream possessed a higher total risk and a higher NER-risk proportion than those

332

upstream. This can be attributed to the flushing and carrying activities of surface runoff as mentioned above

333

and the synchronous biotic process. Since there is a bank filtration area through which surface water leaches

334

into groundwater near the groundwater remediation facilities, a high total risk level and low NER-risk

335

proportion in TC-G1 to TC-G4 is reasonable. For soils, a higher total risk level was found in topsoil, whereas

336

a higher NER-risk proportion was observed in subsoil. This points to the existence of a vertical transport of

337

in particular step 2- and step 3- DDXs. Because of the limited number of soil samples that was available for

338

this study this has to be confirmed in future studies. Non-polar DDXs can also be transported vertically via

339

dissolved organic matter in soil57, so that an EF-risk was also observable in subsoil.

17



340 341

Fig. 6. Quantified remobilization risk of DDXs at each sampling site.

342

3.4. Conceptual model of the formation and fate of non-extractable DDXs

343

According to the aforementioned results, their interpretation and discussed implications, a conceptual

344

model of the fate of non-extractable DDXs in environmental aquatic-terrestrial pathway is proposed (see

345

Fig. 7). Even decades after being released, DDT and its metabolites will still remain at the highest level at

346

the emission point as compared to other sites. The special step 2-products DDPU and DDPS will connect

347

strongly and irreversibly to the sedimentary matrix at the emission site with little remobilization potential.

348

The other commonly studied DDXs, however, are subject to remobilization along with the river flow

349

together with the fine sedimentary particles or can be released after the organic matter structures is getting

350

loose. The step 1-DDXs (DDT, DDD, DDE, DDMU, DDMS and DDNU) have a higher remobilization

351

potential than step 2-DDXs (DDA, DDCN, DDEt and DDOH) and step 3-DDXs (DDM, DBP and DBH),

352

because they are predominantly incorporated in the solid matrix by non-covalent bonds. Although DDA and

353

DDCN (both in step 2) are water soluble, their functional groups will make them more likely to be covalently

354

bound. Consequently, sediments from downstream will collect more free-DDXs as well as NER-DDXs than

355

upstream sediments. In general, the distribution of NER-DDXs in the sedimentary phase will be in the order

356

of step 1 > step 2 > step 3, and a certain amount of step 2 ones can be covalently bound owing to the

357

microbial effects and high DDX level. In soil, DDXs can be from both the exposure to surface water and

358

former flood deposition of sediment material. High microbial activity but fairly low DDX concentration 18


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359

lead to a comparatively low covalent binding extent in soil. Our first insights suggest that step 1-DDXs will

360

be entrapped predominantly in topsoil, whereas step 2-DDXs are subject to vertical transport and can even

361

leach into the groundwater. Therefore, solid particles and colloidal organic matter in groundwater will

362

adsorb or entrap the DDXs from both soil and surface water. These DDXs are bound primarily via non-

363

covalent binding owing to relatively lower microbial effects. As a result, step 2-DDXs will be more in free

364

stage or loosely bound because of their higher remobilization potential and solubility. This model can

365

provide some basic knowledge for risk assessment and further remediation of DDT contaminations in soils

366

and aquatic systems under consideration of both – the extractable and the non-extractable portion of the

367

contamination. However, due to the limited number of samples considered in the present study, the aging

368

of solid substances and the variability of the distribution of DDXs at different locations, this proposed

369

conceptual model has to be regarded as a preliminary model, which has to be confirmed by future in-depth

370

studies.

19



371 372

Fig. 7. Conceptual model of the fate of NER-DDXs along the environmental aquatic-terrestrial pathway.

373

ACKNOWLEDGMENTS

374

We gratefully acknowledge the financial support of the China Scholarships Council program (No.

375

201606400065). We also give our gratitude to Mrs. Annette Schneiderwind and Mrs. Yvonne Esser for their

376

help during the laboratory work.

377

Supporting Information

378

Detailed information on sampling areas; foc measurement; sequential chemical degradation and

379

fractionation; DDXs calibration and quantification; and raw data of the concentration of DDXs after each

380

extraction/chemical degradation process (PDF). 20


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