“Behaviors and Fate of Nonextractable Anthropogenic Organohalogens in Aquatic and Terrestrial Environments: Formation Mechanisms and Remobilization Potential”
Von der Fakultät für Georessourcen und Materialtechnik der Rheinisch-Westfälischen Technischen Hochschule Aachen
zur Erlangung des akademischen Grades eines Doktors der Naturwissenschaften
genehmigte Dissertation vorgelegt von
Xiaojing Zhu, M. Eng. aus Hangzhou, China
Berichter: Herr Prof. Dr. rer. nat. Jan Schwarzbauer Herr Univ.-Prof. Dr. rer. nat. Andreas Schäffer
Tag der mündlichen Prüfung: 23.07.2020
Diese Dissertation ist auf den Internetseiten der Universitätsbibliothek online verfügbar
In memory of the lost souls in this pandemic
Acknowledgments
The work of this thesis was carried out at the Institute of Geology and Geochemistry of Petroleum and Coal (LEK), RWTH Aachen University, from October 2016 to April 2020, with financial support from the China Scholarship Council (No. 201606400065). Results of the present work were submitted to and published in some peer-reviewed journals.
With the help and support from a lot of people, this thesis was finally finished. First, I would like to sincerely thank my supervisor Prof. Dr. Jan Schwarzbauer for providing me the opportunity to conduct this research along with his patient instructions, extensive knowledge sharing, inspiring discussions and kind review of my manuscripts. I also want to express my gratefulness to Dr. Larissa Dsikowitzky for her thoughtful escort during my lab work and manuscript writing. Furthermore, I am gratitude to our head of the LEK, Prof. Dr. Ralf Littke, for organizing the institute in such a professional way and offering us many enjoyable activities. My sample treatments and instrumental measurements were kindly supported by Mrs. Annette Schneiderwind and Mrs. Yvonne Esser. I also would like to acknowledge Prof. Dr. Andreas Schäffer from Institute for Environmental Research, RWTH Aachen University, for co-reviewing my thesis.
Thanks to Mrs. Olga Konechnaya, Dr. Qiulei Guo, Dr. Ting Zhang, Dr. Ping Li, Dr. Ronghui Fang and all the rest of my colleagues, the experience of working in this institute became so lovely.
A special gratitude should be express to my partner Jialin for his long-lasting encouragement and understanding.
Finally, I want to thank my parents and all my friends for continuously caring about me and bringing me strength.
i Abstract
Anthropogenic organohalogens (AOHs) are toxic and persistent pollutants that occur ubiquitously in the environment. An unneglectable portion of them can convert into nonextractable residues (NER) in the natural solid substances. NER-AOHs are not detectable by conventional solvent- extraction, and will get remobilized through changes of surrounding environment. Consequently, the formation and fate of NER-AOHs should be investigated comprehensively. In the present study, both field-collected and lab-incubated samples were investigated with respect to the formation mechanisms and fate of nonextractable anthropogenic organochlorine/bromine and the newly concerned contaminants per-/polyfluoroalkyl substances (PFASs), respectively. The field samples including terrestrial and aquatic ones were applied with solvent extraction, sequential chemical degradation and tetramethylammonium hydroxide thermochemolysis to obtain both extractable and nonextractable AOHs. Over 150 AOHs were detected including DDXs (DDT (bis(chlorophenyl)trichloroethane) and its metabolites/ derivatives). DDT and its first degradation products, DDD (bis(chlorophenyl)dichloroethane) and DDE (bis(chlorophenyl)dichloroethylene), were dominant in the free extractable fraction, whereas DDM (bis(chlorophenyl)methane), DBP (bis(chlorophenyl)methanone) and DDA (bis(chlorophenyl)acetic acid) were observed primarily after chemical degradation. The detection of DDA, DDMUBr (bis(chlorophenyl)bromoethylene), DDPU (bis(chlorophenyl)propene) and DDPS (bis(chlorophenyl)propane) after chemical treatments evidenced the covalent bindings between these DDXs and the organic matrix. The identified NER-DDXs were categorized into three groups according to the three-step degradation process of DDT. Their distribution along the different pathways demonstrated significant specificity. The rest detected AOHs could be either originated from direct industrial emissions or from degradation of their precursors, and therefore were discussed with regard to their molecular structures and preferred NER incorporating mechanisms. Covalent linkages were observed most favorable for the hydrophilic-group-containing monocyclic aromatic AOHs (HiMcAr-AOHs) (e.g. halogenated phenols, benzoic acids and anilines) incorporating into the natural organic matter (NOM) as NER. Physical entrapment mainly contributed to the NER formation of hydrophobic monocyclic aromatic AOHs (HoMcAr-AOHs) and polycyclic aromatic AOHs (PcAr-AOHs). The hypothesized remobilization potential of these NER-AOHs follow the order HiMcAr-AOHs > HoMcAr-AOHs/ aliphatic AOHs > PcAr-AOHs. In addition, the NOM macromolecular structures of the studied samples were analyzed. The 240-day lab-incubation investigation was conducted by
ii spiking perfluorooctanesulfonate (PFOS) and its alternative 6:2 chlorinated polyfluorinated ether sulfonate (F-53B) to a fresh topsoil to mimic their long-term fate in terrestrial environment. Solvent extraction, alkaline hydrolysis and sequential chemical degradation were applied on periodically sampled soil to obtain extractable, moderately bound and deeply bound PFASs, respectively. The results confirmed the formation of NER of both compounds but with different preferences of incorporating mechanisms. NER-PFOS was proofed to be formed predominantly by covalent binding (via head group) and strong adsorption (via tail group). The formation of NER-F-53B was mainly driven by physical entrapment. Both bound compounds within the incubation period showed three-stage behaviors including an initial period with slight release followed by a (re)incorporating stage and a subsequent remobilizing stage. This work provides some basic knowledge and molecular insights for risk assessment and further remediation of AOH contaminations in both soils and aquatic systems. Particularly, the part of PFAS-related work, to the best of our knowledge, contributes the first insights on the long-term fate of PFASs.
iii Zusammenfassung
Anthropogene Organohalogene (AOHs) sind toxische und persistente Schadstoffe, die in der Umwelt allgegenwärtig sind. Ein nicht vernachlässigbarer Teil von ihnen kann sich in den natürlichen festen Substanzen in nicht extrahierbare Rückstände (NER) umwandeln. NER-AOHs sind durch herkömmliche Lösungsmittelextraktion nicht nachweisbar und werden durch Änderungen der Umgebung remobilisiert. Folglich sollten die Bildung und das Schicksal von NER-AOHs umfassend untersucht werden. In der vorliegenden Studie wurden sowohl vor Ort gesammelte als auch im Labor inkubierte Proben hinsichtlich der Bildungsmechanismen und des Schicksals von nicht extrahierbarem anthropogenem Organochlor / Brom und den neu betroffenen Kontaminanten Per-/ Polyfluoralkylsubstanzen (PFASs) untersucht. Die Feldproben, einschließlich terrestrischer und aquatischer Proben, wurden mit Lösungsmittelextraktion, sequentiellem chemischem Abbau und Tetramethylammoniumhydroxid-Thermochemolyse angewendet, um sowohl extrahierbare als auch nicht extrahierbare AOHs zu erhalten. Über 150 AOHs wurden nachgewiesen, einschließlich DDXs (DDT (Bis(chlorphenyl)trichlorethan) und seine Metaboliten/ Derivate). DDT und seine ersten Abbauprodukte, DDD (Bis(chlorphenyl)dichlorethan) und DDE (Bis(chlorphenyl)dichlorethylen), dominierten in der frei extrahierbaren Fraktion, während DDM (Bis(chlorphenyl)methan), DBP (Bis(chlorphenyl)methanon)) und DDA (bis(chlorphenyl)essigsäure) hauptsächlich nach chemischem Abbau beobachtet wurden. Der Nachweis von DDA, DDMUBr (Bis(chlorphenyl)bromethylen), DDPU (Bis(chlorphenyl)propen) und DDPS (Bis(chlorphenyl)propan) nach chemisch-degradativen Behandlungen zeigte die kovalenten Bindungen zwischen diesen DDXs und der organischen Matrix. Die identifizierten NER-DDXs wurden gemäß dem dreistufigen Abbauprozess von DDT in drei Gruppen eingeteilt. Ihre Verteilung entlang der verschiedenen Wege zeigte eine signifikante Spezifität. Die übrigen nachgewiesenen AOHs könnten entweder aus direkten industriellen Emissionen oder aus dem Abbau ihrer Vorläufer stammen und wurden daher im Hinblick auf ihre molekularen Strukturen und bevorzugten NER-Einbaumechanismen diskutiert. Kovalente Bindungen wurden am günstigsten für die hydrophile Gruppen enthaltenden monocyclischen aromatischen AOHs (HiMcAr-AOHs) (z. B. halogenierte Phenole, Benzoesäuren und Aniline) beobachtet, die als NER in die natürliche organische Substanz (NOM) eingebaut wurden. Der physikalische Einschluss trug hauptsächlich zur NER-Bildung von hydrophoben monocyclischen aromatischen AOHs
iv (HoMcAr-AOHs) und polycyclischen aromatischen AOHs (PcAr-AOHs) bei. Das hypothetische Remobilisierungspotential dieser NER-AOHs folgt der Reihenfolge HiMcAr-AOHs > HoMcAr- AOHs/ aliphatische AOHs > PcAr-AOHs. Zusätzlich wurden die makromolekularen NOM- Strukturen der untersuchten Proben analysiert. Eine 240-tägige Laborinkubationsuntersuchung wurde durchgeführt, indem Perfluoroctansulfonat (PFOS) und sein alternatives 6: 2-chloriertes polyfluoriertes Ethersulfonat (F-53B) auf einen frischen Mutterboden versetzt wurden, um ihr langfristiges Schicksal in terrestrischer Umgebung nachzuahmen. Lösungsmittelextraktion, alkalische Hydrolyse und sequentieller chemischer Abbau wurden auf periodisch entnommenen Böden angewendet, um extrahierbare, mäßig gebundene bzw. tief gebundene PFASs zu erhalten. Die Ergebnisse bestätigten die Bildung von NER beider Verbindungen, jedoch mit unterschiedlichen Präferenzen für den Mechansimus des Einbaus. Es wurde nachgewiesen, dass NER-PFOS vorwiegend durch kovalente Bindung (über die Kopfgruppe) und starke Adsorption (über die Schwanzgruppe) gebildet wird. Die Bildung von NER-F-53B wurde hauptsächlich durch physikalische Einschlüsse angetrieben. Beide gebundenen Verbindungen zeigten innerhalb der Inkubationszeit ein dreiphasiges Verhalten, einschließlich einer Anfangsperiode mit leichter Freisetzung, gefolgt von einer (Wieder-) Einarbeitungsphase und einer anschließenden Remobilisierungsphase. Diese Arbeit liefert grundlegende molekulare Erkenntnisse für die Risikobewertung und die weitere Sanierung von AOH-Kontaminationen sowohl in Böden als auch in aquatischen Systemen. Insbesondere der Teil der PFAS-bezogenen Arbeit liefert nach unserem besten Wissen die ersten Erkenntnisse über das langfristige Schicksal von PFASs.
v
Contents
ACKNOWLEDGMENTS ...... I ABSTRACT ...... II ZUSAMMENFASSUNG ...... IV CONTENTS ...... VI LIST OF FIGURES ...... VIII LIST OF TABLES ...... XI ABBREVIATIONS ...... XII 1. INTRODUCTION ...... 1
1.1 NONEXTRACTABLE RESIDUES (NER) ...... 1 1.2 ATHROPOGENIC ORGANOHALOGENS (AOHS) ...... 3 1.3 METHODOLOGICAL APPROACHES FOR NER-AOHS INVESTIGATION ...... 5 1.4 CURRENT KNOWLEDGE ON NER-AOHS ...... 10 1.5 THESIS OVERVIEW ...... 18 2. FORMATION AND FATE OF POINT-SOURCE NONEXTRACTABLE DDT-RELATED COMPOUNDS ON THEIR ENVIRONMENTAL AQUATIC-TERRESTRIAL PATHWAY ...... 20
2.1 INTRODUCTION ...... 21 2.2 MATERIAL AND METHODS ...... 23 2.2.1 Samples ...... 23 2.2.2 Extraction ...... 24 2.2.3 Chemical degradation ...... 25 2.2.4 GC–MS analysis ...... 26 2.3 RESULTS AND DISCUSSION ...... 27 2.3.1 Occurrence and concentrations of extractable/nonextractable DDXs in subaquatic sediments, soils and groundwater sludge ...... 27 2.3.2 Distribution of nonextractable DDXs in the three different types of solid samples ...... 30 2.3.3 DDXs remobilization potential on their environmental pathway ...... 35 2.3.4 Conceptual model of the formation and fate of nonextractable DDXs ...... 36 2.4 CONCLUSION ...... 38 3. MOLECULAR INSIGHTS INTO THE FORMATION AND REMOBILIZATION POTENTIAL OF NONEXTRACTABLE ANTHROPOGENIC ORGANOHALOGENS IN HETEROGENEOUS ENVIRONMENTAL MATRICES ...... 45
3.1 INTRODUCTION ...... 46 3.2 MATERIALS AND METHODS ...... 49 3.2.1 Samples ...... 49 3.2.2 Extraction and Chemical Degradation ...... 49 3.2.3 GC‒MS Analysis ...... 50 3.3 RESULTS AND DISCUSSION ...... 51 3.3.1 Occurrence of AOHs and Their Binding Forces towards the Matrices ...... 51 3.3.2 NER-AOHs with Different Degree of Chlorination ...... 54 3.3.3 Molecular structures of NOM Moieties ...... 57
vi 3.3.4 Conceptual Model of the Formation of Nonextractable AOHs ...... 59 3.3.5 (Re)mobilization Risk of AOHs on Their Environmental Pathway ...... 60 3.4 CONCLUSION ...... 63 4. FIRST INSIGHTS INTO THE LONG-TERM DYNAMIC BEHAVIORS AND FATE OF PERFLUOROOCTANESULFONATE AND ITS ALTERNATIVE 6:2 CHLORINATED POLYFLUORINATED ETHER SULFONATE IN SOIL AS NONEXTRACTABLE RESIDUES ...... 77
4.1 INTRODUCTION ...... 78 4.2 MATERIALS AND METHODS ...... 80 4.2.1 Materials ...... 80 4.2.2 Incubation ...... 81 4.2.3 Extraction ...... 81 4.2.4 Alkaline Hydrolysis ...... 81 4.2.5 Chemical Degradation ...... 81 4.2.6 Analytical Methods ...... 82 4.3 RESULTS AND DISCUSSION ...... 83 4.3.1 Distribution of extractable and nonextractable PFASs in soil along aging...... 83 4.3.2 Elucidation of further bound PFASs ...... 85 4.3.3 Dynamic incorporating-remobilizing behaviors of PFASs...... 88 4.4 CONCLUSION ...... 91 5. GENERAL CONCLUSION AND OUTLOOK ...... 95 REFERENCES ...... 99 APPENDIX A: THE IDENTIFIED DATA OF AOHS (FOR CHAPTER 3) ...... 121 APPENDIX B: THE QUANTIFIED DATA OF AOHS (FOR CHAPTER 3) ...... 141
vii List of figures
Figure 1-1. A diagrammatic representation of natural solid matter as adapted from Masoom et al., (2016)...... 2
Figure 1-2. The incorporating mechanisms of NER, adapted from Claßen (2019)...... 3 Figure 1-3. DDT and its metabolites along with their degradation pathway, adapted from Frische (2011), Mackintosh et al. (2016), Ricking and Schwarzbauer (2012) and Wetterauer et al. (2012)...... 5 Figure 1-4. The summarized workflows of various reported approaches for investigation of NER- AOHs ...... 6 Figure 1-5. Bonds cleavage mechanisms of Na18OH alkaline hydrolysis, adapted from Kalathoor et al. (2015b)...... 7 Figure 1-6. The work scheme of acquiring NER with separating humic substances and silylation, adapted from Celi et al. (1997), Haider et al. (1992), Liebich et al. (1999), Loiseau and Barriuso (2002) and Weiß et al. (2004). The solid arrows point the mainstream of this approach. The dashed arrows show the extra procedures reported in some literatures...... 9 Figure 1-7. The workflow of alkaline hydrolysis, sequential chemical degradation, pyrolysis/thermochemolysis and their corresponding covalent bond-cleaving mechanisms, adapted from Kalathoor et al. (2015b), Kronimus and Schwarzbauer (2007), Riefer et al. (2017 and 2013), Schwarzbauer et al. (2005) and Schwarzbauer and Jovančićević (2020). Likewise, the dashed arrows show the extra procedures that were only reported in a few literatures...... 10
Figure 1-8. AOHs that were targetedly investigated as NER in the previous studies...... 11 Figure 1-9. Proposed mechanisms of chlorinated phenols, chlorinated benzoic acids, chlorinated methylphenoxyacetic acid and chlorinated anilines incorporating into NOM via chemical bindings as described in previous works, adapted from Hatcher et al. (1993), Riefer et al. (2017), Scheunert and Schröder (1998), Thorn et al. (1996) and Yuan et al. (2017)...... 13 Figure 1-10. Proposed mechanisms of DDXs incorporating into NOM as described in previous works, adapted from Frische (2011) and Kronimus et al. (2006)...... 14 Figure 1-11. Proposed incorporating mechanisms of atrazine, anilazine and their metabolites into NOM as reported in previous works, adapted from Dankwardt et al. (1998), Dankwardt and Hock (2001) and Klaus et al. (1998)...... 15 Figure 1-12. Proposed incorporating mechanisms of TBBA and its metabolites into NOM as reported in previous works, adapted from Li et al. (2015), Liu et al. (2013) and Tong et al. (2016)...... 15 Figure 1-13. Proposed association of organofluorines (aliphatic and aromatic) towards NOM from previous works, adapted from Klaus et al. (1998) and Shirzadi et al. (2008a, 2008b)...... 17
Figure 2-1. Sampling locations at the Teltow Canal and in Bitterfeld-Wolfen. Total organic matter (foc, dry weight) was determined by loss on ignition...... 24
viii Figure 2-2. DDX concentrations in sample material from two highly DDT-contaminated sites in Germany. All concentrations are given in nmol g-1 dry weight. Figure 2-2a) shows the freely extractable pollutants and Figures 2-2b), c) and d) the pollutants released after different chemical degradation steps in sequence. The dominant DDXs in each step are marked by yellow bars and the compounds with considerable amount observed exclusively in groundwater sludge are marked by green dash lines...... 29 Figure 2-3. Quantified results of DDXs released by TMAH thermochemolysis (a). The dominant DDXs are marked by yellow bars. Figure 2-3b shows the mass spectra of DDPU and DDPS, two extra DDXs obtained after TMAH thermochemolysis...... 30 Figure 2-4. Comparison of the contents of nonextractable DDT-metabolites from three steps of natural degradation (a) in different types of solid sample (b sediments, c groundwater sludge, d soils)...... 31 Figure 2-5. Possible forming mechanisms of NER-DDXs and the potential releasing pathways of some covalently bound DDXs...... 33
Figure 2-6. Quantified remobilization risk of DDXs at each sampling site...... 36 Figure 2-7. Conceptual model of the fate of NER-DDXs along the environmental aquatic-terrestrial pathway...... 37
Figure S2-1. Procedure of sample treatment and fractionation of each extract. The volumn of solvent for each elution was 5 mL except that F2 from first extraction was obtained by 8.5 mL of solvent mixture...... 44
Figure 3-1. The categorization of the detected AOHs according to their putative binding mechanisms. Their speculative structures corresponding to certain covalent bonds are orange color-marked. The presumed original molecular structures before covalent bond-cleavages are depicted in blue color...... 53 Figure 3-2. Quantified AOH homologous in different solid matrices. Trends of partitions between extractable and nonextractable fractions varied gradually along with the increase of the grades of chlorination of their NER. Corresponding predicted soil adsorption coefficient (Kd, presented as log Kd) data were taken from the chemistry dashboard of US EPA (https://comptox.epa.gov/dashboard). The speculative structures of AOHs corresponding to certain covalent bonds are color-marked. The degradation pathway of HCHs (hexachlorocyclohexanes) was adapted from Ricking and Schwarzbauer (2008)(Ricking and Schwarzbauer, 2008)...... 56
Figure 3-3. The Br-substituted NOM moieties released after BBr3 treatment and/or RuO4 oxidation. R and R’ refer to the NOM moieties that can be released by BBr3 treatment and RuO4 oxidation, respectively...... 58
Figure 3-4. A proposed model of the formation mechanisms of NER-AOHs into NOM...... 60 Figure 3-5. AOHs with gradual changes of distribution patterns on their environmental pathway (the concentration of each group of compounds was calculated as the sum). The speculative structures of AOHs corresponding to certain covalent bonds are color-marked...... 62
ix
Figure S3-1. Sampling locations of the studied samples and their foc (total organic content) were determined by loss on ignition...... 64
Figure S3-2. Procedure of sample treatment and fractionation of each extract...... 64
Figure S3-3. AOHs detected in EF and/or as NER...... 65
Figure 4-1. The concentrations of PFOS (panel a) and F-53B (panel b) in soil (dw) as EF and NER (released by alkaline hydrolysis) after each incubation period...... 84 Figure 4-2. Ion chromatograms (m/z 131, m/z 181 and m/z 231) of the detected PFCs after chemical degradation were applied on the standard compounds (panel a). A representative mass spectrum (retention time 17.8 min) of the released PFCs (panel b)...... 86 Figure 4-3. Ion chromatograms (m/z 131, m/z 181 and m/z 231) of the detected PFCs after chemical degradation were applied on the hydrolyzed soil residues (panel a). A representative mass spectrum (retention time 17.8 min) of the released PFCs (panel b)...... 87 Figure 4-4. Quantified results of the obtained PFASs (panel a) and nonpolar PFCs (panel b) released from the alkaline hydrolyzed soil residues (day-30, day-60 and day-120) by BBr3 treatment and RuO4 oxidation...... 88 Figure 4-5. The dynamic incorporating-remobilizing indexes of PFASs with various possible mechanisms along the increased incubation period. The data of the total bound PFASs, moderately bound PFASs and deeply bound PFASs were given in panel a, b and c, respectively...... 90 Figure 4-6. Three-stage behaviors of bound PFOS (panel a) and F-53B (panel b) with different binding mechanisms...... 91
Figure S4-1. Procedures of each sample treatment and fractionation...... 93 Figure S4-2. The molecular structure and the mass spectrum of perfluoro-1,3-dimethylcyclohexane...... 94
x List of tables
Table S2-1 The quantified results of DDXs after ASE extraction...... 39
Table S2-2 The quantified results of DDXs after alkaline hydrolysis...... 40
Table S2- 3 The quantified results of DDXs after BBr3 treatment...... 41
Table S2-4 The quantified results of DDXs after RuO4 oxidation...... 42
Table S2-5 The quantified results of DDXs after TMAH thermochemolysis...... 43
Table S3- 1 The characteristic ions for quantification ...... 66
Table S3-2 Predicted soil adsorption coefficient (Kd, data are from the chemistry dashboard of US EPA (https://comptox.epa.gov/dashboard)) and possible origins of the studied AOHs (EPA and additional references) ...... 73
xi
Abbreviations
AOHs Anthropogenic organohalogens
BBr3 Boron tribromide
C8H16 Perfluoro-1,3-dimethylcyclohexane
CCl4 Carbon tetrachloride DBH bis(Chlorophenyl)methanol DBP bis(Chlorophenyl)methanone DCM Dichloromethane DDA bis(Chlorophenyl)acetic acid DDCN bis(Chlorophenyl)acetonitrile DDD bis(Chlorophenyl)dichloroethane DDE bis(Chlorophenyl)dichloroethylene DDEt bis(Chlorophenyl)ethane DDM bis(Chlorophenyl)methane DDMS bis(Chlorophenyl)chloroethane DDMU bis(Chlorophenyl)chloroethylene DDMUBr bis(Chlorophenyl)bromoethylene DDNU bis(Chlorophenyl)ethylene DDOH bis(Chlorophenyl)ethanol DDPS bis(Chlorophenyl)propane DDPU bis(Chlorophenyl)propene DDT bis(Chlorophenyl)trichloroethane DDXs DDT and its metabolites dw Dry weight EF Extractable fraction F-53B 6:2 Chlorinated polyfluorinated ether sulfonate FA Fulvic acid
xii foc Total organic matter GC–MS Gas chromatography–mass spectrometry HA Humic acid HCHs Hexachlorocyclohexanes HiMcAr-AOHs Hydrophilic-group-containing monocyclic aromatic AOHs HoMcAr-AOHs Hydrophobic monocyclic aromatic AOHs HPLC‒MS High-performance liquid chromatography‒mass spectrometry
Kd Predicted soil adsorption coefficients (i.e. partition coefficient) KOH Potassium hydroxide LSC Liquid scintillation counting NER Nonextractable residues NMR Nuclear magnetic resonance NOM Natural organic matter PAHs Polycyclic aromatic hydrocarbons PcAr-AOHs Polycyclic aromatic AOHs PCBs Polychlorinated biphenyls PFASs Per- and polyfluoroalkyl substances PFCs Per- and polyfluorinated chemicals PFOA Perfluorooctanoic acid PFOS Perfluorooctanesulfonate
RuO4 Ruthenium tetroxide SOM Soil organic matter SPE Supercritical fluid extraction TBBA Tetrabromobisphenol A TLC-RAD Thin Layer Chromatography with radioactivity detection TMAH Tetramethylammonium hydroxide
xiii
1. Introduction
1.1 Nonextractable residues (NER)
Anthropogenic contaminants can be released from industrial emission, agricultural application and municipal usage, and will transport into various compartments including water systems, atmosphere, natural solid substances and (micro)organism. They can further incorporated into the natural solid matter as nonextractable residues (NER) which are temporally immobilized (Kästner et al., 2014). NER, also known as bound residues, were first recognized in 1960s (Bailey and White, 1964; Chiba and Morley, 1968) as a portion of organic substances temporarily or permanently not free available because of their strong association with the environmental solid phase (e.g. soil, sediment and aquatic suspended particles). The specific definition of NER is longstanding xenobiotics or their metabolites in solid matrix that cannot be released after mild solvent extraction which do not change the physicochemical nature of both the compounds and the adhered solid matrix (Barriuso et al., 2008; Führ et al., 1998; Loos et al., 2012).
In aquatic system, suspended particle and colloidal organic matter in water (surface water or groundwater) and sediments are the major targets for the formation of NER. Soil is the most important solid sink for NER accumulation and transformation in terrestrial system. Humic substances, as an constituent portion of natural organic matter (NOM) particularly in the form of humic-clay complexes, always play an important role as the preferential binding domain (Bollag and Loll, 1983; Heim and Schwarzbauer, 2013; Kästner, 2000; Kästner et al., 2014; Liu et al., 2013; Riefer et al., 2013; Schäffer et al., 2009). Besides, with regard to the macromolecular structures of NOM, carbohydrate, lipids (aliphatic), lignin and microbes are proposed to be the main components. A schematic composition of natural solid matter is given in Figure 1-1 according to the work from Masoom et al. (2016).
1
Figure 1-1. A scheme on proposed structural moieties in natural solid matter as adapted from Masoom et al., (2016).
Thus, these organic contaminants can incorporate into the solid matrix via various physical and chemical interactions. These interactions are controlled mainly by the molecular physicochemical properties of the xenobiotics and the macromolecular variations of the NOM (Fenlon et al., 2011). It has already been concluded by intensive research that the forming mechanisms of NER comprise strong adsorption (including Van der Waals, π binding, hydrophobic binding, ionic binding and hydrogen bonds) (Gevao et al., 2000; Jianyang Guo et al., 2017), physical entrapment (sequestration in the solid matter structures), covalent binding (bondage to the organic macromolecules via covalent linkages with functional groups such as carboxylic, hydroxyl, amino and carbonyl groups) (Schwarzbauer and Jovančićević, 2018; Suddaby et al., 2014) and biogenic turnover after degradation of the compounds. Biogenic processes incorporate xenobiotics (which are mainly easily biodegradable) into living biomass and particularly into proteins, and such
2 biogenic NER will be released by acid hydrolysis for destroying proteins (Schäffer et al., 2018). Biogenic NER that were suggested to be not of environmental concern (Claßen et al., 2019; Kästner et al., 2016; Nowak et al., 2018; Poßberg et al., 2016; Schäffer et al., 2018; Trapp et al., 2018; S. Wang et al., 2016), and ,therefore, were not discussed in the present study. The concerned incorporating mechanisms of xenobiotics into solid matrix as NER are illustrated in Figure 1-2.
Figure 1-2. The incorporating mechanisms of NER, adapted from Claßen (2019).
Nonextractable contaminants and their metabolites are just temporarily not detectable with normal analytical methods, but will get partially bioavailable or remobilized later with physical or chemical changes of surrounding conditions, such as pH variation (e.g. low-weight organic acid may enhance the release (Gao et al., 2015)), modified redox conditions, precipitation, microbial activities and even aging (Alexanderova, 2016; Geng et al., 2015; Hamid et al., 2018; J. Li et al., 2019; Sabatier et al., 2014; Schäffer et al., 2018). According to the research of Umeh et al. (2018) about time- dependent remobilization of nonextractable benzo[a]pyrene in soil, slow intercompartmental partitioning of more sequestrated into less sequestrated NER would lead to its remobilization. Thus, the NER portion of anthropogenic pollutants should not be neglected for their environmental risk assessment, otherwise, the contamination level will possibly be underestimated. Therefore, knowledge on the formation mechanisms, behaviors and fate (for instance the potential of remobilization) of NER in aquatic and terrestrial environments, if possible, in a long-term scale, are critically needed.
1.2 Anthropogenic organohalogens (AOHs)
Once released into the environment. apart from being degraded into unhazardous products and/or mineralized, a prominent portion of AOHs are somehow recalcitrant (such as pesticides, herbicides, flame retardants and chemical additives), or their metabolites will still be pernicious to organisms. Among all the reported AOHs, DDT (bis(chlorophenyl)trichloroethane) and its metabolites (DDXs, their degradation pathway was illustrated in Figure 1-3), PCDD/Fs (polychlorinated dibenzo-p- dioxins and dibenzofurans), PCBs (polychlorinated biphenyls), HCHs (hexachlorocyclohexanes),
3 TBBA (Tetrabromobisphenol A) and atrazine are the most concerned compounds owing to their global detections and persistent nature. Recently, poly- and perfluorinated chemicals (PFCs), particularly the poly- and perfluoroalkyl substances (PFASs, for instance, perfluorooctanoic acid (PFOA) and perfluorooctanesulfonate (PFOS)), are receiving increasing attentions. Alternatives are continuously being synthesized and utilized after the phase out of the well-established PFCs. Consequently, more than a thousand of PFCs were identified from different environmental matrices and biological tissues so far. This series of widely used/emitted but newly environmentally recognized chemicals need to be investigated thoroughly.
4 Figure 1-3. DDT and its metabolites along with their degradation pathway, adapted from Frische (2011), Mackintosh et al. (2016), Ricking and Schwarzbauer (2012) and Wetterauer et al. (2012).
In addition to the uptake (but not digested) by plants, animals and other (micro)organisms, these chemicals can form NER in environmental solid matters, for instance soil, sediments and airborne particles. As elaborated in section 1.1, NER-AOHs could be a 'time bomb' that will pose potential risks to the environment and human health, and therefore should be taken into account when performing risk assessments. According to previous works, PCBs, DDT, TBBA, atrazine, chlorinated benzenes, chlorinated anilines and chlorinated phenols as well as their degradation products were reported to form NER (Bhandari et al., 1997; Dankwardt et al., 1998; Heim et al., 1995, 1994; Kucher and Schwarzbauer, 2017; Liu et al., 2013; Loiseau and Barriuso, 2002; Luo et al., 2010; McAvoy et al., 2016; Park et al., 2004; Puglisi et al., 2007; Schroll et al., 2003; Schwarzbauer et al., 2003a; Tong et al., 2016; Wen et al., 2009; Yuan et al., 2017). However, although the issue of NER has been extensively studied during the last several decades, knowledge on NER-AOHs are still not sufficient because a large part of the efforts was done on the nonhalogenic compounds such as polycyclic aromatic hydrocarbons (PAHs) (He et al., 2008; Northcott and Jones, 2012; Umeh et al., 2018; Yang et al., 2010). Therefore, more work is needed to investigate the formation and fate of NER-AOHs.
1.3 Methodological approaches for NER-AOHs investigation
For elucidating the nature of bound xenobiotics within solid substances (e.g. soils and sediments), many approaches have been reported, including various sample treatments for obtaining NER and several analytical methods to determine their incorporating behaviors and mechanisms. A summary of the workflows for NER-AOHs investigation from the previous works during the last three decades is illustrated in Figure 1-4. Note that although there are more reported methods such as supercritical fluid extraction (SFE), which was once applied for capturing NER (Schäffer et al., 2018), they were only partly appropriate for discriminating both extractable fraction (EF) (too strong that will change the properties of a sample) and NER (not exhaustive enough particularly for covalent bonds cleavage), and therefore are not discussed in this thesis. As shown in Figure 1- 4, previous work was conducted mainly on lab-simulating samples and field-collected precontaminated samples. Soils, sediments, organo-clay complexes and humic substances were involved.
5
Figure 1-4. The summarized workflows of various reported approaches for investigation of NER-AOHs
Studies on lab-simulating samples could be subdivided into two categories as incubation and sorption-desorption experiments. Radiolabeling, as a reliable technique which is proved valid for not only quantifying the bound compounds but also investigating the binding groups/moieties between the compounds and matrix, was applied in a prevailing amount of the works with spiking
6 mainly of 14C labeled compounds. However, also 13C, 15N, 18O and 37Cl labeling were reported to be suitable (Benoit and Preston, 2000; Bollag and Dec, 2001; Hatcher et al., 1993; Kalathoor et al., 2015b; Nowak et al., 2013; Osswald et al., 2016; Riefer et al., 2013; Thorn et al., 1996; S. Wang et al., 2016). LSC (liquid scintillation counting), TLC-RAD (thin layer chromatography with radioactivity detection), radio-HPLC (high-performance liquid chromatography), GC-MS (gas chromatography–mass spectrometry), LC-MS and NMR (nuclear magnetic resonance) spectroscopy together with residue combustion were implemented to measure the bound compounds (Gevao et al., 2005; Liu et al., 2013; McAvoy et al., 2016; Mordaunt et al., 2005; Suddaby et al., 2016).
However, these methods are only feasible for lab-scaled work but cannot be applied on field samples. Therefore, theories that were drawn from those radio-mimicking experiments are more on a theoretical level and need to be proofed by field sample measurements. Moreover, lab-scaled incubation and sorption-desorption experiments could just provide information on very limited (types of) AOHs which are much less than those under natural conditions. In some highly contaminated sites, different pollutants could interact with each other towards their incorporating into solid matrices by co-association, competing the binding sites or elimination. So far, only alkaline hydrolysis with Na18OH has been reported to be a valid protocol to release some NER from field samples with 18O as the indicator of the yielded carboxyl/hydroxyl groups (Kalathoor et al., 2015b; Michaelis et al., 1995). The corresponding mechanisms were illustrated in Figure 1-5. Yet this alkaline treatment could only release NER bound to NOM via hydrolysable carboxyl and/or amide bonds, or by decomposing NOM macromolecules where NER were entrapped. Therefore, further chemical treatments are still needed. This issue is further discussed below.
Figure 1-5. Bonds cleavage mechanisms of Na18OH alkaline hydrolysis, adapted from Kalathoor et al. (2015b).
Some incubations were also conducted without radiolabeling. Apart from the normal strategies, in- situ elucidation of bound AOHs at molecular level by means of some high resolution NMR (for instance cross-polarization magic angle spinning NMR, high-resolution magic angle spinning
7 NMR, saturation-transfer double difference NMR) (Longstaffe et al., 2012, 2010; Shirzadi et al., 2008a, 2008b) and immunolocalization (with specific antibodies) (Dankwardt et al., 1998, 1996; Dankwardt and Hock, 2001) were respectively reported.
For both field-collected and incubated samples, solvent extraction and chemical treatments were carried out to separate EF and NER in the previous works. Uniformly, solvent extraction (either organic solvent(s) or a mixed solution) was first applied to obtain EF. Subsequently, the residue was processed for NER-releasing by means of chemical treatment(s). One popular NER-releasing method is to separate humic substances from the matrix samples (with different acid and alkaline treatment) followed by a silylation (normally with chlorotrimethylsilane) to derivatize the products (Celi et al., 1997; Haider et al., 1992; Klaus et al., 1998; Liebich et al., 1999; Loiseau and Barriuso, 2002; Tong et al., 2016; Weiß et al., 2004; Zhao et al., 2016). The work scheme of this method was illustrated in Figure 1-6. However, despite of how much NER could be detected after silyation, the separated humic substances are just parts of the soil organic matter (SOM). And the terms of humic acid (HA), fulvic acid (FA) and humin are of technical nature and still without a clear definition with respect to their molecular structural features. Moreover, according to Claßen et al. (2019) and Schäffer et al. (2018), xenobiotics released after silylation could be assorted as physical entrapped and strongly adsorbed NER (defined as type I-NER by them), whereas the remaining compounds within the solid matrix should be the covalently bound NER which could not be accessed by this methods.
8
Figure 1-6. The work scheme of acquiring NER with separating humic substances and silylation, adapted from Celi et al. (1997), Haider et al. (1992), Liebich et al. (1999), Loiseau and Barriuso (2002) and Weiß et al. (2004). The solid arrows point the mainstream of this approach. The dashed arrows show the extra procedures reported in some literatures.
Hence, another frequently reported method which is consist of alkaline hydrolysis, sequential chemical degradation (BBr3 treatment and RuO4 oxidation) and pyrolysis/thermochemolysis (mostly with tetramethylammonium hydroxide, to release further NER) seems to be more suitable for understanding NER-AOHs comprehensively by cleaving various chemical bonds corresponding to each step of treatments (Kalathoor et al., 2015b; Kronimus and Schwarzbauer, 2007; Riefer et al., 2017, 2013; Schwarzbauer et al., 2005; Schwarzbauer and Jovančićević, 2020). The workflow of this method as well as the corresponding bond-cleaving mechanisms were depicted in Figure 1-7. This method is promising for investigating the formation mechanisms and fate of NER-AOHs particularly within field samples.
9
Figure 1-7. The workflow of alkaline hydrolysis, sequential chemical degradation, pyrolysis/thermochemolysis and their corresponding covalent bond-cleaving mechanisms, adapted from Kalathoor et al. (2015b), Kronimus and Schwarzbauer (2007), Riefer et al. (2017 and 2013), Schwarzbauer et al. (2005) and Schwarzbauer and Jovančićević (2020). Likewise, the dashed arrows show the extra procedures that were only reported in a few literatures.
1.4 Current knowledge on NER-AOHs
Researches with regard to NER-AOHs formation process have been carried out during the last three decades by applications of various protocols as elaborated in section 1.3. Organochlorines, organobromines and organofluorines with different physicochemical properties were assessed, including both aliphatic and aromatic (monocyclic and polycyclic) ones. A summary of the targeted AOHs as reported by previous works is illustrated in Figure 1-8, accompanied by their molecular structures. For most of the listed compounds, NER (or also called as (irreversibly) bound/ sequestrated/ nondesorbable residues (Bhandari et al., 1997; Liu et al., 2013; Park et al., 2004; Suddaby et al., 2016; Wen et al., 2009; Zhao et al., 2016)) were observed. Works were conducted with respect to the NER conversing proportion, their incorporating mechanisms, binding preferences and further bioavailability (remobilization).
10
Figure 1-8. AOHs that were investigated as NER in previous studies.
11 According to the previous studies, the incorporating mechanisms of NER-AOHs were reported to be via both covalent and noncovalent bindings as well as via entrapment within the particle pores and in between mineral layers. The proposed binding mechanisms of chlorinated monocyclic aromatic compounds, DDXs, atrazine, anilazine, TBBA and organofluorines were as representatives and described in detail below, since these compounds have been quite explored previously and some valuable theories have already been drawn.
The formation of nonextractable chlorinated phenols, chlorinated benzoic acids, chlorinated methylphenoxyacetic acid and chlorinated anilines were concluded to be via covalent bonds (ester, amide, ether and carbon-carbon bonds) and by noncovalent sequestration as depicted in Figure 1- 9. Alkaline hydrolysis with Na18OH which obtained both 18O labeled and unlabeled dichlorobenzoic acids evidenced both their ester bindings and physical entrapment/adsorption (Michaelis et al., 1995). Similar phenomenon was found by Riefer et al. (2017) with 14C labeling of 4-chloro-2-methylphenoxyacetic acid. Ester/amide bonds followed by ether linkages were reported as the main incorporating modes of it into organo-clay complexes. Enzymatic coupling was also proofed to be valid for carbon-carbon (with a substitution on either the Cl site or the H site of the benzene ring) and carbon-oxygen (ether and ester linkages via the hydroxyl group of the compound, with a substitution on the H site of the benzene ring, or with hydrolysis of Cl) bindings between chlorinated phenols and humic acids by means of 13C labeling (Hatcher et al., 1993). Besides, nucleophilic addition reactions were shown to be involved in the chemical binding for anilines to humic substances according to the work of Thorn et al. (1996) with 15N NMR study. The reactions should be triggered between the amino groups of the compounds and the quinones (and other carbonyl groups) within the humic macromolecules. The resulted aniline-humic moieties could be anilinohydroquinone, anilinoquinone, anilide, imine and heterocyclic nitrogen. This theory can be applied to chlorinated anilines. Particularly, Scheunert and Schröder (1998) evidenced it with an incubation experiment with 14C labeled monochloroaniline spiked in an agricultural soil. Moreover, metabolites of 3,4-dichloroaniline, as 3,4-dichloroacetanilide and 3,3’,4,4’-tetrachloroazobenzene, were found to be only in EF by Yuan et al. (2017).
12
Figure 1-9. Proposed mechanisms of chlorinated phenols, chlorinated benzoic acids, chlorinated methylphenoxyacetic acid and chlorinated anilines incorporating into NOM via chemical bindings as described in previous works, adapted from Hatcher et al. (1993), Riefer et al. (2017), Scheunert and Schröder (1998), Thorn et al. (1996) and Yuan et al. (2017).
DDT, with a very long history since the 1940s, is still widely detected at present together with its metabolites. Although NER-DDXs have been reported in many works, knowledge on their incorporating mechanisms are limited. As results from works of Kucher and Schwarzbauer (2017) and Schwarzbauer et al. (2003b), DDT as well as its main degradation product DDD
13 (bis(chlorophenyl)dichloroethane) and DDE (bis(chlorophenyl)dichloroethylene) were more abundant in EF, whereas its further degradation metabolites DDA (bis(chlorophenyl)acetic acid), DBP(bis(chlorophenyl)methanone), DDNU (bis(chlorophenyl)ethylene) and DDM (bis(chlorophenyl)methane) were predominated as NER. DDA released by alkaline hydrolysis (particularly with Na18OH) evidenced the ester bond formation between DDA and NOM macromolecules (Frische, 2011; Kalathoor et al., 2015b). A further carbon-carbon linkage between DDEt (bis(chlorophenyl)ethane) and NOM was discovered after application of TMAH (tetramethylammonium hydroxide) thermochemolysis with released products as DDPS (bis(chlorophenyl)propane) and DDPU (bis(chlorophenyl)propene) (Kronimus et al., 2006). The aforementioned incorporation processes were summarized in Figure 1-10.
Figure 1-10. Proposed mechanisms of DDXs incorporating into NOM as described in previous works, adapted from Frische (2011) and Kronimus et al. (2006).
The NER formation of atrazine and its metabolites (or its congener simazine) was suggested to be due to both physical entrapment and chemical binding after silylation experiments (Hartlieb et al., 2003). Thio, amino and other heteroatomic bonds (after the substitutions of Cl within atrazine molecules) between atrazine and NOM were found to be the main chemical binding (Dankwardt et al., 1998; Dankwardt and Hock, 2001; Loiseau and Barriuso, 2002). Similarly, chemical binding (mainly via oxygen-containing bonds), ionic interactions and sequestration of pristine anilazine and its hydrolysis product dihydroxyanilazine were observed with applying silylation. The involved chemical bonds included ester, ether, thio and amino bonds (Klaus et al., 1998). Their incorporations were depicted in Figure 1-11.
14
Figure 1-11. Proposed incorporating mechanisms of atrazine, anilazine and their metabolites into NOM as reported in previous works, adapted from Dankwardt et al. (1998), Dankwardt and Hock (2001) and Klaus et al. (1998).
For the organobromine TBBA, work has been conducted with respect to both its degradation products and their NER formation mechanisms. As demonstrated in Figure 1-12, the detected metabolites of TBBA included tri-, di- and mono-BBA (bromobiphenyl bisphenol A), BA, TBBA mono- and di-methyl ether, and other monocyclic aromatic products containing hydroxyl group(s) and two Br atoms substituted on the benzene ring. Besides sequestration, TBBA and these products can form into NER via ester, ether and carbon-carbon bonds (Li et al., 2015; Liu et al., 2013; Tong et al., 2016).
Figure 1-12. Proposed incorporating mechanisms of TBBA and its metabolites into NOM as reported in previous works, adapted from Li et al. (2015), Liu et al. (2013) and Tong et al. (2016).
15 Given the fact that the researches on nonextractable organofluorines are not as sufficient as those on nonextractable organochlorines, knowledge of their formation mechanisms (particularly regarding to PFCs) are still in a preliminary stage. Although they have been observed as nondesorbable or irreversibly sequestrated residues in many works (Dasu et al., 2012; Enevoldsen and Juhler, 2010; Lee and Mabury, 2017; Wang et al., 2009; Washington et al., 2014, 2009; Zhao et al., 2016, 2013), more works with respect to their bindings to organic matter were focus on their toxicological behaviors towards human and animal protein tissue such as serum albumin and thyroid hormone transport protein (D’Eon et al., 2010; Weiss et al., 2009). Works from Longstaffe et al. (2012, 2010), Shirzadi et al. (2008a, 2008b), Xiao et al. (2019) and Zhao et al. (2014) with NMR spectroscopic measurement on both aliphatic and aromatic organofluorines with different functional groups revealed that electronegativity and electron density might be the key roles in the binding interactions between the compounds and NOM, followed by hydrophobic effects. Aliphatic and aromatic domains of NOM macromolecules showed preferred binding of PFOA and perfluoronaphthanol, respectively. F and Cl atoms presented stronger affinities towards NOM as compared with carboxyl groups, although carboxyl groups dominated in binding to NOM in the case of halogen atoms being absent. The aforementioned association mechanisms between different organofluorines and NOM were illustrated in Figure 1-13. Note that these associations include both strong and weak bindings, and therefore are not exclusively account for NER formation. Additionally, so far there is still no solid evidence for any chemical binding except of a Michael addition reaction between fluorotelomer unsaturated aldehydes (with an unsaturated β carbon) and biological nucleophilic amino acids which formed some conjugates (Rand and Mabury, 2012).
16
Figure 1-13. Proposed association of organofluorines (aliphatic and aromatic) towards NOM from previous works, adapted from Klaus et al. (1998) and Shirzadi et al. (2008a, 2008b).
NOM (content, constituents and macromolecular structures), aging, aerobic/anaerobic conditions, microbial activities (biochemical reactions) and the nature of the studied AOHs (functional groups, aliphaticiy/aromaticity, ionic charges and degree of halogenation) were demonstrated to be vital to the formation of NER-AOHs. By incubation and sorption-desorption experiments, the content of NOM in soil was shown to be positively correlated to the NER yields of pentachlorophenol, dichloroaniline, TBBA and PFCs (Heim et al., 1995; McAvoy et al., 2016; Wen et al., 2009; Zhao et al., 2016). Likewise, aging was indicated to be able to result a further sequestration (or shifts from desorbable to nondesorbable) of atrazine, PFCs and aminocyclopyrachlor into soils by incubation experiments (Mendes et al., 2017; Park et al., 2004; Zhao et al., 2016). Microbial activities are quite essential for the formation of most NER-AOHs as evidenced by several studies with comparison of sterilized and nonsterilized conditions (Bhandari et al., 1997; Gevao et al., 2005; Hatcher et al., 1993; Heim et al., 1994; Liebich et al., 1999; Liu et al., 2013; Loiseau and Barriuso, 2002; McAvoy et al., 2016). And aerobic-anaerobic conditions can lead to different patterns of AOHs behaviors and fate with regard to NER formation, transformation and mineralization (Bhandari et al., 1997; Liu et al., 2013; McAvoy et al., 2016). Different NOM constituents such as humic acids, fulvic acids and humin will behave as NER sinks for different
17 AOHs (Celi et al., 1997; Heim et al., 1995; Li et al., 2015; Liebich et al., 1999; Loiseau and Barriuso, 2002). The aliphatic and aromatic moieties could attract AOHs with similar structures (aliphatic and aromatic, respectively) (Longstaffe et al., 2010; Tong et al., 2016; Zhao et al., 2014). Also lignin- and protein- derived components played different roles in the bindings of perfluoro-2- naphthol and perfluorooctanoic acid, respectively (Longstaffe et al., 2012, 2010). For the AOH compounds, functional groups are critical for NER formation. As discovered by (Celi et al., 1997; Klaus et al., 1998; Kronimus and Schwarzbauer, 2007; Li et al., 2015; Michaelis et al., 1995; Riefer et al., 2017; Shirzadi et al., 2008b), carboxyl, amino and hydroxyl groups are effective for covalent bindings (in the order of carboxyl > amino > hydroxyl), whereas F and Cl showed a stronger affinity to NOM with nonchemical forces.
Besides, remobilization and bioavailability of formerly nonextractable chloroaniline, anilazine, 4- chloro-2-methylphenol (as the metabolites of bound 4-chloro-2-methylphenoxyacetic acid), TBBA (and its metabolites), chlorotoluron and PFCs were also observed along with the microbial degradations of NOM, during alternation of the redox environment, and under the introduction of higher amount of pollutants (Liebich et al., 1999; Liu et al., 2013; Riefer et al., 2017; Scheunert and Schröder, 1998; Suddaby et al., 2016; Zhao et al., 2016).
1.5 Thesis overview
The aim of the present study was to decode the behaviors and fate of AOHs with different molecular features in heterogenous environmental solid matter as NER, particularly with respect to their formation mechanisms and remobilization potential. Field samples were processed with solvent extraction and sequential chemical degradation, and the obtain data of DDXs and other AOHs were discussed separately in chapter 2 and chapter 3, respectively. In addition, the newly concerned contaminants PFCs were studied with incubation for a long-term observation of the NER (in chapter 4).
Chapter 2 and chapter 3 have been already published in peer-reviewed journals.
CHAPTER 2: Field-collected soil, sediment and groundwater sludge samples were measured for extractable and nonextractable DDT and its metabolites (DDXs). Covalent bindings between DDXs and natural organic matter were evidenced. The distribution of NER-DDXs along their degradation pathway showed significant discrepancies within different kinds of matrices. A
18 conceptual model of the fate of NER-DDXs on different environmental pathways (terrestrial and aquatic) was proposed. Published as: Zhu, X., Dsikowitzky, L., Kucher, S., Ricking, M., Schwarzbauer, J. (2019). Formation and fate of point-source nonextractable DDT-related compounds on their environmental aquatic-terrestrial pathway. Environmental Science & Technology, 53(3), 1305-1314.
CHAPTER 3: Heterogenous precontaminated environmental solid substances were processed to investigate the correlation of the molecular structures of different AOHs and their NER formation mechanisms. The remobilization potential of NER-AOHs with different molecular structures was thence assessed. In addition, the NOM macromolecular structures of the studied samples were also explored by chemical degradation. Based on the results, a conceptual model of the incorporating mechanisms of NER-AOHs is proposed. Published as: Zhu, X., Dsikowitzky, L., Ricking, M., Schwarzbauer, J. (2020). Molecular insights into the formation and remobilization potential of nonextractable anthropogenic organohalogens in heterogeneous environmental matrices. Journal of Hazardous Materials, 381, 120959.
CHAPTER 4: A 240-day lab-incubation with spiking PFOS and its alternative F-53B to a fresh topsoil was conducted to mimic their long-term fate as NER. The extractable, moderately bound and deeply bound PFASs were obtained, respectively, after sequential extraction and chemical treatments. The two compounds showed different preferences of incorporating mechanisms into the soil. Remobilization of the NER-PFASs was also observed. A conceptual model of such long- term dynamic behaviors of NER-PFOS and NER-F-53B with some numerical information was proposed accordingly. Submitted to: Science of the Total Environment Zhu, X., Song X., Schwarzbauer, J. First insights into the long-term dynamic behaviors and fate of perfluorooctanesulfonate and its alternative 6:2 chlorinated polyfluorinated ether sulfonate in soil as nonextractable residues.
19 2. Formation and fate of point-source nonextractable DDT-related compounds on their environmental aquatic-terrestrial pathway
ABSTRACT: Nonextractable residues (NER) are pollutants incorporated into the matrix of natural solid matter via different binding mechanisms. They can become bioavailable or remobilize during physical-chemical changes of the surrounding conditions and should thus not be neglected in environmental risk assessment. Sediments, soils and groundwater sludge contaminated with DDXs (DDT, bis(chlorophenyl)trichloroethane; and its metabolites) were treated with solvent extraction, sequential chemical degradation and thermochemolysis to study the fate of NER-DDX along different environmental aquatic-terrestrial pathways. The results showed that DDT and its first degradation products, DDD (bis(chlorophenyl)dichloroethane) and DDE (bis(chlorophenyl)dichloroethylene), were dominant in the free extractable fraction, whereas DDM (bis(chlorophenyl)methane), DBP (bis(chlorophenyl)methanone) and DDA (bis(chlorophenyl)acetic acid) were observed primarily after chemical degradation. The detection of DDA, DDMUBr (bis(chlorophenyl)bromoethylene), DDPU (bis(chlorophenyl)propene) and DDPS (bis(chlorophenyl)propane) after chemical treatments evidenced the covalent bindings between these DDXs and the organic matrix. The identified NER-DDXs were categorized into three groups according to the three-step degradation process of DDT. Their distribution along the different pathways demonstrated significant specificity. Based on the obtained results, a conceptual model of the fate of NER-DDXs on their different environmental aquatic-terrestrial pathways is proposed. This model provides basic knowledge for risk assessment and remediation of both extractable and nonextractable DDT-related contaminations.
20
Graphic abstract for chapter 2
2.1 Introduction
Anthropogenic organic contaminants from industrial, municipal and agricultural emissions are ubiquitous in the aqueous phase, the atmosphere, in natural solid substances as well as in (micro)organisms. They can further be sequestrated as nonextractable residues (NER), which are immobilized in the natural solid phase (Kästner et al., 2014). NER, also known as bound residues, a portion of organic substances not freely available, either temporarily or permanently, because of their strong association with the environmental particulate matter (e.g. soil, sediment, aquatic suspended particular matter), were first recognized in the 1960s (Bailey and White, 1964; Chiba and Morley, 1968). NER cannot be released from solid matrix after mild solvent extractions that do not cause physicochemical changes to the compounds or the underlying matrix (Barriuso et al., 2008; Loos et al., 2012). Nonextractable contaminants and their metabolites are not detectable with normal analytical methods because of their incorporation into solid matrix via physical and chemical interactions, but will get partially bioavailable and can be remobilized by changes in surrounding conditions (e.g. low-weight organic acids (Gao et al., 2015) and biological activities (Xu et al., 1994) may result in release). Thus, the NER-portion of anthropogenic pollutants should not be neglected in environmental risk assessment. The environmental fate and transport of NER compounds are quite different regarding their aquatic or terrestrial pathways. In aquatic systems,
21 sediment, suspended particular matter and colloidal organic matter are the major targets for NER formation. In terrestrial systems, soil is the principal solid sink for NER. Therefore, the effects of these different environmental pathways on the fate of nonextractable contaminants need to be elucidated.
Bis(chlorophenyl)trichloroethane (DDT) was widely used as pesticide since the 1940s and was banned in Europe in the 1970s and 1980s because of its toxicity and environmental persistence. DDT is associated with serious risks to the environment and human health, including carcinogenesis, endocrine disruption and estrogenic action (Kang et al., 2016; Tang et al., 2014). However, DDT and its metabolites (DDXs), particularly bis(chlorophenyl)dichloroethane (DDD) and bis(chlorophenyl)dichloroethylene (DDE), are still frequently worldwide detectable in considerable amounts even decades after the DDT prohibition (Kang et al., 2016). Those metabolites are also reported to be environmentally and biologically harmful (Ghazali et al., 2010; Mrema et al., 2013; Wetterauer et al., 2012). The ongoing usage of the DDT related pesticide dicofol, as well as specialty applications of DDT in antifouling paints and limited disease vector control efforts, result in substantial emission of DDT/DDXs that are cause for continued environmental concern (Bosch et al., 2015; Jin et al., 2015). Consequently, knowledge on nonextractable DDXs is still necessary for both risk assessment and remediation actions.
The formation potential and bioavailability of nonextractable DDT in soil have been investigated for decades, and the 14C-labelling was the mainstream approach (Lichtenstein et al., 1977; Mehetre and Kale, 2008; Samuel et al., 1988; Samuel and Pillai, 1991; Wheeler et al., 1988). With this approach it was proved that NER-DDT and NER-DDE are formed in soil under different matrix conditions, can be released to a low extent and are potentially bioavailable for plants. Sequential chemical degradation has been applied since 2003 to release sedimentary bound DDXs, and a variety of DDT metabolites have been detectable as NER (Frische, 2011; Kronimus and Schwarzbauer, 2007; Kucher and Schwarzbauer, 2017; Schwarzbauer et al., 2003a, 2003b). More recently, pyrolysis and thermochemolysis have been conducted to obtain further DDT derivatives (Kronimus et al., 2006). Covalent linkage of DDA (bis(chlorophenyl)acetic acid) to soil and sediment, which forms NER-DDA was proved by Kalathoor et al. in 2015 (Kalathoor et al., 2015b). Current knowledge on the fate of NER-DDXs (data on their formation are restricted to field samples), especially those beside DDT, DDD and DDE, in natural solid ambient is quite limited, and there is even less research with respect to the aquatic-terrestrial pathway since most of the
22 aforementioned studies solely considered sediment and/or soil. For that reason, further efforts must be carried out to span this knowledge gap.
For a better understanding of the fate of NER-DDXs, soil, sediment and groundwater sludge samples from two highly DDT-contaminated areas were therefore collected for this study to investigate the distribution variation of NER-DDXs. Solvent extraction and sequential chemical degradation, which have been well developed (Frische, 2011; Kronimus et al., 2006; Kucher and Schwarzbauer, 2017; Schwarzbauer et al., 2003b), were applied on these samples to obtain extractable and nonextractable DDXs, respectively. TMAH (tetramethylammonium hydroxide) thermochemolysis was also conducted to obtain further NER-DDT derivatives. This work provides important insights on the fate of DDXs and their different environmental pathways. Such basic knowledge is a prerequisite for risk assessment and remediation.
2.2 Material and methods
2.2.1 Samples Six subaquatic sediment (TC-S1 to TC-S6) and four groundwater sludge (TC-G1 to TC-G4) samples were collected from the Teltow Canal in Berlin, Germany, a highly DDT-contaminated site. A chemical production plant, Berlin Chemie, which produced extraordinarily high amounts of DDT and other pesticides was once located at the beginning of Teltow Canal (Heberer and Dünnbier, 1999). The groundwater was intensively contaminated by the point source because of a bank filtration area of a former drinking water production plant next to it (Frische, 2011).
Two soils (BFW-SOIL1 (5 – 20 cm topsoil) and BFW-SOIL2 (20 – 40 cm subsoil), from a heavily loaded site) and two subaquatic sediment (BFW-S1 and BFW-S2) samples were obtained from the river system of northeastern Bitterfeld-Wolfen, Germany, an industrial megacity, where a wide range of chemical products including pesticides were manufactured (Berger, 2016). For further information on sampling areas and methods see our previous works (Berger et al., 2016; Frische, 2011; Kucher et al., 2018). Sampling locations and information on the samples are shown in Figure 2-1.
23
Figure 2-1. Sampling locations at the Teltow Canal and in Bitterfeld-Wolfen. Total organic matter (foc, dry weight) was determined by loss on ignition.
2.2.2 Extraction Prior to chemical degradation, 10 g of preair dried sample aliquots were pre-extracted sequentially with acetone, mixtures of acetone and n-hexane (1: 1) and n-hexane by accelerated solvent extraction (Dionex ASE 150, Thermo Fisher Scientific, Waltham, MA, USA) to obtain the free extractable fraction (EF). The extracts from the aforementioned three solvents were combined and rotary evaporated at room temperature to reduce the volume. After drying by anhydrous sodium sulfate (Na2SO4) and the removal of elemental sulfur with activated cooper powder, each extract was separated into six factions by micro silica column chromatography using mixtures of n-pentane, dichloromethane and methanol as eluents, as formerly described in detail (Kucher and Schwarzbauer, 2017). A surrogate standard containing 5.82 ng μL-1 4’-fluoroacetophenone, 6.28
-1 -1 ng μL d10-benzophenone and 6.03 ng μL d34-hexadecane was added to each fraction before gas chromatography–mass spectrometry (GC–MS) analyses.
24 2.2.3 Chemical degradation Chemical degradation was carried out according to Schwarzbauer et al. (Schwarzbauer et al., 2005, 2003b) and Kronimus et al. (Kronimus et al., 2006; Kronimus and Schwarzbauer, 2007) to release the nonextractable residues. Pre-extracted samples were sequentially treated with KOH (potassium hydroxide)/methanol, boron tribromide (BBr3) and ruthenium tetroxide (RuO4). TMAH thermochemolysis was also applied on the pre-extracted samples.
2.2.3.1. Alkaline hydrolysis and TMAH thermochemolysis 5.0 g of each pre-extracted sample was placed in a closeable centrifuge glass tube with 2.5 g KOH, 2 mL ultrapure water and 20 mL methanol. After 15 min treatment in an ultrasonic bath, the closed tube was heated at 105 ºC for 24 h. After cooling and decanting the sample solution into a separating funnel, 50 mL ultrapure water was added, and the mixture was acidified to pH 4 to 5. Thereafter, the solution was extracted three times with 30 mL dichloromethane. The combined organic solution was dried and desulfurized.
An aliquot of 150 mg pre-extracted sample was transferred into a DURAN glass tube (8 mm i.d. × 200 mm length) with 200 μL TMAH solution (0.1 mol L-1 in methanol). After gentle hand-shaking of the tube, the methanol was evaporated under a gentle nitrogen stream. The tube was sealed and heated at 250 ºC for 40 min and hereafter cooled and opened. Then, 1 mL n-hexane was added, and the mixture was sonicated for 3 min. After decanting the extract into a glass flask, the residue was extracted again with 1 mL n-hexane. A second round of extraction was repeated with acetone. Then, 5 mL n-hexane was added to the combined extract and the solution was rotary evaporated to remove the acetone. The condensed extract was also dried and desulfurized.
After fractionation with micro silica column chromatography, the surrogate standard solution (Section 2.2.2) was added before GC–MS analysis.
2.2.3.2. BBr3 treatment -1 A BBr3 solution (10 mL, 1.0 mol L in dichloromethane) was placed with a 1.0 g aliquot of an alkaline hydrolyzed sample into a glass centrifuge tube. This mixture was stirred for 30 min at room temperature, sonicated for 15 min and stirred again for 24 h. This procedure was repeated, and subsequently the mixture was sonicated for 15 min. Diethyl ether (10 mL) and 5 mL ultrapure water were added, and the tube was centrifuged for 10 min at 4000g. The supernatant was decanted
25 into a separating funnel and washed twice with 5 mL of ultrapure water. The organic layer was then evaporated, dried, desulfurized, fractionated and the surrogate standard was added.
2.2.3.3. RuO4 oxidation
A 500 mg aliquot of BBr3-treated sample was added into a glass centrifuge tube together with 500 mg sodium periodate, 5 mg RuO4, 8 mL acetonitrile and 8 mL carbon tetrachloride. The mixture was stirred for 4 h at room temperature. Afterwards, 50 μL methanol and two drops concentrated sulfuric acid were added to stop the reaction. The tube was centrifuged for 10 min at 4000g and the supernatant was decanted into a separating funnel. The remaining solid residue was washed again with 3 mL carbon tetrachloride. Ultrapure water (5 mL) was added, and the combined supernatant was washed five times with 10 mL diethyl ether. After evaporation to 0.5 mL, the organic extract was washed again with 0.5 mL of saturated sodium thiosulfate pentahydrate solution. The obtained extract was then dried, desulfurized, fractionated and added with the surrogate standard.
2.2.4 GC–MS analysis GC–MS analyses were performed on a Finnigan PolarisQ ion trap mass spectrometer linked to a trace gas chromatograph (Thermo Finnigan) equipped with a 30 m × 0.25 mm i.d. × 0.25 μm film zebron ZB–5 fused silica capillary column (Phenomenex, Aschaffenburg). The GC oven was programmed as follows: 3 min isothermal time at 60 ºC, followed by heating at 3 ºC min-1 to 310 ºC and held for 20 min. The injection volume was 1 μL in splitless mode at 270 ºC injector temperature. Carrier gas was helium set to a velocity of 30 cm s-1. The mass spectrometer was operated in an EI+ full scan mode scanning from 50 to 650 m/z with a scan time of 0.58 s. Ion source temperature was set at 200 ºC.
DDX identification was conducted by comparison of the detected mass spectra with mass spectral libraries (NIST/EPA/NIH Mass Spectral Library NIST14, Wiley/NBS Registry of Mass Spectral Data, 7th ed.) and was verified by reference compounds. Quantification was carried out by integrating the peak areas of selected ion chromatograms from the total ion current of a measured sample under consideration of the individual response factors of the GC–MS device. Response factors were determined from linear regression functions based on calibration measurements with different concentrations of authentic reference materials (5 points, concentrations ranged within the linear detection range (Frische, 2011; Frische et al., 2010)). The surrogate standard was used
26 for the correction of injection volume and sample volume inaccuracies. The characteristic ions used for DDX quantification are listed in Table S1.
2.3 Results and discussion
2.3.1 Occurrence and concentrations of extractable/nonextractable DDXs in subaquatic sediments, soils and groundwater sludge The DDX concentrations in the sample extracts and the sample residues after each chemical degradation step are shown in Figure 2-2. All concentration values are given in detail in Table S2- 1to S2-5. Interestingly, DDT and its first two anaerobic degradation products, DDD and DDE, were most prominent only in the free extractable fraction (EF) of all sample types. A huge amount of DDT (26000 nmol kg-1) was detected in the EF of sample TC-S1, at the chemical plant outlet to the Teltow Canal. In contrast to samples TC-S2 and TC-S3 from the same location, its foc (total organic matter) was quite low so that DDT occupied even one-third of it. The production and discharge of DDT at this site was stopped decades ago, and thus the sampling site TC-S1 might be an occasional point where DDT largely accumulated and microbial degradation was strongly inhibited.
Alkaline hydrolysis releases compounds incorporated by ester bonds. Further compounds bound by ester and ether bonds to the solid matrix can be obtained by BBr3 treatment. RuO4 oxidation, finally, attacks aromatic structures and activated carbon-carbon bonds (Kronimus and Schwarzbauer, 2007). Consistent with previous studies (Schwarzbauer et al., 2005, 2003b), bis(chlorophenyl)methane (DDM), bis(chlorophenyl)methanone (DBP) and DDA were the main products released after sequential chemical degradation (see Figure 2-2). DDA was only released after alkaline hydrolysis and BBr3 treatment (Figure 2-2 b,c), illustrating that the compound is covalently bound to the solid matrix via ester bonds. A new DDT-related compound, whose mass spectrum (Figure 2-5) was not matched to any database or previous studies, was detected in samples
TC-S2 and TC-S3 after BBr3 treatment, and was tentatively identified as bis(chlorophenyl)bromoethylene (DDMUBr) after comparison with the mass spectrum of DDMU (bis(chlorophenyl)chloroethylene). This compound originates very likely from the cleavage of an ether bond between bis(chlorophenyl)ethylene (DDNU) and the organic matrix by BBr3 (proposed original structure and mechanism see Figure 2-5). Moreover, it was quite interesting that DBP was dominantly detected after RuO4 oxidation. Its possible origins could be the oxidation of other
27 DDXs (e.g. DDNU, DDM and bis(chlorophenyl)methanol (DBH), referring to Memeo et al. (Memeo et al., 2011) and Schouten et al. (Schouten et al., 1998)), a covalently bound DDX obtained by bond cleavage (depicted in Figure 2-5, referring to Memeo et al. (Memeo et al., 2011), Rup et al. (Rup et al., 2010) and Schouten et al. (Schouten et al., 1998), and this origin is more reasonable since former chemical processes have already released a considerable amount of humic moieties) and/or the entrapped DBP released after the structural decomposition of the organic matrix.
It should be noted that the distribution of DDXs in the free extractable and hydrolysis released fractions from the groundwater sludge was quite unique as compared to those in soils and sediments (Figure 2-2a). Bis(chlorophenyl)acetonitrile (DDCN) was detected at a considerably high amount in the free extractable fraction of groundwater sludge. Its specific position in the DDT degradation pathway is to date unknown, but it is surely produced under anaerobic conditions (Albone, E. S., Eglinton, G., Evans, N. C., Rhead, 1972; Heberer and Dünnbier, 1999; Heim and Schwarzbauer, 2013). Those sludge samples were collected only after physical filtration of groundwater without any biological or chemical treatments. A relatively low DDCN concentration was also detected as a free extractable component in sediments from both sampling areas. The polarity of DDCN (Jensen, S., Göthe, R., Kindstedt, 1972) might be a reason, such that it is mobile in the aqueous phase and could be freely transported into groundwater. DDCN was also reported in the EF of submarine sediments (Kucher and Schwarzbauer, 2017) and riverine sediments (Kronimus and Schwarzbauer, 2007). Relatively high amounts of DDT and DDD were released from groundwater sludge after hydrolysis rather than from the other two solid substances. This could be attributed to the high foc level of groundwater sludge (see Figure 2-1), in which DDT and DDD could temporarily be adsorbed and entrapped. And they could get released after the destruction of solid/colloidal organic matter by the hydrolysis procedure. Groundwater samples from the same polluted area were analyzed by Frische et al in 2010 (Frische et al., 2010), and DDT, DDD, DDE, DDA, DDCN, DDMU and DBP were detected.
28
Figure 2-2. DDX concentrations in sample material from two highly DDT-contaminated sites in Germany. All concentrations are given in nmol g-1 dry weight. Figure 2-2a) shows the freely extractable pollutants and Figures 2-2b), c) and d) the pollutants released after different chemical degradation steps in sequence. The dominant DDXs in each step are marked by yellow bars and the compounds with considerable amount observed exclusively in groundwater sludge are marked by green dash lines.
Nonextractable DDXs released after TMAH thermochemolysis are shown in Figure 2-3a. The most abundant products were DDM and DBP, concentrations of which were higher than of those from sequential chemical degradation. Yet total molar concentrations of DDXs obtained by the two different degradation procedures were basically equal. This indicates that several DDXs are transformed to DDM and/or DBP (both could be formed under natural conditions) during thermochemolysis and/or pydrolysis, which was also proved by Kronimus et al. (Kronimus et al., 2006).
Apart from those well-known DDT metabolites, small amounts of bis(chlorophenyl)propene (DDPU) and bis(chlorophenyl)propane (DDPS) (mass spectra see Figure 2-3b) were detectable at the chemical plant outlet to the Teltow Canal. These two compounds were firstly and only reported by Kronimus et al. in 2006 (Kronimus et al., 2006) in sediments from the Teltow Canal (at the emission point) after TMAH thermochemolysis. DDPU was also detectable after pyrolysis of the same samples. They suggested that DDPU and DDPS were incorporated into the matrix by carbon-
29 carbon bonds, and that the additional carbon atoms might belong to the linked macromolecules. However, in our study these two compounds were only detectable at TC-S2 and TC-S3, and at significantly lower concentrations than other dominant DDXs. There are two reasonable assumptions to explain the absence of DDPU and DDPS in the other samples. The first one is that they were solely formed at the emission point with specific ambient conditions such as microbial activity and natural organic matter makeup/concentration. Another possible reason is that the concentrations of DDPU and DDPS in the rest of the samples were too low to be detected.
Figure 2-3. Quantified results of DDXs released by TMAH thermochemolysis (a). The dominant DDXs are marked by yellow bars. Figure 2-3b shows the mass spectra of DDPU and DDPS, two extra DDXs obtained after TMAH thermochemolysis.
2.3.2 Distribution of nonextractable DDXs in the three different types of solid samples The degradation pathway of DDT in subaquatic sediments, groundwater and subsoil can be considered as anaerobic. Based on previous studies (Frische, 2011; Mackintosh et al., 2016; Ricking and Schwarzbauer, 2012; Wetterauer et al., 2012) and with some minor modifications, the anaerobic transformation pathway of DDT (although no available mass balance data for the whole process of direct metabolism) is described in Figure 2-4. For a better understanding of the effects of environmental factors on the behavior and fate of nonextractable DDXs, and according to some previous studies (Frische, 2011; Ricking and Schwarzbauer, 2012), the sequential chemical degradation released NER-DDXs were re-assorted into three groups which are also illustrated in Figure 2-4. First, DDT is dehalogenated and the number of chlorine substituents are reduced (step
30 1). Then, a functional group is formed on the dechlorinated carbon atom (step 2). Finally, the newly formed functional group is removed together with the related carbon atom (step 3). For the topsoil BFW-SOIL1, an aerobic and an anaerobic degradation pathway have to be considered. According to the classic aerobic DDT degradation pathway, DDE, DDMU and DBP should be the main metabolites. Research on DDXs in several surface soils (Huang et al., 2018) proposed that DDT is degraded to DDD, DDE, DDMU, DDNU and DBP in surface soils during both, aerobic and anaerobic processes. Hence, the three-group classification can also be applied to BFW-SOIL1. The total concentrations of step 1, step 2 and step 3 NER-DDXs in the different solid sample types are also displayed in Figure 2-4. These results showed significant distinctions of the NER-DDXs distribution trends between the different types of solid substances. An increase and a decrease of NER-DDXs content along with the degradation process were observed in sediments and soils, respectively (Figure 2-4b and d). In the groundwater sludge, an extremely low amount of step 2 NER-DDXs were found, with a clearly higher level of step-1 and step-3 NER-DDXs. The possible reason for this discrimination could be the incorporating mechanisms.
Figure 2-4. Comparison of the contents of nonextractable DDT-metabolites from three steps of natural degradation (a) in different types of solid sample (b sediments, c groundwater sludge, d soils).
Natural solid particles consist mostly of minerals and organic matter, and the organic matter covers the mineral surfaces or are present as organo-mineral complexes. The incorporation of DDXs into
31 natural solid particles as nonextractable residues can be i) strong adsorption on solid surfaces as well as on the structural pores (e.g., van-der Waals forces and ionic interactions), ii) physical entrapment in the solid matrix structures, or iii) bondage to the organic macromolecules via covalent linkages with functional groups (e.g. carboxylic, hydroxyl, amino and carbonyl groups) (Schwarzbauer and Jovančićević, 2018; Suddaby et al., 2014). These incorporation mechanisms and potential releasing pathways of covalently bound DDXs as well as the detection of corresponding products are illustrated in Figure 2-5. The formation of NER-DDA with sediment organic matter has already been proved to be via ester bonds (Kalathoor et al., 2015b). As mentioned above, DDMUBr was likely released when an ether bond between DDNU and the organic matrix was broken by BBr3, DBP might have been formed after the attack of carbon-carbon double bonds by RuO4, and DDPU and DDPS were suggested to be released after carbon-carbon bond cleavage by TMAH thermochemolysis (Kronimus et al., 2006). The structures of these likely covalently bound precursors (illustrated in Figure 2-5) were all in the step 2-DDX-scope. DDMUBr, DDPU and DDPS were detected only in TC-S2 and TC-S3, and temporally could not be quantified. The amounts of DDA released after alkaline hydrolysis in both, sediments and soils, were much more than those released after BBr3 treatment (Figure 2-5). In contrast, in groundwater sludge samples, DDA was not detected after hydrolysis but after BBr3 treatment. This released portion of NER-DDA was therefore likely non-covalently bound. In this study, the total DDXs level was significantly higher in sediments than in groundwater sludge, whereas the distribution of DBP after
RuO4 oxidation was relatively equal (Figure 2-5). A potential explanation of this is that DBP might preferably tend to accumulate in solid particles of aquatic mobile phases because of the aqueous- affinity of its keto group. A rough quantitative estimation of covalently and non-covalently bound DDXs was made according to the aforementioned hypothesis (the concentrations of alkaline hydrolysis released DDA and bis(chlorophenyl)ethanol (DDOH) and RuO4 oxidation released DBP were summed up as the covalently bound level). 14 to 70 %, 1.7 to 26 %, 25 % and 40 % of NER- DDXs were calculated as covalently bound in sediments, groundwater sludge, BFW-SOIL1 and BFW-SOIL2, respectively.
32
Figure 2-5. Possible forming mechanisms of NER-DDXs and the potential releasing pathways of some covalently bound DDXs.
In general, the DDT metabolites that were detected at the different sampling sites may originate from two sources. The source 1-portion derives from the transport and accumulation of already degraded DDXs from upstream areas. The source 2-portion originates from the in situ degradation of DDXs at this site. In groundwater sludge, the DDT metabolites likely predominantly derived from source 1. The initial metabolites (step 1 NER-DDXs) and the more degraded DDXs (step 3 NER-DDXs) were mostly transported adsorbed on or entrapped in suspended solid particles/colloidal organic matter. Research on the transformation of phenanthrene into NER revealed that microbial activity is essential for the NER formation process with covalent bonds (Wang et al. (Y. Wang et al., 2017)). An active microbial community was also believed to promote irreversible sorption of pesticides to soil (Barriuso et al., 2008; Gevao et al., 2005; Suddaby et al., 2014). This could explain the low proportion of covalently bound DDXs in groundwater sludge,
33 where in general relatively lower and divergent microbial effects occur than surface water (Cecilia et al., 2008; Hazen et al., 1991; Sapkota et al., 2007) as a result of land filtration and the comparatively lower temperature.
The step 2-DDXs, with a relatively higher polarity, have a higher water solubility so that they could have been transported within the aqueous phase over longer distances. During the past DDT production, fine sediment particles with extremely high DDXs loads might have been carried downstream along with the river flow. A considerable portion of those non-strongly bound DDXs then probably partitioned into the aqueous phase and was transported further within this phase. This could explain the low step 1-NER-DDXs content in sediment samples as compared to groundwater sludge and soils, since their incorporation into sediments should be principally non- covalent and thus more reversible. As monitored by Schwarzbauer and Ricking (Schwarzbauer and Ricking, 2010), DDD, DDE, bis(chlorophenyl)chloroethane (DDMS), DDNU, DDA, DDOH, DBP were formerly detected in surface water of the Teltow Canal. DDT metabolites in sediments could be from both sources but source 2 might be more dominant corresponding to their high step 3- NER-DDXs level.
Besides, from the soil incubation experiments of the herbicide 4-chloro-2-methylphenoxyacetic acid into soil by Riefer et al. (Riefer et al., 2017), more significant effects of microbial activity were observed on highly loaded samples. It is reasonable to believe that higher DDXs concentration would promote their incorporation into sediment matrix via covalent bonds. Without frequent rinse of surface water, step 1-DDXs could have been accumulated and have a high residence time in soils. The relatively higher proportion of step 2-DDXs in subsoil than in topsoil indicates the possibility of their vertical transport via soil porous water since they are more hydrophilic than step 1-DDXs. This portion of DDXs might leach to groundwater eventually and the loss of step 2-DDXs would lead to a lower generation of step 3-DDXs in soils. Soil profiles at the same location were formerly studied and DDD was predominantly detected at the depth interval 0 to 6 cm, whereas a lot lower amount of it was observed below 6 cm (even lower than that in downstream sediments) (Berger and Schwarzbauer, 2016). In that case, DDD was suggested to be not vertically mobile. To the best of our knowledge, there is no literature so far dealing with the vertical distribution/transport behaviors of the step 2-DDXs in soils. Therefore, further efforts have to be done to verify the aforementioned leaching potential of step 2-DDXs in soil profiles.
34 2.3.3 DDXs remobilization potential on their environmental pathway The temporarily stored or sequestrated DDXs in the solid phase are possibly remobilized under appropriate conditions. As Li et al. (Li et al., 2018) investigated, the lateral remobilized DDXs from soil contributed to 20 to 42 % of the total DDXs fluxes. Consequently, it is quite critical to understand the remobilization potential of NER-DDXs for risk assessments, as it is important for the bioavailability in aquatic and terrestrial systems and the (eco)toxic potential (Kalathoor et al., 2015a). The risk of DDXs would be underestimated if without considering NER-DDXs. Concentrations of DDA, its precursors as well as of DDCN in EF (defined as EF-risk) plus concentrations of DDA, its metabolites and DDCN in the nonextractable fractions (defined as NER-risk) were calculated as the quantified total DDXs remobilization risk according to Frische, Ricking and Schwarzbauer (Frische, 2011; Ricking and Schwarzbauer, 2012). Considering the extraordinarily high EF-DDT level in TC-S1, EF-risk (concentrations of DDA, its precursors and DDCN in the extractable fraction) at this location was quantified as the average value of TC-S2 and TC-S3. The results are illustrated in Figure 2-6. For sediments, besides the high level at the emission point, the samples downstream possessed a higher total risk and a higher NER-risk proportion than those upstream. This can be attributed to the flushing and carrying activities of surface runoff as mentioned above and the synchronous biotic process. Since there is a bank filtration area through which surface water leaches into groundwater near the groundwater remediation facilities, a high total risk level and low NER-risk proportion in TC-G1 to TC-G4 is reasonable. For soils, a higher total risk level was found in topsoil, whereas a higher NER-risk proportion was observed in subsoil. This points to the existence of a vertical transport of in particular step 2- and step 3- DDXs. Because of the limited number of soil samples that was available for this study this has to be confirmed in future studies. Non-polar DDXs can also be transported vertically via dissolved organic matter in soil (Chabauty et al., 2016), so that an EF- risk was also observable in subsoil.
35
Figure 2-6. Quantified remobilization risk of DDXs at each sampling site.
2.3.4 Conceptual model of the formation and fate of nonextractable DDXs According to the aforementioned results, their interpretation and discussed implications, a conceptual model of the fate of nonextractable DDXs in environmental aquatic-terrestrial pathway is proposed (see Figure 2-7). Even decades after being released, DDT and its metabolites will still remain at the highest level at the emission point as compared to other sites. The special step 2- products DDPU and DDPS will connect strongly and irreversibly to the sedimentary matrix at the emission site with little remobilization potential. The other commonly studied DDXs, however, are subject to remobilization along with the river flow together with the fine sedimentary particles or can be released after the organic matter structures is getting loose. The step 1-DDXs (DDT, DDD, DDE, DDMU, DDMS and DDNU) have a higher remobilization potential than step 2-DDXs (DDA, DDCN, bis(chlorophenyl)ethane (DDEt) and DDOH) and step 3-DDXs (DDM, DBP and DBH), because they are predominantly incorporated in the solid matrix by non-covalent bonds. Although DDA and DDCN (both in step 2) are water soluble, their functional groups will make them more likely to be covalently bound. Consequently, sediments from downstream will collect more free- DDXs as well as NER-DDXs than upstream sediments. In general, the distribution of NER-DDXs in the sedimentary phase will be in the order of step 1 > step 2 > step 3, and a certain amount of step 2 ones can be covalently bound owing to the microbial effects and high DDX level. In soil, DDXs can be from both the exposure to surface water and former flood deposition of sediment material. High microbial activity but fairly low DDX concentration lead to a comparatively low covalent binding extent in soil. Our first insights suggest that step 1-DDXs will be entrapped predominantly in topsoil, whereas step 2-DDXs are subject to vertical transport and can even leach
36 into the groundwater. Therefore, solid particles and colloidal organic matter in groundwater will adsorb or entrap the DDXs from both soil and surface water. These DDXs are bound primarily via non-covalent binding owing to relatively lower microbial effects. As a result, step 2-DDXs will be more in free stage or loosely bound because of their higher remobilization potential and solubility. This model can provide some basic knowledge for risk assessment and further remediation of DDT contaminations in soils and aquatic systems under consideration of both the extractable and the nonextractable portion of the contamination. Due to the limited number of samples, the aging of solid substances and the variability of the distribution of DDXs in different locations should also be considered for later on developing the conceptual model.
Figure 2-7. Conceptual model of the fate of NER-DDXs along the environmental aquatic-terrestrial pathway.
37 2.4 Conclusion
As soil, sediment and groundwater sludge samples from two pre-contaminated sites were processed with solvent extraction, sequential chemical degradation and TMAH thermochemolysis, DDT as well as a lot of its metabolites were detected. Among all the DDXs, DDT and its first degradation products DDD and DDE were dominant in EF, whereas DDM, DBP and DDA were observed primarily after chemical degradation within the tested field samples. The detection of DDA,
DDMUBr, DDPU and DDPS after different chemical treatments (alkaline hydrolysis, BBr3 treatment and TMAH thermochemolysis, respectively) evidenced the covalent bindings between DDXs and the natural organic matrix. The identified NER-DDXs as categorized into three groups according to the three-step natural degradation process of DDT (step 1: DDT, DDD, DDE, DDMU, DDMS and DDNU; step 2: DDA, DDCN, DDEt and DDOH; step 3: DDM, DBP and DBH) showed distinct distributions within different environmental matrices. In sediments, the load of NER-DDXs was in the order of step 1 > step 2 > step 3. In soils, step 1-DDXs was entrapped dominantly in the upper layer, whereas step 2-DDXs could transport deeply even into the groundwater. However, in groundwater sludge, the load of NER-DDXs in step 2 was much lower than those in step 1 and step 3 because of low microbial activities which is vital for covalent bindings (occurred mostly between step 2-DDXs and NOM). In addition, the remobilization potential of DDXs showed a decrease along with the natural degradation process. Based on the results, a conceptual model of the fate of NER-DDXs on their different environmental aquatic- terrestrial pathways was proposed.
38 Table S2-1 The quantified results of DDXs after ASE extraction.
Free extractable (nmol g-1) Samples DDM DDNU DDEt DBP DBH DDCN DDOH DDA DDMU DDMS DDE DDD DDT
TC-S1 3.8 0.16 0.91 59 - 1.3 0.48 24 43 - 110 690 26000
TC-S2 0.76 1.7 5.1 49 33 26 - 4.2 8.2 39 6.8 130 1.6
TC-S3 1.2 2.4 7.7 42 130 25 - 2.4 9.6 57 8.4 110 1.4
TC-S4 0.03 0.07 0.06 0.38 - 0.15 0.12 - - 1.7 0.17 4.7 0.4
TC-S5 0.02 0.13 0.21 1.2 - 0.45 0.55 - 0.83 3.1 1.1 8.3 3.3
TC-S6 0.54 1.2 3.4 24 - 6.0 0.36 0.13 7.2 44 5.6 90 21
TC-G1 0.01 - 0.01 11 - 110 - - 0.60 < 0.01 0.50 71 0.3
TC-G2 - 0.06 0.04 6.1 - 63 - - 0.75 - 1.6 37 8.8
TC-G3 - - 0.02 8.7 - 110 - - 0.85 - 0.92 30 7.3
TC-G4 0.01 0.01 - 68 20 20 - - 0.70 - 1.5 38 63
BFW-S1 1.2 0.23 0.04 0.16 - 0.12 - - 0.45 0.07 0.77 3.9 3.3
BFW-S2 5.6 0.81 0.19 4.6 - 3.7 - - 1.1 - 1.8 12 8.9
BFW-SOIL1 8.8 2.8 0.33 9.1 2.2 0.24 0.95 0.02 1.5 - 1.3 5.7 45
BFW-SOIL2 0.21 0.12 0.02 0.27 - < 0.01 - 0.01 0.04 - 0.03 0.16 2.0 a not detected or under limit of detection.
39 Table S2-2 The quantified results of DDXs after alkaline hydrolysis.
Released after alkaline hydrolysis (nmol g-1) Samples DDM DDNU DDEt DBP DBH DDCN DDOH DDA DDMU DDMS DDE DDD DDT
TC-S1 7.8 0.08 - 5.3 1.8 - - 46 1.2 0.10 0.73 - 0.42
TC-S2 13 5.9 0.20 53 - - - 4.4 9.0 - 1.3 2.3 -
TC-S3 25 12 1.20 44 36.0 - - 13 6.9 - 0.49 0.43 -
TC-S4 ------
TC-S5 3.1 0.93 - 2.4 - - - 0.17 - - - - -
TC-S6 19 6.5 - 8.6 - - 3.40 ------
TC-G1 3.3 - - 0.89 - - - 0.02 8.6 - - 23 44
TC-G2 2.1 0.04 - 0.36 - - - - 0.59 - 0.29 - -
TC-G3 5.3 0.07 - 0.80 - - - - 0.45 - 0.06 0.03 -
TC-G4 1.6 0.04 - 4.8 - - - - 1.7 - 2.6 - -
BFW-S1 ------0.22 -
BFW-S2 0.35 0.22 ------1.8 - - 1.4 -
BFW-SOIL1 - 0.52 - - - - - 1.1 - - - 2.1 -
BFW-SOIL2 0.12 ------0.39 - - - 0.46 -
40
Table S2- 3 The quantified results of DDXs after BBr3 treatment.
-1 Released after BBr3 treatment (nmol g ) Samples DDM DDNU DDEt DBP DBH DDCN DDOH DDA DDMU DDMS DDE DDD DDT
TC-S1 0.10 - - 0.24 - - - 1.3 0.11 - 0.13 - -
TC-S2 4.8 0.72 2.2 14 - - - 0.28 0.85 - - - -
TC-S3 0.25 0.08 0.25 2.6 - - - 0.15 - - - - -
TC-S4 ------
TC-S5 ------
TC-S6 ------
TC-G1 - - - 0.01 - - - - < 0.01 - - 0.01 0.02
TC-G2 0.04 ------0.19 - - - - -
TC-G3 0.05 ------1.4 - - - - -
TC-G4 - - - 0.16 - - - 0.97 - - - - -
BFW-S1 ------
BFW-S2 ------
BFW-SOIL1 0.29 0.05 - - - - - 0.36 - - - - -
BFW-SOIL2 ------
41 Table S2-4 The quantified results of DDXs after RuO4 oxidation.
-1 Released after RuO4 oxidation (nmol g ) Samples DDM DDNU DDEt DBP DBH DDCN DDOH DDA DDMU DDMS DDE DDD DDT
TC-S1 0.93 - - 1.3 ------2.2 - -
TC-S2 - - - 31 ------
TC-S3 - - - 1.5 - - - - - 0.52 - - -
TC-S4 ------
TC-S5 - - - 1.4 ------
TC-S6 - - - 2.2 ------
TC-G1 - - - 1.3 ------
TC-G2 - - - 0.46 ------
TC-G3 - - - 2.9 ------
TC-G4 - - - 2.1 ------
BFW-S1 ------
BFW-S2 ------
BFW-SOIL1 ------
BFW-SOIL2 ------
42 Table S2-5 The quantified results of DDXs after TMAH thermochemolysis.
Released after TMAH thermochemolysis (n mol/g) Samples DDM DDNU DDEt DBP DBH DDCN DDOH DDA DDMU DDMS DDE DDD DDT
TC-S1 33 0.05 - 130 - - - 100 2.60 0.63 4.50 0.58 0.42
TC-S2 84 0.52 9.8 34 - - 0.30 - - - 0.02 - 0.07
TC-S3 170 3.2 20 48 - 0.11 0.33 - 0.05 0.05 - - -
TC-S4 1.1 0.10 0.26 0.10 - - - - 0.10 - - 0.03 0.07
TC-S5 1.0 0.07 0.28 84 - - - - 0.22 - - - 0.12
TC-S6 0.83 0.17 0.10 0.25 ------0.03 0.13
TC-G1 0.34 ------
TC-G2 6.1 - 0.05 ------
TC-G3 0.20 - - 31 - 0.01 ------
TC-G4 9.6 0.03 0.05 0.74 - - 0.20 - 0.05 - 0.01 - -
BFW-S1 ------
BFW-S2 0.43 0.02 0.04 77 - - - - 0.05 - 0.04 - -
BFW-SOIL1 1.1 0.15 0.16 60 - - - - 0.24 - 0.20 - -
BFW-SOIL2 - - - 8.9 ------0.02 - -
43 Alkaline BBr3 RuO Samples Extraction 4 Hydrolysis Treatment Oxidation
TMAH Thermochemolysis
Fractionation F1: Pentane F2: Pentane/DCM Fractionation Fractionation Fractionation Fractionation (95:5 v:v) F1: Pentane F1: Pentane/DCM F1: Pentane/DCM F1: DCM F3: Pentane/DCM F2: Pentane/DCM (50:50 v:v) (95:5 v:v) F2: Ether/Methanol (90:10) (40:60 v:v) F2: DCM F2: DCM (40:60 v:v) F4: Pentane/DCM F3: DCM F3: Methanol F3: Methanol (40:60 v:v) F4: Methanol F5: DCM F6: Methanol
Figure S2-1. Procedure of sample treatment and fractionation of each extract. The volumn of solvent for each elution was 5 mL except that F2 from first extraction was obtained by 8.5 mL of solvent mixture.
44 3. Molecular Insights into the Formation and Remobilization Potential of Nonextractable Anthropogenic Organohalogens in Heterogeneous Environmental Matrices
ABSTRACT: Anthropogenic organohalogens (AOHs) are toxic and persistent pollutants that occur ubiquitously in the environment. An unneglectable portion of them can convert into nonextractable residues (NER) in the natural solid substances. NER-AOHs are not detectable by conventional solvent-extraction, and will get remobilized through changes of surrounding environment. Consequently, the formation and fate of NER-AOHs should be investigated comprehensively. In this study, solvent extraction, sequential chemical degradation and thermochemolysis were applied on different sample matrices (sediments, soils and groundwater sludge, collected from industrial areas) to release extractable and nonextractable AOHs. Covalent linkages were observed most favorable for the hydrophilic-group-containing monocyclic aromatic AOHs (HiMcAr-AOHs) (e.g. halogenated phenols, benzoic acids and anilines) incorporating into the natural organic matter (NOM) as NER. Physical entrapment mainly contributed to the NER formation of hydrophobic monocyclic aromatic AOHs (HoMcAr-AOHs) and polycyclic aromatic AOHs (PcAr-AOHs). The hypothesized remobilization potential of these NER-AOHs follow the order HiMcAr-AOHs > HoMcAr-AOHs/ aliphatic AOHs > PcAr-AOHs. In addition, the NOM macromolecular structures of the studied samples were analyzed. Based on the derived results, a conceptual model of the formation mechanisms of NER-AOHs is proposed. This model provides basic molecular insights that are of high value for risk assessment and remediation of AOHs.
45
Graphic abstract for chapter 3
3.1 Introduction
Anthropogenic organohalogens (AOHs), such as polychlorinated biphenyls (PCBs), organochlorine pesticides and organobromine flame retardants as well as their metabolites, are ubiquitously detected pollutants that originate mainly from industrial productions, municipal emissions and agricultural activities (Alonso et al., 2017; Atashgahi et al., 2018; Makokha et al., 2018). Their toxic nature, massive distribution and persistence in aqueous, terrestrial and atmospheric compartments pose a continuous threat to the biosphere (Alonso et al., 2017; Atashgahi et al., 2018; Jiehong Guo et al., 2017; Isaac-Olive et al., 2018; Makokha et al., 2018). As widely reported, an unneglectable portion of AOHs (e.g. PCBs, DDT (bis(chlorophenyl)trichloroethane), TBBA (tetrabromobisphenol A), atrazine, chlorinated benzenes, chlorinated anilines and chlorinated phenols as well as their degradation products) (Bhandari et al., 1997; Dankwardt et al., 1998; Heim et al., 1995, 1994; Kucher and Schwarzbauer, 2017; Liu et al., 2013; Loiseau and Barriuso, 2002; Luo et al., 2010; McAvoy et al., 2016; Park et al., 2004; Puglisi et al., 2007; Schroll et al., 2003; Schwarzbauer et al., 2003a; Tong et al., 2016; Wen et al., 2009; Yuan et al., 2017) can be incorporated into natural solid substances (soil and sediment, for instance) as nonextractable residues (NER), the so-called bound residues. NER are defined as the portion of organic chemicals immobilized or sequestrated in the solid matrix,
46 particularly in the natural organic matter (NOM). They cannot be released by mild solvent extractions which do not change the physicochemical properties of either the associated matrix or the compounds (Jiehong Guo et al., 2017; Kästner et al., 2016; Poßberg et al., 2016; Trapp et al., 2018; S. Wang et al., 2016; Zhang et al., 2017), but may get remobilized through changes of surrounding environment. These changes include variations of pH values (caused by the weather and/or by the application of acidic herbicides), microbial, plants and earthworms activities, UV radiation and even a certain aging period (Geng et al., 2015; Hartlieb et al., 2003; Liebich et al., 1999; Liu et al., 2015; Riefer et al., 2017; Sabatier et al., 2014; Scheunert and Schröder, 1998; F. Wang et al., 2017). Therefore, the potential risk of NER-AOHs should not be neglected, and a comprehensive interpretation is indispensable.
Several previous studies on the formation route of xenobiotic NER were carried out. Covalent linkage as well as noncovalent adsorption/sequestration have been reported to control the conversion of organic contaminants from free stage to NER. Coherent data showed that the physicochemical properties of both, the compounds and the receiving solid substances, have a strong impact on the formation mechanisms, strength and extent of NER (Fenlon et al., 2011). Concerning the properties of solid substances, fine particles (< 2 µm) were found to be the dominant loading portion for NER, and NOM was the most frequently studied component (Loiseau and Barriuso, 2002). The subfractions of NOM (humin, humic acid, fulvic acid and nonhumics) were compared for the strong capture of heterogeneous xenobiotics including those containing chemical reactive groups (e.g. dichloroaniline, dichlorophenol, bromoxynil, atrazine, anilazine and aromatic amines) and polycyclic aromatic hydrocarbons (PAHs) (Alexanderova, 2016; He et al., 2008; Heim et al., 1995; Liebich et al., 1999; Loiseau and Barriuso, 2002; Nowak et al., 2018; Y. Wang et al., 2017; Yuan et al., 2017). Distinctive NER distribution trends in NOM subfractions, even for the same compounds in different studies, were reported (Heim et al., 1995; Yuan et al., 2017). Therefore, a new perspective, for instance the NOM molecular structure, may contribute to the completion of such knowledge.
Binding of xenobiotics with various molecular structures towards the solid matrix (particularly humic substances) as one form of NER formation was also studied. Barriuso et al (2018) concluded that compounds with chemical reactive groups (e.g. aniline and phenol) yielded larger NER portions, and compounds possessing benzene ring(s) were found to have a higher NER level than those with N-containing heterocycle(s) (Barriuso et al., 2008). Ester, ether, amine and carbon-
47 carbon bonds were evidenced as the covalent bonds linking organic acids (e.g. dichlorobenzoic acid, 4-chloro-2-methylphenoxyacetic acid and metalaxyl acid), phenolic compounds (e.g. TBBA and nonylphenol) and anilines to the NOM, respectively, as NER (Alexanderova, 2016; Gevao et al., 2000; Kalathoor et al., 2015a; Li et al., 2015; Michaelis et al., 1995; Riefer et al., 2017, 2013; Tong et al., 2016). Thence, a higher amount of TBBA was reported to form NER than polybrominated diphenyl ethers owing to the hydroxyl groups of TBBA (Luo et al., 2010). Similar trends were also found between simazine and pyrene (Hartlieb et al., 2003), and between dicamba and trifluralin (Mordaunt et al., 2005). Among the amines, aromatic amines were proved to be covalently bound to humic acid, whereas there was no direct evidence for such binding of aliphatic amines (Alexanderova, 2016). PCBs and PAHs were only reported to be noncovalently bound, unless they were biogenically degraded into their metabolites substituted with oxygen-containing functional group(s). And PAHs with smaller molecular size (two or three rings) were observed to be more favorable for NER formation (Burauel and Führ, 2000; Jiehong Guo et al., 2017; He et al., 2008; Richnow et al., 1994; Yang et al., 2010).
Previous research was mainly conducted using incubation experiments with isotopic labeled target compounds. Field samples were seldom involved, even though they are more suitable to elucidate the composition and distribution patterns of NER-AOHs (Jianyang Guo et al., 2017; He et al., 2008; Luo et al., 2010; Sabatier et al., 2014). Consequently, current knowledge on this issue is still at a more theoretical stage within laboratory scope and lack of sufficient verification in real environmental systems. Besides, previous researches dominantly focused on a single type of sample such as soils or sediments. Therefore, some pieces are still missing for the puzzle regarding the fate of NER-AOHs on their environmental pathway. In the present study, to interrogate the incorporation mechanisms and remobilization potential of NER-AOHs, sediment, soil and groundwater sludge samples from two former industrial areas were collected and proceeded with solvent extraction and sequential chemical degradation. Gas chromatography-mass spectrometry (GC‒MS) based nontarget screening of organohalogens were conducted to investigate the occurrence and structural diversities of NER-AOHs. In addition, the molecular structures of the cleavable NOM moieties of the studied samples were analyzed. This work thus provides general insights on the formation mechanisms of NER-AOHs and on the prognostication of their environmental fate at a molecular level, and will contribute to the basic knowledge for AOHs risk assessment and remediation.
48 3.2 Materials and methods
3.2.1 Samples In total, 14 samples including subaquatic sediments (TC-S1 to TC-S6, BFW-S1 and BFW-S2), soils (BFW-SOIL1 and BFW-SOIL2) and groundwater sludge (TC-G1 to TC-G4) were collected from the area of Teltow Canal, Berlin, Germmany (where formerly a chemical plant was located) and the river system in vicinity to Bitterfeld-Wolfen (an industrial megacity), Germany. TC-S1 to TC-S3 were taken from the outlet site of the chemical plant, and TC-S4 to TC-S6 were taken from the downstream of the canal. The groundwater sludge samples were obtained only after physical filtration from a groundwater treatment facility next to a bank filtration area of a former drinking water mining plant, via which the groundwater has already been proved contaminated (Frische, 2011; Frische et al., 2010). BFW-SOIL1 was topsoil (5 ‒ 20 cm) and BFW-SOIL2 was subsoil (20 ‒ 40 cm). For further information please refer to Frische (2011) and Kucher et al. (2018) (Frische, 2011; Kucher et al., 2018). Detailed sample information are illustrated in Figure S3-1.
3.2.2 Extraction and Chemical Degradation Accelerated solvent extraction (Dionex ASE 150, Thermo Fisher Scientific, Waltham, MA, USA), sequentially with solvents of decreasing polarities (acetone, mixture of acetone and n-hexane (1: 1) and n-hexane) was firstly applied on 10 g air-dried sample aliquots to obtain their free extractable fractions (EF).
To acquire NER, sequential chemical degradation including alkaline hydrolysis, boron tribromide
(BBr3) treatment and ruthenium tetroxide (RuO4) oxidation was carried out on the preextracted samples. In parallel, the solvent-extracted samples were processed with TMAH (tetramethylammonium hydroxide) thermochemolysis (Zhu et al., 2019).
For alkaline hydrolysis, a closeable centrifuge glass tube containing 5.0 g of each preextracted sample, 2.5 g potassium hydroxide, 2 mL ultrapure water and 20 mL methanol was heated for 24 h at 105 ºC after 15 min of ultrasonication. The solution part was then separated, acidified to a pH value of 4 to 5 and added to 50 mL ultrapure water. Afterwards, the liquid mixture was extracted three times with 30 mL dichloromethane and the organic layers were pooled together.
After alkaline hydrolysis, 1.0 g solid residue was placed into a glass centrifuge tube with 10 mL ‒1 BBr3 solution (1.0 mol L in dichloromethane), which was stirred for 30 min followed by 15 min
49 ultrasonication and another stirring for 24 h. The sonicating-stirring step was repeated and the mixture was processed for a final ultrasonication for 15 min. The well-reacted mixture was then enriched with a mixture of 10 mL diethyl ether and 5 mL ultrapure water. The liquid phase was thereafter separated and washed twice with 5 mL ultrapure water to get a BBr3-free organic layer.
Thereafter, 500 mg of each BBr3 treated solid material, 500 mg sodium periodate, 5 mg RuO4, 8 mL acetonitrile and 8 mL carbon tetrachloride were transferred into a new glass centrifuge tube and stirred for 4 h. The reaction was then stopped by adding 50 μL methanol and two drops of concentrated sulfuric acid. The liquid phase was separated and the solid residue was washed with 3 mL carbon tetrachloride. The obtained two liquid phases were combined and 5 mL ultrapure water was added. This mixture was then washed five times with 10 mL diethyl ether to acquire all organic constituents.
In parallel, an aliquot of 150 mg pre-extracted sample and 200 μL of 0.1 mol L‒1 TMAH solution (in methanol) was placed into a DURAN glass tube (8 mm i.d. × 200 mm length) and was hand- shaken. Methanol and oxygen were sequentially removed by evaporation with nitrogen and the tube was sealed immediately. After heating for 40 min at 250 ºC, the tube was opened and the mixture was extracted twice with 1 mL n-hexane and then twice with 1 mL acetone. Afterwards, 5 mL n-hexane was added to the obtained extract and acetone was removed by rotary evaporation.
The extract volumes from each extraction or chemical degradation step were further reduced to 0.5 mL by rotary evaporation. Anhydrous sodium sulfate was used to dry the extract and activated cooper powder was added to remove elemental sulfur. Each extract was fractionated into several fractions by micro silica column chromatography with organic solvents as eluents. Detailed fractionation information was listed in Figure S3-2. Prior to GC‒MS analyses, a surrogate standard ‒1 ‒1 ‒1 containing 5.8 ng μL 4’-fluoroacetophenone, 6.3 ng μL d10-benzophenone and 6.0 ng μL d34- hexadecane was added to each fraction.
3.2.3 GC‒MS Analysis GC‒MS analyses were performed on a trace gas chromatograph (Thermo Finnigan, Germany) equipped with a zebron ZB-5 fused silica capillary column (Phenomenex, Aschaffenburg, Germany, 30 m × 0.25 mm i.d. × 0.25 μm film thickness) and linked to an ion trap mass spectrometer (Finnigan PolarisQ, Germany). The injection mode was splitless at 270 ºC and the injection volume was 1 μL. Carrier gas (helium) velocity was 30 cm s‒1. The oven temperature was
50 set initially at 60 ºC for 3 min, followed by heating to 310 ºC (held for 20 min) at a rate of 3 ºC min‒1. The mass spectrometer was operated in EI+ mode (full scan) with a source temperature of 200 ºC and a scan time of 0.58 s scanning from 50 to 650 m/z.
Mass spectral libraries (NIST/EPA/NIH Mass Spectral Library NIST14, Wiley/NBS Registry of Mass Spectral Data, 7th ed.) were used for initial compound identification. Further on, corresponding reference compounds were measured to verify the compound identification. Quantification of a selected compound was conducted by integration of its characteristic ion chromatograms (see Table S3-1) and consideration of the individual response factor, which was obtained by external calibrations with authentic reference material (linear regression function, 5 point-calibration). The inaccuracies of injection volume and sample volume were corrected with the surrogate standard (Kucher and Schwarzbauer, 2017; Zhu et al., 2019).
3.3 Results and discussion
3.3.1 Occurrence of AOHs and Their Binding Forces towards the Matrices Over 150 AOHs were detected in each sampling area. High concentrations of DDT and its metabolites (DDXs) were found. Owing to their well explored natural degradation pathway (Frische, 2011; Mackintosh et al., 2016; Ricking and Schwarzbauer, 2012; Wetterauer et al., 2012), DDXs were discussed separately in a parallel work (Zhu et al., 2019). The other compounds that were detected frequently from the two areas could either originated from direct industrial emissions or from degradation of their precursors. In the present study, these compounds were discussed with respect to their incorporation into the solid matrix as NER by various mechanisms. These are controlled mainly by their molecular physicochemical properties and the macromolecular discrepancies of the NOM (Fenlon et al., 2011). It has already been concluded by massive works that the forming mechanisms of NER comprise strong adsorption, physical entrapment, covalent binding and biogenic turnover after degradation of the compounds. Biogenic processes incorporate xenobiotics (which are mainly easily biodegradable and with high mineralization rates) into living biomass and particularly into proteins, and such biogenic NER will be released by acid hydrolysis which is valid for destroying proteins (Schäffer et al., 2018). Biogenic NER that were suggested to be of no environmental concern (Kästner et al., 2016; Nowak et al., 2018; Poßberg et al., 2016; Schäffer et al., 2018; Trapp et al., 2018; S. Wang et al., 2016), were not discussed in the present study.
51 The AOHs considered for the present study are illustrated in Figure S3-3 (for their possible industrial and/or municipal origins see Table S3-2). Most AOHs were observed in both EF and as NER, and aromatic halogens were observed more frequently than aliphatic ones. Some (groups of) compounds were exclusively detectable in EF or as NER. Pentachloroethylbenzene, 4-chloro-1- methoxy-2-(phenylmethyl)-benzene, chlorinated methoxydiphenyl ethers, dichlorodiphenyl disulfone, tris(2-chloroisopropyl)phosphate and methoxychlor were only detectable in the EF. Compounds that were detectable as NER are shown in Figure 3-1. These NER compounds were grouped according to their speculated binding forces towards the natural organic matter. Their binding forces were evaluated via their potential of being covalently bound, since different chemical treatments have certain target molecular structures, and therefore, will release products with corresponding functional groups. Alkaline hydrolysis can decompose esters and amides, and will release compounds containing carboxyl, hydroxyl or amino groups. BBr3 treatment attacks ether bonds and the remaining ester bonds, and the macromolecules will be destructed randomly into hydroxyl-substituted and bromine-substituted compounds. Further covalently bound compounds released after RuO4 oxidation are from the carbon-carbon bond cleavage. In parallel, TMAH thermochemolysis methylates some reactive functional groups such as hydroxyls and carboxyl, or will cleave an ether bond via which an aromatic compound was associated to the organic matrix and thereafter, release the corresponding anisole. In our previous investigations, compounds supposed to be formerly bound by carbon-carbon bonds were also detected with an additional carbon which was possibly from the natural organic macromolecule (Kronimus et al., 2006; Kronimus and Schwarzbauer, 2007; Schwarzbauer et al., 2005, 2003a). The supposed bond cleavages are also depicted in Figure 3-1. In this study, those compounds that were released exclusively after specific chemical treatments corresponding to their functional groups were defined as strong covalently bound AOHs, if they were only detected as NER. If they were also found in EF they were assorted as secondary covalently bound AOHs. NER released after multiple chemical procedures, including structurally related (which can generate their characteristic functional groups) and unrelated ones, were categorized as potential covalently bound AOHs. The last group was named as noncovalent bound AOHs as their NER were obtained only after structurally unrelated chemical treatments.
52
Figure 3-1. The categorization of the detected AOHs according to their putative binding mechanisms. Their speculative structures corresponding to certain covalent bonds are orange color-marked. The presumed original molecular structures before covalent bond-cleavages are depicted in blue color.
53 As shown in Figure 3-1, a high structural diversity (for instance, aliphatic, heterocyclic, monocyclic and polycyclic groups) was found among the AOHs bound to the matrix via ethers, amides and carbon-carbon bonds. Whereas compounds incorporated via ester groups were only chlorinated (methyl)phenols and chlorinated benzoic acids. These covalent bonds between NER-xenobiotics and NOM were corroborated by massive previous studies (Alexanderova, 2016; Barriuso et al., 2008; Hartlieb et al., 2003; Kalathoor et al., 2015a; Kronimus et al., 2006; Li et al., 2015; Loos et al., 2012; Michaelis et al., 1995; Mordaunt et al., 2005; Riefer et al., 2013, 2017; Tong et al., 2016; Yuan et al., 2017). Covalently bound AOHs in the present study were predominantly monocyclic. Aliphatic and polycyclic aromatic AOHs were mainly noncovalently bound. As compared to our previous works, similar NER-AOHs occurrence patterns were observed. Chlorinated bromobenzene, ditert-butyl-chlorophenol and chlorinated anisoles were obtained exclusively after
BBr3 treatment, RuO4 oxidation and TMAH thermochemolysis, respectively (Schwarzbauer et al., 2005). Chlorinated ethanes, chlorinated butadienes, chlorinated benzenes, chlorinated naphthalenes and PCBs were released after multiple chemical procedures (Kronimus and Schwarzbauer, 2007; Schwarzbauer et al., 2005, 2003a). Besides, it should be clarified that even though some compounds are assorted into the strong covalently bound group, a portion of them is possibly also noncovalently bound.
3.3.2 NER-AOHs with Different Degree of Chlorination In total 28 (groups of) representative AOHs (92 compounds) were selected for quantification. The partition trends of different AOH congeners/homologues in the extractable and nonextractable fractions varied significantly along with the increase of the chlorination grade. The number of substituted chlorines normally promotes the adsorption potential of the compound to the natural solid substances, particularly the NOM. Other featured molecular structures, for instance aliphatic chains, aromatic rings and groups which are hydrophobic or hydrophilic, simultaneously affect their adsorption potential. As adsorption (including Van der Waals, π binding, hydrophobic binding and hydrogen bonds) (Gevao et al., 2000; Jianyang Guo et al., 2017) always plays a certain role in the formation of NER, it is crucial to understand to which extent it contributes for different AOHs. Thus, the quantified AOH congeners/homologues were arranged with gradual changes of both their NER-Cl degree and their assumed binding forces as illustrated in Figure 3-2. The predicted soil adsorption coefficients (Kd, presented as log Kd) were also included in Figure 3-2. The structure of a compound correlated to a possible covalent bond cleavage is orange-marked in consistence with
54 Figure 3-1. The noncovalently bound compounds were sorted by their EF/NER ratios for the sequence of binding forces.
55 Figure 3-2. Quantified AOH homologous in different solid matrices. Trends of partitions between extractable and nonextractable fractions varied gradually along with the increase of the grades of chlorination of their NER.
Corresponding predicted soil adsorption coefficient (Kd, presented as log Kd) data were taken from the chemistry dashboard of US EPA (https://comptox.epa.gov/dashboard). The speculative structures of AOHs corresponding to certain covalent bonds are color-marked. The degradation pathway of HCHs (hexachlorocyclohexanes) was adapted from Ricking and Schwarzbauer (2008)(Ricking and Schwarzbauer, 2008).
Figure 3-2 shows that the NER-AOH in the left two columns were constituted by their congeners/homologues with only low degree of chlorination. This indicated that the adsorption was not the predominant activity for the NER formation of these AOHs. This concerned in particular chlorinated naphthalenes and chlorinated triphenylmethanes, whose Kd values are much higher than those of the monocyclic chemicals. The NER formation of the noncovalently bound AOHs, should hence be attributed mainly to physical entrapment in voids of the solid matrix. For chlorinated styrenes and chlorinated thioanisoles, covalent binding and physical sequestration both contributed considerably for their NER yields. A further possible reason for the low NER-Cl degree of chlorinated styrenes (the assumed covalent bond cleavage by TMAH thermochemolysis see Figure 3-1) is that the congeners with more Cl substituted on the aliphatic double bond would be less likely stem from the supposed carbon-carbon bond cleavage by TMAH thermochemolysis. This could also be a reasonable explication of the low NER-Cl degree of chlorinated toluenes, chlorinated xylenes and chlorinated dimethylphenols. Compounds in the middle column showed an equivalent degree of chlorination of their EF and NER. In this case, adsorption seems to play a certain role in course of the NER formation of these compounds, and may even be competitive to covalent binding and physical entrapment. The AOHs in the right two columns, in contrast, displayed clearly a higher degree of chlorination in their NER as compared to their EF. This reveals that adsorption was very likely the primary mechanism for NER formation. As concluded by Burauel and Führ (2000), Van der Waals forces and lipophilic bindings composites the interaction between PCBs and soil organic matter because of the hydrophobic nature of PCBs (Burauel and Führ, 2000). In addition, chlorinated anilines and chlorinated benzoic acids are considered to have high potential of covalent bonding as discussed in section 3.3.1. Thence, the NER of these two groups of compounds could be from both covalent binding and adsorption, and these two actions might be complementary.
56 3.3.3 Molecular structures of NOM Moieties As the major acceptor of organic pollutants in the natural solid matrix, NOM, mostly in the form of organo-mineral complexes, plays a crucial role in the formation of NER by means of covalent binding, adsorption and physical sequestration (Gevao et al., 2000; He et al., 2008; Heim et al., 1995; Liebich et al., 1999; Loiseau and Barriuso, 2002; McAvoy et al., 2016; Nowak et al., 2018; Y. Wang et al., 2017; Yuan et al., 2017). Macromolecular structural properties of NOM vary in different substances and their effects on the formation of NER-AOHs should therefore be investigated. In this study, there is a further portion of compounds which were obtained exclusively after BBr3 treatment and/or RuO4 oxidation with Br as the solely contained halogen. These compounds were considered as moieties cleaved from the NOM macromolecules with a Br- substitution on the carbon atoms where ether and/or ester bonds were located. The possible release mechanisms are illustrated in Figure 3-3. These moieties could provide a better insight into the NOM macromolecular structure, since most of the nonhalogenated xenobiotics were already removed after alkaline hydrolysis. This structural information is considered to have a strong relevance for the obtained NER-AOHs because they were released in course of the same chemical procedures.
Br-substituted compounds were categorized in two groups as aliphatic and aromatic, respectively, for a comprehensive understanding. Each group was quantified roughly by their TIC (total ion current) peak areas, and the results were calibrated with the response factors of two representing reference compounds (dibromopentane and dibromobenzene as aliphatic and aromatic, respectively). The detected representative Br-substituted NOM moieties are shown in Figure 3-3. Note that each Br atom indicates a former ether/ ester group. Nevertheless, it should be expounded that these compounds do not demonstrate the whole NOM structure but just a part of the outer organic layer and that NOM fragments released by alkaline hydrolysis are excluded.
57
Figure 3-3. The Br-substituted NOM moieties released after BBr3 treatment and/or RuO4 oxidation. R and R’ refer to the NOM moieties that can be released by BBr3 treatment and RuO4 oxidation, respectively.
In general, as depicted in Figure 3-3, the aliphatic moieties were the predominant products in terms of quantities, and the distribution of aromatic contents followed the order sediments > groundwater sludge > soil. RuO4 oxidation released a slightly higher proportion of aromatic moieties than BBr3 treatment, indicating an increasing aliphaticity from the internal towards the surface of NOM macromolecules. This is in agreement with the classic NOM model that the macromolecule consists of an aromatic core and an aliphatic surface layer (De Melo et al., 2016; Gerke, 2018; Trapp et al., 2018). Besides, for the organo-mineral complexes, it was demonstrated that aromatic and polycyclic aromatic NOM associate closely to the inner mineral layer, and that the lignin-like NOM and aliphatic NOM are subsequently bound as the intermediate layer and the outer-sphere, respectively (Coward et al., 2019).
To the point of molecular structures, apart from the commonly detected moieties, some differences between solid matrices and after different treatment steps were observed. A higher Br level was found in the moieties of sediments and groundwater sludge as compared to soils (see Table S3-14
58 to Table S3-17), suggesting a higher content of ether and/or ester groups in aqueous organic substances than in terrestrial ones. Some specific functional groups particularly those containing nitrogen(s) (nitrile and pyridine, for instance) were observed only in groundwater sludge. Greater structural diversities were inspected among BBr3 released moieties with some unique groups comprising some aromatic heterocycles (e.g. furan, thiazole and pyrimidinol), benzoquinone, benzothiophene, and veratrole. In addition, more aromatic moieties containing hydroxyls and/or carboxyls were found after BBr3 treatment. This evidenced the preferred locating of polar and hydrophilic groups in the surface area of NOM structures (Avneri-Katz et al., 2017).
3.3.4 Conceptual Model of the Formation of Nonextractable AOHs Based on the results and the abovementioned interpretations, a conceptual model of the formation mechanisms of NER-AOHs into NOM including molecular details was developed (Figure 3-4). In accordance with the obtained NOM moieties, a characteristic NOM macromolecule is delineated in Figure 3-4 with substantial aliphatic loads in the outer-sphere and more aromatic rings occupying the central core. Hydrophilic groups (e.g. carboxyls and hydroxyls) and heterocycles are located preferably in the outer-sphere. The detected NER-AOHs are classified as aliphatic AOHs (Al- AOHs), monocyclic aromatic AOHs (McAr-AOHs) and polycyclic aromatic AOHs (PcAr-AOHs). The monocyclic aromatic AOHs are further divided into hydrophobic ones (HoMcAr-AOHs) and hydrophilic functional groups equipped ones (HiMcAr-AOHs), according to their molecular structures. Representative structures of each AOH type are shown in Figure 3-4. The formation of NER-AOHs comprises activities of covalent binding, strong adsorption and physical entrapment. Ester, ether, amide and carbon-carbon bonds are the main covalent bonds between AOHs and NOM. Therefore, the covalent bound AOHs consist mainly of HiMcAr-AOHs. Physical sequestration, on the contrary, prefers HoMcAr-AOHs and PcAr-AOHs rather than HiMcAr-AOHs. The distribution of each type of adsorbed NER-AOHs is more balanced with only small preferences of HiMcAr- AOHs and PcAr-AOHs because of their greater structural affinities to the hydrophilic and hydrophobic NOM constituents, respectively. The occurrence of nonextractable Al-AOHs in this study is less with lower contents than the other groups, and they are mainly noncovalently bound to NOM. Note that the number of the samples (particularly the soils) is quite limited, and the AOHs are from specific point sources with limited types of chemicals. Therefore, this model needs to be confirmed and developed further in future studies.
59
Figure 3-4. A proposed model of the formation mechanisms of NER-AOHs into NOM.
3.3.5 (Re)mobilization Risk of AOHs on Their Environmental Pathway The remobilization potential of NER-AOHs should also be assessed, since nonbiogenic NER are considered to have a toxic potential (Kästner et al., 2016; Nowak et al., 2018; Poßberg et al., 2016; Trapp et al., 2018; S. Wang et al., 2016). The occurrence of AOHs in EF and/or as NER with structural specificities in different types of samples can offer some further insights into their mobility and/or remobilization potential. This information helps to form a basis for risk assessment and remediation. Therefore, the quantified compounds were re-arranged as illustrated in Figure 3- 5. The arrow which shows the increasing direction of binding forces was made with the consideration of both, the possibilities of covalent bonds and the EF/NER ratios. Compounds with high EF/NER ratios are considered to be weaklier bound. The direction of (re)mobilization potential was determined by comparing the total distribution levels (EF + NER) of each compound (congeners/homologues) in the samples from the emission point (TC-S1, TC-S2 and TC-S3) to those in the downstream sample (rest of the sediments, groundwater sludge and soils). Note that the studied samples (particularly the soils) are quite limited to their quantity and pollution sources. Therefore, further studies are needed to confirm and develop these theoretical results.
60
61 Figure 3-5. AOHs with gradual changes of distribution patterns on their environmental pathway (the concentration of each group of compounds was calculated as the sum). The speculative structures of AOHs corresponding to certain covalent bonds are color-marked.
AOHs with lower (re)mobilization potential have an explicitly higher free extractable proportion than the compounds with higher (re)mobilization potential. These high EF/NER-ratio compounds are principally pesticides, antimicrobials and their metabolites (Table S3-2). These compounds are less likely transferred into the surrounding solid matrix via microbial activities, which are quite vital for the formation of NER (Barriuso et al., 2008; Nowak et al., 2013; Osswald et al., 2016). One explanation of their seemingly low mobility could be their high emissions at the industrial outlet. NER loads of chlorinated thiophenes, chlorinated benzenes and chlorinated naphthalenes, as considered to be noncovalently bound, were distinctly higher at the emission point and consisted of components with a low chlorination degree. These lower chlorinated compounds could partly be products of the in situ biogenic dechlorinating process (mainly contributed by microbial activities) of higher chlorinated precursors. These degradation products were simultaneously encased by the solid matrix with limited diffusion because of their hydrophobicities, as discussed in section 3.3.2.
PCBs are a group of widely applied chemicals from both industrial and municipal discharges and are ubiquitous (Meijer et al., 2003). They were detected in fairly equal levels and EF/NER ratios in different types of matrices. The high levels in downstream substances and very low EF/NER- ratios of chlorinated benzaldehydes, chlorinated veratroles, chlorinated benzenethiols and chlorinated chloromethylthiobenzenes, whose NER were believed to be noncovalently bound, could be attributed to their hydrophilic functional groups which might contribute to their aqueous (re)mobilities. In addition, more (groups of) AOHs whose NER are possible to be covalently bound were found to have higher (re)mobilization potential. Ester/amide/ether bonds between xenobiotics and NOM are reported to be reversible, and these NER (in the present study they were chlorinated phenols, chlorinated anilines, chlorinated benzoic acids and dichloromethoxylpyridinamine) may later be released through changes of the surrounding matrix condition (e.g., pH value changes) (Geng et al., 2015; Riefer et al., 2017, 2013). Thereby, the higher (re)mobilization potential of these NER compounds could be attributed either to their reversible binding or to their hydrophilic functional groups.
62 3.4 Conclusion
After solvent extraction, sequential chemical degradation and TMAH thermochemolysis were applied on different sample matrices (sediments, soils and groundwater sludge, collected from industrial areas) over 150 AOHs were detected in EF and/or as NER. AOHs with different molecular structures showed distinct NER formation mechanisms. Based on their occurrence after each step of chemical degradation/ thermochemolysis, those detected NER-AOHs were categorized into groups with regard to their possibilities of being covalently bound according to their specific functional groups and the target chemical bond(s) of each treatment. Covalent bindings were found most favorable for monocyclic aromatic AOHs containing hydrophilic groups (HiMcAr-AOHs, for instance halogenated phenols, benzoic acids and anilines) incorporating into NOM as NER. Many detected compounds were congeners and/or homologues with different levels of chlorine-substitution，and were therefore quantified both in EF and as NER. The change of ratio EF/NER with increasing degree chlorination of each homologous group inflected the role of strong adsorption played during their NER formation. Yet the NER formation of hydrophobic monocyclic aromatic AOHs (HoMcAr-AOHs) and polycyclic aromatic AOHs (PcAr-AOHs) was mainly attributed to physical entrapment. Besides, the hypothesized remobilization potential of these NER-AOHs follow the order HiMcAr-AOHs > HoMcAr-AOHs/ aliphatic AOHs > PcAr- AOHs. In addition, the NOM macromolecular structures of the studied samples were investigated in accordance with the Br-containing products exclusively obtained after BBr3 treatment and/or
RuO4 oxidation. The results suggested a higher content of ether and/or ester groups in aqueous NOM macromolecules than in terrestrial ones. Groundwater sludge was observed to contain more nitrogen-containing groups (for instance nitrile and pyridine). Based on the aforementioned results, a conceptual model of the formation mechanisms of NER-AOHs is proposed.
63
Figure S3-1. Sampling locations of the studied samples and their foc (total organic content) were determined by loss on ignition.
Alkaline BBr RuO Samples Extraction 3 4 Hydrolysis Treatment Oxidation
TMAH Thermochemolysis
Fractionation F1: Pentane F2: Pentane/DCM Fractionation Fractionation Fractionation Fractionation (95:5 v:v) F1: Pentane F1: F1: F1: DCM F3: Pentane/DCM F2: Pentane/DCM Pentane/DCM Pentane/DCM F2: Ether/Methanol (90:10) (40:60 v:v) (50:50 v:v) (95:5 v:v) (40:60 v:v) F4: Pentane/DCM F2: DCM F2: DCM (40:60 v:v) F3: DCM F3: Methanol F5: DCM F4: Methanol F3: Methanol F6: Methanol
Figure S3-2. Procedure of sample treatment and fractionation of each extract.
64
Figure S3-3. AOHs detected in EF and/or as NER.
65 Table S3- 1 The characteristic ions for quantification
Representative Characteristic No. Compounds Formula MW structure ion
1 Dichlorothiophene C4H2Cl2S 152 152
2 Trichlorothiophene C4HCl3S 186 186
3 Tetrachlorothiophene C4Cl4S 220 222
Tetrachlorocyclohexene 4 C6H6Cl4 218 147 TCCH
Pentachlorocyclohexene 5 C6H5Cl5 252 181 PCCH
Hexachlorocyclohexane 6 C6H6Cl6 288 219 HCH
7 Dichlorobenzene C6H4Cl2 146 146
8 Trichlorobenzene C6H3Cl3 180 180
9 Tetrachlorobenzene C6H2Cl4 214 214
10 Pentachlorobenzene C6HCl5 248 250
11 Hexachlorobenzene C6Cl6 282 284
12 Dibromochlorobenzene C6H3Br2Cl 268 270
13 Chlorostyrene C 8 H 7 Cl 138 138
14 Dichlorostyrene C 8 H 6 Cl 2 172 172
15 Trichlorostyrene C 8 H 5 Cl 3 206 206
16 Tetrachlorostyrene C 8 H 4 Cl 4 240 240
66
17 Pentachlorostyrene C8H3Cl5 274 276
18 Hexachlorostyrene C8H2Cl6 308 310
19 Heptachlorostyrene C8HCl7 342 344
20 Octachlorostyrene C8Cl8 376 380
21 Chlorophenol C6H4Cl2O 128 128
22 Dichlorophenol C6H4Cl2O 162 162
23 TriChlorophenol C6H3Cl3O 196 196
24 Tetrachlorophenol C6H2Cl4O 230 232
25 Ditert-butyl-chlorophenol C14H21ClO 240 225
26 Chlorobenzaldehyde C7H5ClO 140 139
27 Dichlorobenzaldehyde C7H4Cl2O 174 173
28 Trichlorobenzaldehyde C7H3Cl3O 208 207
29 Chloroanisole C7H7ClO 142 142
30 Dichloroanisole C7H6Cl2O 176 176
67
31 Trichloroanisole C7H5Cl3O 210 212
32 Tetrachloroanisole C7H4Cl4O 244 246
33 Pentachloroanisole C7H3Cl5O 278 280
34 Chloroveratrole C8H9ClO2 172 172
35 Dichloroveratrole C8H8Cl2O2 206 206
36 Trichloroveratrole C8H7Cl3O2 240 240
37 Tetrachloroveratrole C8H6Cl4O2 274 276
38 Chlorobenzoic acid C7H5ClO2 156 139
39 Trichlorobenzoic acid C7H3Cl3O2 224 224
40 Chloroaniline C6H6ClN 127 127
41 Dichloroaniline C6H5Cl2N 161 161
42 Trichloroaniline C6H4Cl3N 195 195
43 Pentachloroaniline C6H2Cl5N 263 265
44 dichloropyridineamino C5H4Cl2N2 162 162
Dichloro-methoxy-pyridin- 45 C6H6Cl2N2O 192 192 ylamine
68
46 Chlorobenzenethiol C6H5ClS 144 144
47 Dichlorobenzenethiol C6H4Cl2S 178 178
48 Chlorothioanisole C7H7ClS 158 158
49 Dichlorothioanisole C7H6Cl2S 192 192
50 Trichlorothioanisole C7H5Cl3S 226 226
51 Tetrachlorothioanisole C7H4Cl4S 260 262
52 Pentachlorothioanisole C7H3Cl5S 294 296
53 Chloromethyl phenyl sulfide C7H7ClS 158 123
chloro-(chloromethyl)thio- 54 C7H6Cl2S 192 157 benzene
Diochloro- 55 C7H5Cl3S 226 191 (chloromethyl)thio-benzene
56 Chloronaphthalene C10H7Cl 162 162
57 Dichloronaphthalene C10H6Cl2 196 196
58 Trichloronaphthalene C10H5Cl3 230 230
59 Tetrachloronaphthalene C10H4Cl4 264 266
69
60 Pentachloronaphthalene C10H3Cl5 298 300
61 Hexachloronaphthalene C10H2Cl6 332 334
62 Heptachloronaphthalene C10HCl7 366 368
63 Octachloronaphthalene C10Cl8 400 404
64 Chloro-Isopropylnaphthalene C13H13Cl 204 189
Dichloro- 65 C13H12Cl2 238 223 Isopropylnaphthalene
Trichloro- 66 C13H11Cl3 272 257 Isopropylnaphthalene
67 Chlorobiphenyl PCB-Cl1 C12H9Cl 188 188
68 Dichlorobiphenyl PCB-Cl2 C12H8Cl2 222 222
69 Trichlorobiphenyl PCB-Cl3 C12H7Cl3 256 256
70 Tetrachlorobiphenyl PCB-Cl4 C12H6Cl4 290 292
71 Pentachlorobiphenyl PCB-Cl5 C12H5Cl5 324 326
72 Hexachlorobiphenyl PCB-Cl6 C12H4Cl6 358 360
73 Heptachlorobiphenyl PCB-Cl7 C12H3Cl7 392 394
70
74 Chlorophenoxybenzene C12H9ClO 204 204
75 Dichlorophenoxybenzene C12H8Cl2O 238 238
76 Trichlorophenoxybenzene C12H7Cl3O 272 272
77 Tetrachlorophenoxybenzene C12H6Cl4O 306 236
Trichloromethoxydiphenyl 78 C13H9Cl3O2 302 302 ether
Tetrachloromethoxydiphenyl 79 C13H8Cl4O2 336 338 ether
Pentachloromethoxydiphenyl 80 C13H7Cl5O2 370 372 ether
81 Dichlorodiphenylsulphide C12H8Cl2S 254 254
82 Trichlorodiphenylsulphide C12H7Cl3S 288 288
83 MDT C16H15Cl3O2 344 227
84 MDD C16H16Cl2O2 310 227
85 MDMU C16H15ClO2 274 274
86 Chlorotriphenylmethane C19H15Cl 278 165
87 Dichlorotriphenylmathane C19H14Cl2 312 165
88 Trichlorotriphenylmethane C19H13Cl3 346 199
71
Trichloro-methyl- 89 C20H15Cl3 360 125 triphenylmethane
1,1-dichloro-2- 90 (chlorophenyl)-2- C14H11Cl3 284 201 phenylethane
91 Dichlorodiphenylsulfone C12H8Cl2SO2 286 286
92 Clobazam C16H13ClN2O2 300 259
72 Table S3-2 Predicted soil adsorption coefficient (Kd, data are from the chemistry dashboard of US EPA (https://comptox.epa.gov/dashboard)) and possible origins of the studied AOHs (EPA and additional references)
predicted average Soil log Compounds Origin Adsorption Kd Coefficient (Kd) L/kg Dichlorobenzene 572 2.8 Trichlorobenzene 1623 3.2 Tetrachlorobenzene 5007 3.7 Colorant, flame retardant, raw material for manufacturing, pesticide, rubber additive, antimicrobial (EPA) Pentachlorobenzene 6790 3.8 Hexachlorobenzene 12200 4.1 Chlorotoluene 376 2.6 Dichlorotoluene 614 2.8 Trichlorotoluene 1932 3.3 reactants of various industrial chemicals such as dyes and Tetrachlorotoluene 1432 3.2 pesticides (EPA: colorant, antimicrobial, hair dye, Pentachlorotoluene 10400 4.0 photoinitiator, antioxidant, catalyst, fragrance) Hexachlorotoluene 15900 4.2 Chloroxylene 426 2.6 polymer precursors (EPA: colorant, antimicrobial, photoinitiator, Dichloroxylene 459 2.7 fragrance) Trichloroxylene 614 2.8 Bromochlorotoluene 668 2.8 Dibromochlorotoluene 3477 3.5 unknown Tribromochlorotoluene no record - Dibromochlorobenzene 5525 3.7 Bromodichlorobenzene 2997 3.5 Bromotrichlorobenzene 5520 3.7 monomers for producing PCBs Bromotetrachlorobenzene 7480 3.9 Bromopentachlorobenzene 11400 4.1 EPA: colorant, antimicrobial, hair dye, skin conditioner, flame Chlorophenol 231 2.4 retardant EPA: colorant, antimicrobial, hair dye, flame retardant, Dichlorophenol 675 2.8 antioxidant Trichlorophenol 2245 3.4 EPA: colorant, antimicrobial, hair dye, flame retardant Tetrachlorophenol 5314 3.7 EPA: colorant, antimicrobial, rubber additive, flame retardant Chloro-methylphenol 197 2.3 EPA: colorant, antimicrobial, hair dye,flavorant, antioxidant, skin Dichloromethylphenol 728 2.9 conditioner, fragrance, flame retardant Trichloromethylphenol 1290 3.1 Chlorodimethylphenol 662 2.8 EPA: antimicrobial, hair dye,flavorant, antioxidant, skin Dichlorodimethylphenol 766 2.9 conditioner, fragrance, flame retardant, UV absorber, heat Trichlorodimethylphenol 927 3.0 stabilizer, catalyst Chlorobenzaldehyde 96 2.0 antioxidant monomer colorant antimicrobial fragrance Dichlorobenzaldehyde 287 2.5 flavorant catalyst (US EPA) Trichlorobenzaldehyde 573 2.8 Chloroaniline 127 2.1 EPA: colorant, hair dye, crosslinker, flame retardant Dichloroaniline 457 2.7 EPA: colorant, hair dye, crosslinker, flame retardant Trichloroaniline 1042 3.0 EPA: colorant, hair dye, flame retardant
73 Pentachloroaniline 23300 4.4 EPA: colorant, antimicrobial, flame retardant, rubber additive Dibromochloroaniline 5520 3.7 Bromodichloroaniline 2390 3.4 unknown Tribromochloroaniline no record - a metabolite of phosphodiesterase 4 inhibitor Roflumilast (Its Dichloropyridinamine 40 1.6 primary clinical use is in the prevention of exacerbations (lung attacks) in severe COPD) Dichloro-methoxy-pyridinamine 31 1.5 unknown Dichloro-ortho-phenylenediamine 122 2.1 EPA: colorant, hair dye, crosslinker, flame retardant Chloroanisole 147 2.2 Dichloroanisole 513 2.7 Trichloroanisole 920 3.0 EPA: colorant, antimicrobial, flame retardant, rubber additive Tetrachloroanisole 4060 3.6 Pentachloroanisole 9110 4.0 Chloroveratrole 707 2.8 Dichloroveratrole 859 2.9 EPA: colorant, antimicrobial Trichloroveratrole 932 3.0 Tetrachloroveratrole 4760 3.7 Dichlorothiophene 349 2.5 Trichlorothiophene 801 2.9 monomers for conducting polymers Tetrachlorothiophene 3250 3.5 Tetrachlorocyclohexene TCCH 1030 3.0 metabolites of HCH Pentachlorocyclohexene PCCH 1060 3.0 Hexachlorocyclohexane HCH 2750 3.4 pesticide (EPA: flame retardant) Chlorostyrene 152 2.2 Dichlorostyrene 550 2.7 Trichlorostyrene 683 2.8 Tetrachlorostyrene 1720 3.2 Chlorinated styrenes (CSs) are not manufactured commercially. They are by-products of electrolytic processes combining carbon Pentachlorostyrene 8235 3.9 and chlorine under elevated temperatures Hexachlorostyrene 14600 4.2 Heptachlorostyrene 42400 4.6 Octachlorostyrene 261000 5.4 Chlorobenzoic acid 37 1.6 EPA: preservative, colorant, antimicrobial, catalyst, skin conditioner, antioxidant, film forming agent, buffer, fragrance, Trichlorobenzoic acid 275 2.4 flavorant
Chlorobiphenyl PCB-Cl1 3147 3.5 Because of their longevity, PCBs are still widely in use, even though their manufacture has declined drastically since the 1960s, Dichlorobiphenyl PCB-Cl2 13017 4.1 their production was banned by United States federal law in
Trichlorobiphenyl PCB-Cl3 43889 4.6 1978, and by the Stockholm Convention on Persistent Organic Pollutants in 2001. Tetrachlorobiphenyl PCB-Cl4 89692 5.0 (wikipedia: plasticizers in paints and cements, stabilizing Pentachlorobiphenyl PCB-Cl5 196625 5.3 additives in flexible PVC coatings of electrical cables and electronic components, pesticide extenders, reactive flame Hexachlorobiphenyl PCB-Cl 832727 5.9 6 retardants and sealants for caulking, adhesives, wood floor finishes and other indutrial usage such as coolants and insulating Heptachlorobiphenyl PCB-Cl7 836500 5.9 fluids (transformer oil) for transformers and capacitors) Chloronaphthalene 1805 3.3 Because of their good electrical properties, weather resistance, Dichloronaphthalene 1608 3.2 low flammability, high chemical, and thermal stability, they were
74 Trichloronaphthalene 8374 3.9 produced and used since the 1930s. PCNs were used as cable insulators and dielectric fluids in transformers and capacitors due Tetrachloronaphthalene 58533 4.8 to their high thermal stability and inertness. They were also used Pentachloronaphthalene 127000 5.1 as engine oil additives, wood preservatives, electroplating masking products and feedstocks for dye production. They are Hexachloronaphthalene 482400 5.7 also byproducts of combustion and chlorinating processes. As Heptachloronaphthalene 1045000 6.0 PCNs are microcontaminants in technical PCB mixtures, they are released into the environment through the use of PCBs and Octachloronaphthalene 859000 5.9 therefore likely to be transported together, having almost the same physical and chemical properties. (EPA: colorant) Chloro-trimethylnaphthalene 3618 3.6 Dichloro-trimethylnaphthalene 27500 4.4 unknown Trichlorotrimethylnaphthalene no record - Chlorobenzenethiol 459 2.7 for a very wide range of industrial use such as synthesis of Dichlorobenzenethiol 1097 3.0 organic dyes Chlorothioanisole 649 2.8 antimicrobial (EPA) metabolite of pentachlorothioanisole Dichlorothioanisole 1544 3.2 Trichlorothioanisole 1357 3.1 metabolites of pentachlorothioanisole Tetrachlorothioanisole 10300 4.0 Pentachlorothioanisole 14500 4.2 flame retardant rubber additive antimicrobial colorant Chloromethyl phenyl sulfide 468 2.7
Chloro-(chloromethyl)thio-benzene 717 2.9 for a very wide range of industrial use such as synthesis Diochloro-(chloromethyl)thio-benzene no record - Trichlorobutadiene 333 2.5
EPA: pesticide, inert ingredient, manufacturing, raw Tetrachloroethane 156 2.2 material, personal care, toothbrush, dental, automotive, probably
foamer, flame retardant, antimicrobial, colorant
Tetrachlorobutadiene 875 2.9 Pentachloroethane 361 2.6 EPA: flame retardant, antimicrobial, colorant Pentachlorobutadiene 849 2.9
Hexachloroethane 2000 3.3 EPA: flame retardant, antimicrobial, colorant Hexachlorobutadiene 9930 4.0 EPA: monomer, fragrance least half of the total production is production total the of half least cleaning and degreasing fluids some of the compounds concerned are used are used concerned of the compounds some as intermediates in the chemical industry, at Hexachlorohexadiene 6770 3.8 as employed industrial solvents and as dry Tris(2-chloroisopropyl)phosphate 229 2.4 flame retardant, building material, for plastic manufactute
Chlorotriphenylmethane no record - dye, pesticides as digestion-affecting poisons applied in the Dichlorotriphenylmethane 25800 4.4 dyebath when the wool is dyed where they penetrate the wool Trichlorotriphenylmethane 45900 4.7 fiber. Trichloro-methyl-triphenylmethane no record - MDT 38500 4.6 MDD 27500 4.4 pesticide and its metabolites According to EPA, MDT could MDE 25600 4.4 also be from flame retardant MDMU no record - 1,1-dichloro-2-(chlorophenyl)-2- industrial impurities of DDT and/or its metabolites (specific 74600 4.9 phenylethane origin is still unknown)
75 Chlorodiphenylmethane 2495 3.4 Trichlorodiphenylmethane 56900 4.8 Tetrachlorodiphenylmethane 56533 4.8 Pentachlorodiphenylmethane no record - 4-chloro-1-methoxy-2-(phenylmethyl)- 5553 3.7 unknown benzene Monochlorophenoxybenzene 7390 3.9 Dichlorophenoxybenzene 30433 4.5 industrial chemicals which are used in heat exchange and Trichlorophenoxybenzene 41084 4.6 chemical synthesis Tetrachlorophenoxybenzene 38040 4.6 combustion products released from municipal and industrial Dichlorodibenzodioxin 82025 4.9 incinerators a metabolite of Triclosan ( a commonly used antimicrobial Trichloromethoxydiphenyl ether 34000 4.5 agent), more lipophilic than the parent compound Tetrachloromethoxydiphenyl ether 57700 4.8 metabolite of Tetraclosan and Pentaclosan (intermediates produced while waste water treatments (oxidative Pentachloromethoxydiphenyl ether 493000 5.7 degradation(UV+chlorination) ) Dichlorodiphenylsulphide 29380 4.5 as high-temperature lubricants, flame-retardants, pesticides and fungicides, and pharmaceuticals for acidophilic granulocyte- Trichlorodiphenylsulphide 84150 4.9 related diseases antimicrobial, flame retardant (EPA) or the oxidation product of Dichlorodiphenylsulfone 2700 3.4 dichlorodiphenylsulphide Dichlorodiphenyl disulfone 747 2.9 a starting material for copolymers manufacture Dibromochloronitromethane 75 1.9 a disinfection by-product in drinking water Ditert-butyl-chlorophenol 10800 4.0 a stabilizers for synthetic rubber
metabolite of avobenzone(sunscreen) (intermediates produced Tert-butyl-dichlorophenol 5480 3.7 while waste water treatments (UV+chlorination)
It is used in formulation or re-packing, at industrial sites and in manufacturing. (https://echa.europa.eu/substance-information/- 1-(3-chlorophenyl)-4-(3- 867 2.9 /substanceinfo/100.052.744) synthetic intermediates/reagents chloropropyl)piperazine for organic reactions/ an impurity of nefazodone hydrochloride (antidepressant)
plasticizer over 20 years after the general shutdown of Pentachloroethylbenzene 1050 3.0 industrial production Clobazam 828 2.9 a medication used as an anxiolytic and an anticonvulsant
76 4. First insights into the long-term dynamic behaviors and fate of perfluorooctanesulfonate and its alternative 6:2 chlorinated polyfluorinated ether sulfonate in soil as nonextractable residues
ABSTRACT: Per- and polyfluoroalkyl substances (PFASs) are persistent and toxic contaminants that are ubiquitous in the environment. They can incorporate into soil as nonextractable residues (NER) which would get remobilized and thus be bioavailable with changes of surrounding conditions. Therefore, there is a need to investigate thoroughly the long-term fate of NER-PFASs. In this study, a 240-day incubation of perfluorooctanesulfonate (PFOS) and its alternative 6:2 chlorinated polyfluorinated ether sulfonate (F-53B) in a topsoil was carried out. Solvent extraction, alkaline hydrolysis and sequential chemical degradation were applied on periodically sampled soil to obtain extractable, moderately bound and deeply bound PFASs, respectively. The results confirmed the formation of NER of both compounds but with different preferences of incorporating mechanisms. NER-PFOS was proofed to be formed predominantly by covalent binding (via head group) and strong adsorption (via tail group). The formation of NER-F-53B was mainly driven by physical entrapment. Both bound compounds within the incubation period showed three-stage behaviors including an initial period with slight release followed by a (re)incorporating stage and a subsequent remobilizing stage. This work provides some first insights on the long-term dynamic behaviors of nonextractable PFASs and will be conducive to their risk assessment and remediation.
77
Graphic abstract for chapter 4
4.1 Introduction
Per- and polyfluoroalkyl substances (PFASs) are well-known xenobiotics and have been extensively utilized for multiple industrial and municipal applications (e.g. food packaging, textiles and aqueous film-forming foams) as well as for consumer products since the 1940s (Hamid et al., 2018; Joerss et al., 2019; Munoz et al., 2019; Wu et al., 2019; Ye et al., 2015). These compounds form a group of persistent organic pollutants owing to their highly thermal and chemical recalcitrance but limited removal efficiency by wastewater treatment plants (Houtz et al., 2013; Huang and Jaffé, 2019; Wang et al., 2019; Zhou et al., 2019). As a result, their pervasive detection in different environmental compartments and organisms has been widely reported during the last decades (Ding et al., 2018; Høisæter et al., 2019; Houtz et al., 2013; Jing et al., 2019; Joerss et al., 2019; Lin et al., 2017; Liu et al., 2019, 2017; MacInnis et al., 2019; Mourier et al., 2019; Munoz et al., 2019; Mussabek et al., 2019; Ruan et al., 2015; Sepulvado et al., 2011; Wang et al., 2013; Wilkinson et al., 2017). According to the previous toxicological studies, it is evident that these contaminants are bio-accumulative and can cause adverse health effects on humans and animals, for instance hormone disruption and cancer (Hansmeier et al., 2014; Kim et al., 2019; Shi et al., 2019; Shoeib et al., 2011; Wang et al., 2018). Once they were emitted into the environment, they will incorporate into the natural solid matter (e.g. soil and sediments) as nonextractable residues
78 (NER, also known as bound residues), aside from other detectable portions in different phases. These nonextractable residues are not detectable by conventional solvent-extraction protocols (which will not change the physicochemical properties of both the compounds and the matrix), but will be likely released after some certain changes of surrounding conditions, such as pH variation, modified redox conditions, precipitation, microbial activities and even aging (Alexanderova, 2016; Geng et al., 2015; Hamid et al., 2018; J. Li et al., 2019; Sabatier et al., 2014; Schäffer et al., 2018). Consequently, negligence of the nonextractable PFASs will lead to an underestimate of PFAS contamination level, and therefore, will bias PFAS risk assessment.
Among them, perfluorooctanesulfonate (PFOS) is the most ubiquitously detected PFAS particularly in solids (Ding et al., 2018; Fitzgerald et al., 2018; Høisæter et al., 2019; Joerss et al., 2019; Liu et al., 2019; Munoz et al., 2019; Mussabek et al., 2019; Oliver et al., 2019; Rankin et al., 2016; Sammut et al., 2017, 2019; Wang et al., 2019, 2020; Zhang et al., 2019), and has been reported to be extremely persistent with a stable environmental loading level in recent years even after its gradually global phase out since 2000 (Christensen et al., 2019; MacInnis et al., 2019; Mourier et al., 2019). 6:2 chlorinated polyfluorinated ether sulfonate (commercially named as F- 53B), as one of the alternatives of PFOS, has been used for over 30 years in the electroplating industry and is being increasingly detected in the environment possibly due to the phase out of PFOS (Wang et al., 2019, 2013; Zhou et al., 2019). It shares similar physicochemical properties with PFOS and has been proved to be as persistent and toxic as PFOS (Chen et al., 2018; Cheng and Ng, 2018; Ti et al., 2018; Wang et al., 2019; Zhang et al., 2015). Therefore, knowledge on the formation mechanisms and long-term fate of NER-PFOS and NER-F-53B are obligatory.
The formation of NER-PFASs in solid substances has been confirmed in numerous previous research (Dasu et al., 2012; Enevoldsen and Juhler, 2010; Lee and Mabury, 2017; Liu et al., 2010a, 2010b; Liu and Lee, 2007, 2005; Wang et al., 2012, 2009; Washington et al., 2009, 2014; Xiao et al., 2019; Zhao et al., 2016, 2013). However, most of these researches are focus on the sorption behaviors of PFASs on different sorbents, and only limited works were conducted with respect to NER-PFOS and NER-F-53B (Enevoldsen and Juhler, 2010; Høisæter et al., 2019; Wei et al., 2019; Zhao et al., 2016). The physicochemical properties of both the compounds (the hydrophobic C-F chain length and the functional head groups) (D’Eon et al., 2010; Feng et al., 2018; Masoom et al., 2015; Shirzadi et al., 2008a, 2008b; Yang et al., 2017) and the solid matrix (e.g. the content and the components of natural organic matter, mineral constituents and specific surface area)
79 (Cabrerizo et al., 2018; F. Li et al., 2019; K. Li et al., 2019; Longstaffe et al., 2012, 2010; Mussabek et al., 2019; Oliver et al., 2019; Pereira et al., 2018; Uwayezu et al., 2019; Wei et al., 2019) were reported to have effects on the formation of NER-PFASs (or the extent of PFASs associating with natural and/or artificial solid sorbents). Hydrophobic and electrostatic interactions, hydrogen bonding and physical sequestration are the main mechanisms for the binding of PFASs to different solid substances according to previous works (Chi et al., 2018; Guo et al., 2019; K. Li et al., 2019; Siriwardena et al., 2019; Uwayezu et al., 2019; Uwayezu, 2018; Wang et al., 2015; Y. Wang et al., 2016; Xiao et al., 2019; Zhou et al., 2010). Nevertheless, aside from the fundamental formation mechanism, the stability of NER- PFASs is even more crucial for risk assessment yet with scarce exploration. To the best of our knowledge still no work has been documented with regard to a long- term (over three months) observation of the behaviors and fate of nonextractable PFASs. Thus, data are lacking for predicting their remobilization potential.
In the present study, a long-term incubation experiment (up to 240 days) was carried out with spiking PFOS and F-53B to an uncontaminated fresh topsoil under aerobic condition. Solvent extraction and alkaline hydrolysis were applied on the periodically sampled soil matter to obtain the extractable and nonextractable PFASs, respectively. A sequential chemical degradation was conducted on the hydrolyzed samples to search for further deeply bound PFASs. This work, to the best of our knowledge, is the first study on the long-term behaviors and fate of nonextractable PFOS and F-53B, therefore provides some first insights and will contribute vital information to their risk assessment and remediation strategies.
4.2 Materials and methods
4.2.1 Materials The topsoil used for incubation was taken from a private garden in western Germany and was tested to be PFAS free. After being air dried, the soil was passed through a 2 mm sieve to obtain fine particles and was thereafter kept in dark at 4 ºC before use. The total organic matter content of the sieved soil was determined to be 11% by loss on ignition (for detailed procedures see Supporting Information).
PFOS (97.4%) was purchased from Th. Geyer GmbH (Renningen, Germany). F-53B (> 98%) was provided by Prof. Xin Song (Institute of Soil Science, Chinese Academy of Sciences, Nanjing) obtained from Shanghai Synica Co., Ltd. (China) (Wei et al., 2019).
80 4.2.2 Incubation The incubation experiment was carried out in triplicate. Each aliquot of 140 g soil was spiked with PFOS and F-53B dissolved in methanol to acquire a mixture with concentrations of 1.95 μg g -1 PFOS (dry weight (dw)) and 2.33 μg g -1 F-53B (dw). After methanol was evaporated, deionized water was subsequently added to each spiked soil to reach 60% of the maximum water-holding capacity (the detailed determination method is given in Supporting Information). An extra aliquot of 140 g soil treated with the same procedure, without PFOS and F-53B, was referred to as the blank control. The incubation was operated in dark and aerobic condition at room temperature. Aliquots of 20 g (dw equivalent) soil mixture were sampled after 5, 10, 30, 60, 120, 180, 240 days.
4.2.3 Extraction Each aforementioned sample (20 g dw) was placed in a flask together with 100 mL of acetonitrile: ultrapure water mixed solution (60: 40). The mixture was first stirred for 30 min and for 24 h connected by a 15 min ultrasonic bath. The mixture was thereafter processed by a second ultrasonic bath followed by separating the liquid phase with 10 min of centrifuge at 4000g. The liquid was reduced by evaporation to 0.5 mL and dried with anhydrous sodium sulfate. Afterwards, the extract (as the extractable fraction (EF)) was reconstituted in acetonitrile with 25% methanol (v/v) before high-performance liquid chromatography‒mass spectrometry (HPLC‒MS) analyses.
4.2.4 Alkaline Hydrolysis 15 g of each preextracted soil was divided into three aliquots, and each aliquot was put in a centrifuge tube with 2.5 g potassium hydroxide, 2 mL ultrapure water and 20 mL methanol. The closed tubes were ultrasonicated for 15 min and heated for 24 h at 105 ºC. Subsequently, the mixtures of each sample were pooled together in a flask with 50 mL of acetonitrile: ultrapure water mixed solution (60: 40) and acidified to a pH value of 4‒5. The flask was then proceeded for 15 min of ultrasonication and 24 h of stirring. The liquid was then separated, evaporated and dehydrated with the same procedures as described in section 4.2.3. Ultimately, the extracts (as NER) were further cleaned up by micro silica column chromatography (see Figure S4-1) and reconstituted in acetonitrile with 25% methanol (v/v) before HPLC‒MS measurement.
4.2.5 Chemical Degradation Sequential chemical degradation was applied on the hydrolyzed substances which were sampled after 30, 60 and 120 days, respectively, to explore some further bound compounds. Briefly, 10 mL
81 -1 BBr3 solution (1.0 mol L in dichloromethane) was added into a glass centrifuge tube containing 1.0 g hydrolyzed residue. The mixture was first stirred for 30 min and ultrasonicated for 15 min. Then a 24 h stirring followed by a 15 min ultrasonication was repeated twice. Afterwards, the mixture was diluted by 10 mL diethyl ether and 5 mL ultrapure water. The liquid phase was separated into a funnel and washed twice with 5 mL ultrapure water. The organic layer was collected, evaporated and dehydrated. Activated cooper powder was used to remove elemental sulfur. The final extract was hereafter fractionated into three fractions by micro silica column chromatography (see Figure S4-1 for detailed information on the eluents) before gas chromatography‒mass spectrometry (GC‒MS) analyses. The ultimate fraction which was eluted by methanol was thereafter reconstituted in acetonitrile with 25% methanol (v/v) for HPLC‒MS measurement.
Subsequently, 500 mg BBr3 treated substance was mixed with 5 mg RuO4, 500 mg sodium periodate, 8 mL acetonitrile and 8 mL carbon tetrachloride (CCl4) in a new glass centrifuge tube. After stirring the tube for 4 h, 50 μL methanol and two drops of concentrated sulfuric acid were added to stop the reaction. 3 mL CCl4 was used to extract the solid residue after the liquid phase was separated into a funnel. The CCl4 extract was also transferred into the funnel together with 5 mL ultrapure water. The liquid mixture was extracted five times with 10 mL diethyl ether, and the resulted organic layers were collected and combined. The extract was thereafter dried, evaporated to a volume of 0.5 mL, washed with 0.5 mL of saturated sodium thiosulfate pentahydrate solution, and was dried again. Finally, the extract was desulfurized, fractionated and proceeded for GC‒MS and the following up HPLC‒MS measurements (details are also given in Figure S4-1).
In parallel, the procedures of BBr3 treatment and RuO4 oxidation were performed respectively on a soil-free sample containing exclusively 19.48 μg PFOS and 23.32 μg F-53B to investigate the changes of their molecular structures by the chemical degradation.
4.2.6 Analytical Methods The analysis of samples in acetonitrile with 25% methanol (v/v) was performed on a HPLC‒MS system. LC separation was operated with a Thermo Finnigan Spectra SYSTEM (P4000) gradient pump equipped with a Phenomenex Kinetex 2.6μ C18 100Å column (150 mm × 3 mm). 10 mmol L-1 ammonium acetate in acetonitrile with 25% methanol (v/v) and in ultrapure water with 25% methanol (v/v) were used as mobile phases A and B, respectively, with a gradient elution program
82 (0‒3 min, 90% B; 3‒20 min, 0% B; 20‒25 min, 0% B; 25‒30 min, 90% B; 30‒35 min, 90% B). The flow rate was 0.3 mL min-1 and the injection volume was 5 μL. The LC system was coupled to a Finnigan MAT LCQ quadrupole ion trap mass spectrometer (Finnigan MAT) equipped with an electrospray ion source (in negative ion mode). The capillary voltage was set at 10 V and the capillary temperature was 200 ºC. Quantification of PFOS and F-53B was carried out by integrating the peak areas of their selected ion chromatograms (m/z 499 and m/z 531, respectively) from the total ion current.
Samples resulted from BBr3 treatment and RuO4 oxidation were measured on a Thermo Finnigan trace gas chromatograph equipped with a Phenomenex ZB-5 fused silica capillary column (30 m × 0.25 mm i.d. × 0.25 μm film thickness) and linked to a Finnigan PolarisQ ion trap mass spectrometer operated in EI+ mode (full scan, from 50 to 650 m/z, 0.58 s scan time) with a source temperature of 200 ºC. The injection volume was 1 μL in splitless mode at 300 ºC. The temperature program was set as follow: isothermal time was 5 min at 40 ºC followed by a heating ramp of 3 ºC min-1 to 160 ºC holding for 3 min, thereafter, the oven was heated again at 10 ºC min-1 to 310 ºC and held for 3 min.
4.3 Results and discussion
4.3.1 Distribution of extractable and nonextractable PFASs in soil along aging. The quantified results of PFOS and F-53B after increasing incubation periods were illustrated in Figure 4-1a and Figure 4-1b, respectively. Note that PFASs obtained after alkaline hydrolysis are temporarily defined as NER (also stated in section 4.2.4) in the present study, and the mass loss is determined as the spiked level of each compound (1.95 and 2.33 μg g -1 (dw) for PFOS and F-53B, respectively) minus EF and NER. Both of the compounds showed consistently raised EF level from day-5 to day-10 together with stable amounts of NER. This early stage could be either considered as a fluctuation period or recognized as a first release of bound compounds. From day-10 to day- 60, there was a conspicuous decline of EF and an increase of NER, suggesting a continuous incorporation of both PFASs into the soil substances. Nevertheless, an opposite trend was observed from day-60 to day-180. And on day-180, the EF-PFASs level was nearly as high as the spiked amount. This is indicative of not only the extreme persistence of both compounds (without volatilization or biodegradation) but also a presumable remobilization of the previously bound
83 PFASs. Afterwards, within the time scale of incubation (as long as 240 days), the distribution of EF and NER kept in steady and similarly to that on day-180.
Figure 4-1. The concentrations of PFOS (panel a) and F-53B (panel b) in soil (dw) as EF and NER (released by alkaline hydrolysis) after each incubation period.
As compared to F-53B, PFOS showed a distinctly higher NER/EF ratio and lower proportion of EF in each incubation period. This indicates that PFOS is more inclined to be bound particularly via sulfonic ester bonds to the soil organic matter (SOM). As the two compounds contain the same sulfonate head group and C-F chain length, the extra ether and CClF2 groups in F-53B molecule might be the cause. Shirzadi et al. tested the affinity of several pesticides containing heterogenous functional groups to soil and humic acid (HA), and found that CF3 played a leading role in the interaction as comparison with carboxyl, sulfonyl, ether, peptide and nitroso groups (Shirzadi et al., 2008b, 2008a). Similar findings were also reported by Longstaffe et al., Masoom et al. and D’Eon et al. that aliphatic organofluorine (like perfluorooctanoic acid) could interact preferentially with the protein-derived constituents of SOM, and such interaction was driven vitally by the CF3 moiety (D’Eon et al., 2010; Longstaffe et al., 2012, 2010; Masoom et al., 2015). Therefore, it can be presumed that owing to the CF3 moiety of PFOS the incorporation of PFOS into the soil substances was greater than that of F-53B, whereas the effects of the ether group within F-53B
84 molecule could be negligible or even negative. Besides, it should be cause of concern that the mass loss of F-53B during the whole incubation period was prominently higher than that of PFOS. As discussed above, the high detection level of both compounds on day-180 and day-240 suggested little possibility of volatilization and biodegradation. Thus, the so-called mass loss would probably be some further deeply bound compounds since alkaline hydrolysis affects only the (sulfonic) ester bonds.
4.3.2 Elucidation of further bound PFASs To decipher the way of the additional portion of PFASs (the so-called mass loss) was bound to the soil, sequential chemical degradation (BBr3 treatment and RuO4 oxidation, as described in section 4.2.5) was applied on the hydrolyzed soil residues of day-30, day-60 and day-120. These chemical treatments were also applied respectively on a mixture of standard PFOS and F-53B (19.48 μg and 23.32 μg, respectively) to study the exclusive effects of each treatment on the free compounds. After degradation reaction each extract was fractionated into polar and nonpolar fractions for GC- MS and HPLC-MS measurements, respectively. The quantification of the obtained per- /polyfluorinated compounds (PFCs) by GC-MS analysis was conducted by integrating peak areas of the characteristic ion (m/z 131) from the total ion current. Since the specific molecular structures of these alterated PFCs were unknown in the present study, their calibration was roughly determined by an aliphatic polyfluorinated chemical perfluoro-1,3-dimethylcyclohexane
(hereinafter referred to as C8H16, purchased from abcr GmbH & Co. KG, Germany, detailed information of this compound was given in Supporting Information).
After BBr3 treatment, the standard mixture compounds produced 0.24 μg PFCs (ion chromatograms were illustrated in Figure 4-2a, the representative mass spectrum of the peak at 17.8 min was depicted in Figure 4-2b) whereas no detectable PFOS or F-53B, indicating a possibly complete transformation of PFOS and F-53B into nonpolar PFCs with high mass weight (could be even Br-substituted) and longer C-F chain. Note that the retention time of C8H16 was 0.76 min, which is a lot earlier than that of the first PFC peak at 9.8 min, therefore, the authentic yield of PFCs should be higher than the calculated level. Production of polar PFCs under this procedure was also possible but could not be evidenced owing to the lack of valid analytical protocol in the present study. Nevertheless, 5.76 μg PFOS and 18.74 μg F-53B were obtained after RuO4 oxidation but with no detection of PFCs, revealing that PFOS is more sensitive to the chemical attack by
RuO4 as compared with F-53B. As RuO4 can cleave carbon-carbon bonds, PFOS and F-53B were
85 probably cut into smaller molecules (for instance polyfluorinated alkanes and acids) with short C- F chain and thus could not be detected via either of the measurements.
Figure 4-2. Ion chromatograms (m/z 131, m/z 181 and m/z 231) of the detected PFCs after chemical degradation were applied on the standard compounds (panel a). A representative mass spectrum (retention time 17.8 min) of the released PFCs (panel b).
The representative ion chromatograms of the released PFCs from the hydrolyzed soil substances after chemical degradation are demonstrated in Figure 4-3a. Likewise a typical mass spectrum (retention time 17.8 min, day-120) was as shown in Figure 4-3b. These peaks were identified at consistent retention time with that of the standard compounds, yet two extra peaks were observed at 7.1 min and 31.6 min, respectively. This indicates that the soil-released PFCs were similar to that of the free compounds under the same chemical treatments but with somehow more diverse molecular structures. The peaks in Figure 4-3a that are smaller as compared with those in Figure 4-2a should be attributed to the low level of PFASs remaining in the soil after solvent extraction and alkaline hydrolysis.
86 Figure 4-3. Ion chromatograms (m/z 131, m/z 181 and m/z 231) of the detected PFCs after chemical degradation were applied on the hydrolyzed soil residues (panel a). A representative mass spectrum (retention time 17.8 min) of the released PFCs (panel b).
The quantified results of the identified PFASs and PFCs were given in Figure 4-4a and Figure 4- 4b, respectively. Unlike the results from the standard compounds, both PFASs and PFCs were detected after applying BBr3 treatment on the soil residues. The detection of PFASs suggests that they were formerly bound to the soil matrix and were probably released via the cleavage of sulfonic ester bonds between the compounds and SOM and/or via the disintegration of SOM macromolecules by BBr3. In addition, the soil residues showed a distinctly higher ratio of the -1 average acquired amount of PFCs (96 ng g ) to the total pristine PFASs level prior to BBr3 treatment (in average 277 ng g -1 PFOS and 876 ng g -1 F-53B as the so-called mass loss in Figure 4-1) than the standard compounds (0.24 μg PFCs out of 19.48 μg PFOS and 23.32 μg F-53B). This means that the soil-released compounds were more inclined to be restructured into nonpolar PFCs with higher mass weight (with longer C-F chain and/or higher Br-substitution) by BBr3. Besides,
PFCs (retention time 20.3, 22.7, 25.0 and 27.4 min) and PFASs were detected after RuO4 oxidation from day-30 and day-60 soil residues, respectively. This evidences the further (and perhaps more deeply) bound PFASs. Furthermore, the amount of the soil-released F-53B was generally higher than that of PFOS, but the mass gap was a lot slighter than that between their mass loss in Figure 4-1. A reasonable speculation for this is that the ether bond, which is one of the main targets of
BBr3 treatment, within the molecule of F-53B was attacked, and an individual compound was then decomposed into two smaller ones with randomly a hydroxyl group or a Br atom substituted on the former ether sites. These small compounds were also undetectable in the present work. Comparing the data of day-30, day-60 and day-120, day-30 and day-60 could be considered as a mid to late NER formation stage since higher amount of deeply bound PFASs and PFCs were observed, and day-120 could be the initial NER remobilization stage on account of the lower detection level of bound NER (identified from both alkaline hydrolysis and chemical degradation). More discussion about the temporal change of the formation and remobilization of NER-PFASs were given in section 4.3.3. It should further be noted that there might be some extra refractory NER-PFASs which could not be released by our sequential chemical degradation. Therefore, the quantified amount of PFASs and PFCs in Figure 4-4 might not cover the whole mass loss in Figure 4-1.
87
Figure 4-4. Quantified results of the obtained PFASs (panel a) and nonpolar PFCs (panel b) released from the alkaline hydrolyzed soil residues (day-30, day-60 and day-120) by BBr3 treatment and RuO4 oxidation.
4.3.3 Dynamic incorporating-remobilizing behaviors of PFASs. As discussed in the above sections, the status of PFASs was not static after incorporated into soil matrix as NER, but was dynamic during the whole period of incubation, particularly with extensive remobilization after day-120. To illuminate such dynamic behaviors of PFASs, the increment and reduction of EF for each compound between two adjacent incubation time points were calculated as the remobilized and (re)incorporated amounts, respectively. The ratios of them to the spiked level was accordingly determined as remobilization rate and (re)incorporation rate (as the indexes to describe the fate of PFASs, given in Figure 4-5a). Since the first recorded time point was day-5, the first pair of indexes were obtained on day-10. As displayed in Figure 4-5a, the incorporating- remobilizing trends of PFOS and F-53B were generally consistent, yet F-53B showed a visibly higher peak level of (re)incorporation rate than PFOS. In addition, there was a short period of remobilization for both compounds before day-30 (around day-20 according to the fitted curves). Afterwards, a (re)incorporation stage was followed by a second remobilization with their cut-off time point at around day-110. Possible mechanisms of the incorporation of PFASs into soil matrix were also depicted in Figure 4-5a. Covalent binding normally happens between the sulfonate head groups of PFASs and the hydroxyl groups of SOM. According to the researches of Masson et al. and D’Eon et al., the nonpolar tail particularly the CF3 moiety of a PFAS was proven to have vitally strong affinity to solid phase and protein (human serum) (D’Eon et al., 2010; Masoom et al., 2015). Feng et al. investigated the mobility of perfluorinated sulfonic acids in a water/SOM system, and presumed that mostly their tail groups (the perfluorinated alkyl chains) associated to or even combined into the SOM layer, whereas their head groups faced towards to the water phase (Feng
88 et al., 2018). Therefore, the strong sorption of PFASs to soil matrix should be driven by the per- /polyfluorinated aliphatic tails towards the hydrophobic soil components. Physical entrapment, on the contrary, has no preference of PFAS molecular structure, and can take place in either SOM or the minerals.
In order to further understand the role of each mechanism on the fate of PFASs over time, the moderately bound PFASs (released by alkaline hydrolysis) and the deeply bound PFASs (i.e. the so-called mass loss in Figure 4-1) were likewise respectively calculated for their corresponding indexes (as illustrated in Figure 4-5b and Figure 4-5c, respectively). The incorporation mechanisms of the moderately bound PFASs could include all of the three aforementioned ones. The corresponding remobilization could be attributed to the breaking of (sulfonic)ester bonds (either between PFASs and SOM or within SOM macromolecules) probably by microbial activities. Nevertheless, the formation of the deeply bound PFASs was considered to be mainly via physical entrapment, thus their remobilization could be the disintegration of both the SOM macromolecules and the soil structure. As revealed in Figure 4-5b, the activities of moderately bound PFOS was significantly more intensive than F-53B. Comparing to Figure 4-5a, the moderately bound PFOS and F-53B showed earlier time points of peak (re)incorporation rate and peak remobilization rate, respectively, and without the short remobilization before day-30. Their cut-off time points were brought forward to around day-90. In turn, the dynamic behaviors of the deeply bound F-53B was more conspicuous as compared with that of PFOS as delineated in Figure 4-5c. A smooth and delayed peak (re)incorporation rate of deeply bound PFOS was observed. Besides, the cut-off time points of PFOS and F-53B were also retarded to around day-140 and day-120, respectively.
89
Figure 4-5. The dynamic incorporating-remobilizing indexes of PFASs with various possible mechanisms along the increased incubation period. The data of the total bound PFASs, moderately bound PFASs and deeply bound PFASs were given in panel a, b and c, respectively.
Therefore, based on the abovementioned results, the deduced long-term behaviors of bound PFASs could be summarized into three stages. Before day-30, as stage one, a small portion of each physically entrapped PFAS got remobilized shortly. After day-30, the dynamic behaviors of bound F-53B was mainly controlled by its physically entrapped constituent, whose stage two (i.e. the (re)incorporating process) ended later (on day-120) than that of the covalently bound and/or adsorbed constituents (on day-90). The second remobilization process was then defined as stage three. In contrast, the covalently bound and/or adsorbed PFOS played a more significant role in the last two stages than the physically entrapped one. Likewise, the stage three of the covalently bound and/or adsorbed PFOS was initiated earlier (on day-90) than that of the rest bound PFOS (on day- 140). Such dynamic patterns of the long-term behaviors of nonextractable PFOS and F-53B were then demonstrated in Figure 4-6 as a conceptual model with some numerical information. Note that this work was a first approach with limited sample type (topsoil) and therefore would provide just some rough insights. Further works with respect to this topic on more solid substances (for instance soils and sediments with different organic/mineral constituents) need to be carried out as supplement to the knowledge on the long-term fate of PFASs in the environment.
90
Figure 4-6. Three-stage behaviors of bound PFOS (panel a) and F-53B (panel b) with different binding mechanisms.
4.4 Conclusion
By spiking PFOS and its alternative F-53B to a topsoil, a 240-day incubation was carried out with applications of solvent extraction, alkaline hydrolysis and sequential chemical degradation on periodically sampled soil. The extractable, moderately bound and deeply bound PFASs, respectively, were obtained after each period of time. The results evidenced the formation of both compounds into NER. The different releasing dominations of the two compounds after the aforementioned procedures suggested their different preferences of incorporating mechanisms. NER-PFOS was proofed to be formed predominantly by covalent binding (via its functional head group) and strong adsorption (via its hydrophobic tail group). The formation of NER-F-53B was, on the contrary, mainly driven by physical entrapment. Both bound compounds within the incubation period showed three-stage behaviors including an initial period with slight release (stage one) followed by a (re)incorporating stage (stage two) and a subsequent remobilizing stage (stage three) with different lengths of time for different compounds and binding mechanisms. A conceptual model of such long-term dynamic behaviors of nonextractable PFOS and F-53B with
91 some numerical information was thence proposed according to the obtained results and the derived speculation.
92 Alkaline BBr3 RuO Extraction 4 Samples Hydrolysis Treatment Oxidation
Clean up Fractionation Fractionation F1: Pentane (3 mL) F1: Pentane/DCM F1: DCM F2: Acetonitrile (7 mL) (95:5 v:v) F2: Ether/Methanol for HPLC‒MS measurement F2: DCM (40:60 v:v) F3: Methanol F3: Methanol
GC-MS measurement
F3: reconstituted in F2: reconstituted in 100 µL acetonitrile : 100 µL acetonitrile : methanol (75 : 25, v : methanol (75 : 25, v : v ) for HPLC‒MS v ) for HPLC‒MS measurement measurement
Figure S4-1. Procedures of each sample treatment and fractionation.
Supporting Information: 1. Total organic matter measurement The total organic matter content in each sample was determined as loss on ignition. The specific procedure was as follows: (1) An empty crucible was dried at 150 ºC for 20 min;
(2) After cooling in desiccator, the crucible was weighted with a scale to 4 decimal places (m1); (3) Approximately 500 mg of each sample was added into the dried crucible; (4) The filled crucible was heated again at 106 ºC for 3 h to remove water;
(5) The dried crucible with dried sample was weighted again (m2); (6) The filled crucible was then heated at 550 ºC for 1 h;
(7) After cooling in desiccator for about 30 min, the crucible was weighted (m3).
The total organic matter content in dry weight was then calculated as !!"!" × 100%. !!"!#
93 2. Determination of maximum water-holding capacity The determination of the maximum water-holding capacity of the studied soil was carried out as follows:
(1) A clean and dry separating funnel with a little glass wool placed at the switch was weighted as m1;
(2) About 100 g dry soil was added and the filled funnel was weighted again as m2; (3) 300 mL deionized water was the added in to the funnel; (4) The funnel was shaken gently to allow the soil to be saturated, and the soil was soaked for 2 h; (5) The funnel switch was thereafter turned on and the system was drained for 2 h;
(6) After the drainage was completed, the funnel was weighted at last as m3;
The maximum water-holding capacity of the studied soil was !""!! × 100%. !""!#
3. Information of perfluoro-1,3-dimethylcyclohexane
Formula: C8F16, MW: 400, CAS: 335-27-3, the molecular structure and the mass spectrum of the chemical are as shown in Fig. S2 (from NIST/ EPA/NIH Mass Spectral Library NIST14).
100 69
F
F F F F F
50 F FF
F F
F F 131 F 181 F F 100 31 119 19 50 62 143 162 193 212 231 293 312 381 0 30 60 90 120 150 180 210 240 270 300 330 360 390 (mainlib) Perfluoro-1,3-dimethylcyclohexane
Figure S4-2. The molecular structure and the mass spectrum of perfluoro-1,3-dimethylcyclohexane.
94 5. General conclusion and outlook
In the present study, both field-collected and lab-incubated samples were investigated with respect to the formation mechanisms and fate of nonextractable anthropogenic organochlorine/bromine and the newly concerned per-/polyfluoroalkyl substances, respectively.
The field samples including soils, sediments and groundwater treatment sludge were from two contaminated sites. Solvent extraction, sequential chemical degradation and TMAH thermochemolysis were applied on these samples to obtain both extractable and nonextractable AOHs. Over 150 AOHs were detected including DDXs (DDT and its metabolites/ derivatives), PCBs, chlorinated benzenes, HCHs, and so on.
Owing to their well explored natural degradation pathway, DDXs were discussed separately. DDT and its first degradation products DDD and DDE were dominant in EF, whereas DDM, DBP and DDA were observed primarily after chemical degradation within the tested field samples. The detection of DDA, DDMUBr, DDPU and DDPS after alkaline hydrolysis, BBr3 treatment and TMAH thermochemolysis, respectively, evidenced the covalent bindings between DDXs and NOM. The identified NER-DDXs as categorized into three groups according to the three-step natural degradation process of DDT (step 1: DDT, DDD, DDE, DDMU, DDMS and DDNU; step 2: DDA, DDCN, DDEt and DDOH; step 3: DDM, DBP and DBH) showed distinct distributions within different environmental matrices. In sediments, the load of NER-DDXs was in the order of step 1 > step 2 > step 3. In soils, step 1-DDXs was entrapped dominantly in the upper layer, whereas step 2-DDXs could transport deeply even into the groundwater. However, in groundwater sludge, the load of NER-DDXs in step 2 was much lower than those in step 1 and step 3 because of low microbial activities which is vital for covalent bindings (occurred mostly between step 2-DDXs and NOM). In addition, the remobilization potential of DDXs showed a decrease along with the natural degradation process. Based on the results, a conceptual model of the fate of NER-DDXs on their different environmental aquatic-terrestrial pathways was proposed.
The further detected AOHs could be originated either from direct industrial emissions or from degradation of their precursors, and therefore were discussed with regard to their molecular structures and preferred NER incorporating mechanisms. These AOHs with different molecular structures showed distinct NER formation mechanisms. Based on their occurrence after each step of chemical degradation/ thermochemolysis, those detected NER-AOHs were categorized into
95 groups with regard to their possibilities of being covalently bound according to their specific functional groups and the target chemical bond(s) of each treatment. Covalent bindings were found most favorable for monocyclic aromatic AOHs containing hydrophilic groups (HiMcAr-AOHs, for instance halogenated phenols, benzoic acids and anilines) incorporating into NOM as NER. Many detected compounds were congeners and/or homologues with different levels of chlorine- substitution，and were therefore quantified both in EF and as NER. The change of ratio EF/NER with increasing degree chlorination of each homologous group inflected the role of strong adsorption played during their NER formation. Yet the NER formation of hydrophobic monocyclic aromatic AOHs (HoMcAr-AOHs) and polycyclic aromatic AOHs (PcAr-AOHs) was mainly attributed to physical entrapment. Besides, the hypothesized remobilization potential of these NER- AOHs follow the order HiMcAr-AOHs > HoMcAr-AOHs/ aliphatic AOHs > PcAr-AOHs. In addition, the NOM macromolecular structures of the studied samples were investigated in accordance with the Br-containing products exclusively obtained after BBr3 treatment and/or RuO4 oxidation. The results suggested a higher content of ether and/or ester groups in aqueous NOM macromolecules than in terrestrial ones. Groundwater sludge was observed to contain more nitrogen-containing groups (for instance nitriles and pyridines). Based on the aforementioned results, a conceptual model of the formation mechanisms of NER-AOHs is proposed.
For the 240-day lab-incubation investigation, PFOS and its alternative F-53B were spiked to a fresh topsoil to mimic their long-term fate in terrestrial environment. Solvent extraction, alkaline hydrolysis and sequential chemical degradation were applied on periodically sampled soil substances. The extractable, moderately bound and deeply bound PFASs, respectively, were obtained after each period of time. The results evidenced the formation of both compounds into NER. The different releasing dominations of the two compounds after the aforementioned procedures suggested their different preferences of incorporating mechanisms. NER-PFOS was proofed to be formed predominantly by covalent binding (via its functional head group) and strong adsorption (via its hydrophobic tail group). The formation of NER-F-53B was, on the contrary, mainly driven by physical entrapment. Both bound compounds within the incubation period showed three-stage behaviors including an initial period with slight release (stage one) followed by a (re)incorporating stage (stage two) and a subsequent remobilizing stage (stage three) with different lengths of time for different compounds and binding mechanisms. A conceptual model of
96 such long-term dynamic behaviors of nonextractable PFOS and F-53B with some numerical information was thence proposed according to the obtained results and the derived speculation.
As the NER incorporating preferences and mechanisms as well as the NER remobilization potential of different AOHs were assessed with specific regard to their molecular structural features (functional groups, degree of chlorination, aliphaticity and aromaticity), this work provides some fundamental knowledge and molecular insights valuable especially for risk assessments and further remediation of AOH contaminations in both soils and aquatic systems. Particularly, the part of PFAS-related work, to the best of our knowledge, contributes the first insights on the long-term fate of PFASs.
Due to the limited number of samples, the conceptual models proposed in this work could just provide some rough insights. The combination of alkaline hydrolysis and sequential chemical degradation is a suitable method for analyzing filed-collected samples. However, only limited previous work was conducted with application of such method. Therefore, samples from more precontaminated sites in different compartments (apart from soils and sediments, aquatic suspended solid particles and air-borne particles are also intriguing matrices for NER investigation) are needed to be uniformly processed with this method to enrich and correct the knowledge of NER-AOHs in practical conditions. Meanwhile, lab-scaled works are still indispensable for perspective of the specific interactions between a (type of) AOH and the solid matrix. Incubations can continue being carried out with spiking of AOH(s) with more diverse physicochemical properties and structural variations under different conditions (e.g. solid matrices with different NOM loads, pH, ionic strength, mineral composition, redox condition as well as microbial (or enzyme) activities). Since the chemical interactions between AOH and NOM are crucial, further works on observation the bindings of AOH(s) towards small organic molecules (as representative of different NOM macromolecular moieties) or natural derived NOM constituents (e.g. acidic/basic/neutral-hydrophilic/hydrophobic fractions which can be obtained by using resin adsorbents) with probably fluorescence spectroscopy (quenching of chemical groups) and/or multiphase NMR technique together with sequential chemical degradation and GC-MS analysis.
Besides, at present there are still a lot of puzzles remaining unsolved in this knowledge system (although have been intensively investigated for decades), for instance the specific macromolecular structures of NOM (with different categorizes), its conversion with respect to functional groups staying exposed to xenobiotics after associated with mineral particles with different properties
97 (surface charges, surface area and the interlayer structures), and even the discrimination of natural and anthropogenic organohalogens. Further works are therefore required, or at least a literature survey on the relevant studies should be updated periodically.
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120
Appendix A: The identified data of AOHs (for chapter 3)
Free extractable
S2 S1 - -
Compounds S1 S2 S3 S4 S5 S6 S3 G1 G2 G4 ------SOIL1 SOIL2 - - - - - TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Trichlorobutadiene + Tetrachloroethane + + + + + + Tetrachlorobutadiene + + + + Pentachloroethane + + + + Pentachlorobutadiene + + + + Hexachloroethane + + + + Hexachlorobutadiene + + + + Hexachlorohexadiene + + + + + + Dichlorothiophene + + + Trichlorothiophene + + + + + Tetrachlorothiophene + + + + + + + + Tetrachlorocyclohexene TCCH + + + + Pentachlorocyclohexene PCCH + + + + + + + + + + + Hexachlorocyclohexane HCH + + + + + + + + + + + + + + Dichlorobenzene + + + + + + + + + + + + + + Trichlorobenzene + + + + + + + + + + + + + + Tetrachlorobenzene + + + + + + + + + + + + Pentachlorobenzene + + + + + + + Hexachlorobenzene + + + + Chlorotoluene + + Dichlorotoluene + + + + + + + + + + + Trichlorotoluene + + + Tetrachlorotoluene + + + + Pentachlorotoluene + + + Hexachlorotoluene + Chloroxylene Dichloroxylene + + + Trichloroxylene + + + Pentachloroethylbenzene + + + + + Bromochlorotoluene Dibromochlorotoluene Tribromochlorotoluene Dibromochlorobenzene Bromodichlorobenzene + Bromo-trichlorobenzene + + Bromotetrachlorobenzene + Bromopentachlorobenzene + +
121
Free extractable
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 S3 G1 G2 G4 ------SOIL1 SOIL2 - - - - - TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorostyrene + + + + + + + + Dichlorostyrene + + + + + + + + + + + + + Trichlorostyrene + + + + + + + + + + + + + + Tetrachlorostyrene + + + + + + + + Pentachlorostyrene + + Hexachlorostyrene + + Heptachlorostyrene + + + Octachlorostyrene + + + Chlorophenol + + + + + Dichlorophenol + + + + + + + + Trichlorophenol + + + + + + + Tetrachlorophenol + Chloromethylphenol Dichloromethylphenol + + + + Trichloromethylphenol Chlorodimethylphenol Dihlorodimethylphenol + + + Trichlorodimethylphenol + + + Ditert-butyl-chlorophenol Tert-butyl-dichlorophenol Chlorobenzaldehyde + + + + + + + Dichlorobenzaldehyde Trichlorobenzaldehyde Chloroanisole + + + + + + + + + Dichloroanisole + + + + + + + + + + + + + + Trichloroanisole + + + + + + + + + + Tetrachloroanisole + + + + + Pentachloroanisole + + + Chloroveratrole + + Dichloroveratrole + + Trichloroveratrole + + + Tetrachloroveratrole + + + Chlorobenzoic acid + + + + + + + + + Trichlorobenzoic acid Chloroaniline + + + Dichloroaniline + + + + + + Trichloroaniline Pentachloroaniline + + Dibromochloroaniline Bromodichloroaniline Tribromochloroaniline Dichloropyridinamine + + + Dichloro-methoxy-pyridinamine + + + Dichloro-ortho-phenylenediamine + + Chlorobenzenethiol + Dichlorobenzenethiol
122
Free extractable
S2 S1 - -
Compounds S1 S2 S3 S4 S5 S6 S3 G1 G2 G4 ------SOIL1 SOIL2 - - - - - TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorothioanisole + Dichlorothioanisole + Trichlorothioanisole + + + Tetrachlorothioanisole + + + Pentachlorothioanisole + + + Chloromethyl phenyl sulfide Chloro-(chloromethyl)thio-benzene + Diochloro-(chloromethyl)thio-benzene + + + Chloronaphthalene + + + + + + Dichloronaphthalene + + + + + + + + + + + + Trichloronaphthalene + + + + + + + Tetrachloronaphthalene + + + Pentachloronaphthalene + + Hexachloronaphthalene + Heptachloronaphthalene + Octachloronaphthalene + Chlorotrimethylnaphthalene Dichlorotrimethylnaphthalene + Trichlorotrimethylnaphthalene
Chlorobiphenyl PCB-Cl1 + + + + + + + + + + + +
Dichlorobiphenyl PCB-Cl2 + + + + + + + +
Trichlorobiphenyl PCB-Cl3 + + + + + + + + + + + + +
Tetrachlorobiphenyl PCB-Cl4 + + + + + + + + + +
Pentachlorobiphenyl PCB-Cl5 + + +
Hexachlorobiphenyl PCB-Cl6 + +
Heptachlorobiphenyl PCB-Cl7 + Chlorodiphenylmethane + + + + Trichlorodiphenylmethane + + + + Tetrachlorodiphenylmethane + Pentachlorodiphenylmethane + + + Chlorophenoxybenzene + + + + + + + Dichlorophenoxybenzene + + + + + Trichlorophenoxybenzene + + Tetrachlorophenoxybenzene Trichloromethoxydiphenyl ether + + + + + + + + + + + Tetrachloromethoxydiphenyl ether + + + Pentachloromethoxydiphenyl ether + + Dichlorodiphenylsulphide + + + + + + Trichlorodiphenylsulphide + + + + MDT + + + + + + + + + + + MDD + + + + + + + + + + MDE + + + + + + + + + MDMU +
123
Free extractable
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 S3 G1 G2 G4 ------SOIL2 SOIL1 - - - - - TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW 1,1-dichloro-2-(chlorophenyl)-2- + + + + + + + phenylethane 4-chloro-1-methoxy-2-(phenylmethyl)- + + + + + + benzene Dichlorodibenzo-p-dioxin + + + Dichlorodiphenylsulfone + + + + + + + Dichlorodiphenyl disulfone + + + Chlorotriphenylmethane + + + + Dichlorotriphenylmethane + + Trichlorotriphenylmethane + + + + + + + + + + + + Trichloro-methyl-triphenylmethane Dibromochloronitromethane 1-(3-chlorophenyl)-4-(3- + chloropropyl)piperazine Clobazam Tris(2-chloroisopropyl)phosphate + + + +
124
Released after alkaline hydrolysis
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Trichlorobutadiene Tetrachloroethane Tetrachlorobutadiene Pentachloroethane Pentachlorobutadiene Hexachloroethane Hexachlorobutadiene Hexachlorohexadiene Dichlorothiophene + Trichlorothiophene + Tetrachlorothiophene Tetrachlorocyclohexene TCCH Pentachlorocyclohexene PCCH Hexachlorocyclohexane HCH + + + + Dichlorobenzene + + + + + + + + + + + + + Trichlorobenzene + + + + + + + + + + + + Tetrachlorobenzene + + + + + + Pentachlorobenzene Hexachlorobenzene Chlorotoluene + + + + Dichlorotoluene + + Trichlorotoluene Tetrachlorotoluene Pentachlorotoluene Hexachlorotoluene Chloroxylene Dichloroxylene Trichloroxylene Pentachloroethylbenzene Bromochlorotoluene Dibromochlorotoluene Tribromochlorotoluene Dibromochlorobenzene Bromodichlorobenzene Bromo-trichlorobenzene Bromotetrachlorobenzene Bromopentachlorobenzene
125
Released after alkaline hydrolysis
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorostyrene + + + + + + Dichlorostyrene + + + + + Trichlorostyrene + + + Tetrachlorostyrene Pentachlorostyrene Hexachlorostyrene Heptachlorostyrene Octachlorostyrene Chlorophenol + + + + Dichlorophenol + + + + + + + + + Trichlorophenol + + Tetrachlorophenol + + + Chloromethylphenol + Dichloromethylphenol + + + + Trichloromethylphenol + Chlorodimethylphenol Dihlorodimethylphenol Trichlorodimethylphenol Ditert-butyl-chlorophenol Tert-butyl-dichlorophenol Chlorobenzaldehyde + + + + + + Dichlorobenzaldehyde Trichlorobenzaldehyde Chloroanisole Dichloroanisole + + + + Trichloroanisole + + + Tetrachloroanisole Pentachloroanisole Chloroveratrole + + + + + + + Dichloroveratrole + + + + + Trichloroveratrole Tetrachloroveratrole Chlorobenzoic acid + + + + + + + Trichlorobenzoic acid + + + Chloroaniline + + + Dichloroaniline + + + + + Trichloroaniline + + + + + Pentachloroaniline Dibromochloroaniline Bromodichloroaniline Tribromochloroaniline Dichloropyridinamine + + + + Dichloro-methoxy-pyridinamine + + + + + Dichloro-ortho-phenylenediamine Chlorobenzenethiol + + + + Dichlorobenzenethiol +
126
Released after alkaline hydrolysis
S2 S1 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorothioanisole + + + + Dichlorothioanisole + + + + Trichlorothioanisole Tetrachlorothioanisole Pentachlorothioanisole Chloromethyl phenyl sulfide + + + + Chloro-(chloromethyl)thio-benzene + + + + + Diochloro-(chloromethyl)thio-benzene + + + Chloronaphthalene + + Dichloronaphthalene + + + + Trichloronaphthalene Tetrachloronaphthalene Pentachloronaphthalene Hexachloronaphthalene Heptachloronaphthalene Octachloronaphthalene Chlorotrimethylnaphthalene Dichlorotrimethylnaphthalene Trichlorotrimethylnaphthalene
Chlorobiphenyl PCB-Cl1 + + + + + +
Dichlorobiphenyl PCB-Cl2 + + + + + +
Trichlorobiphenyl PCB-Cl3 + + + + +
Tetrachlorobiphenyl PCB-Cl4 + + +
Pentachlorobiphenyl PCB-Cl5
Hexachlorobiphenyl PCB-Cl6
Heptachlorobiphenyl PCB-Cl7 Chlorodiphenylmethane Trichlorodiphenylmethane Tetrachlorodiphenylmethane Pentachlorodiphenylmethane Chlorophenoxybenzene + Dichlorophenoxybenzene + + + + Trichlorophenoxybenzene Tetrachlorophenoxybenzene + + + + Trichloromethoxydiphenyl ether Tetrachloromethoxydiphenyl ether Pentachloromethoxydiphenyl ether Dichlorodiphenylsulphide Trichlorodiphenylsulphide MDT MDD MDE + MDMU + + +
127
Released after alkaline hydrolysis
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL2 SOIL1 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW 1,1-dichloro-2-(chlorophenyl)-2- phenylethane 4-chloro-1-methoxy-2-(phenylmethyl)- benzene Dichlorodibenzo-p-dioxin + Dichlorodiphenylsulfone Dichlorodiphenyl disulfone Chlorotriphenylmethane + Dichlorotriphenylmethane Trichlorotriphenylmethane + + Trichloro-methyl-triphenylmethane Dibromochloronitromethane 1-(3-chlorophenyl)-4-(3- + + + chloropropyl)piperazine Clobazam Tris(2-chloroisopropyl)phosphate
128
Released after BBr3 treatment
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Trichlorobutadiene Tetrachloroethane + + Tetrachlorobutadiene + Pentachloroethane + Pentachlorobutadiene + Hexachloroethane Hexachlorobutadiene Hexachlorohexadiene Dichlorothiophene + Trichlorothiophene Tetrachlorothiophene Tetrachlorocyclohexene TCCH Pentachlorocyclohexene PCCH Hexachlorocyclohexane HCH Dichlorobenzene + + Trichlorobenzene + + + Tetrachlorobenzene Pentachlorobenzene Hexachlorobenzene Chlorotoluene + Dichlorotoluene Trichlorotoluene Tetrachlorotoluene Pentachlorotoluene Hexachlorotoluene Chloroxylene Dichloroxylene Trichloroxylene Pentachloroethylbenzene Bromochlorotoluene + Dibromochlorotoluene + + Tribromochlorotoluene Dibromochlorobenzene + + + + + + + + + + Bromodichlorobenzene + + Bromo-trichlorobenzene + Bromotetrachlorobenzene Bromopentachlorobenzene
129
Released after BBr3 treatment
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorostyrene Dichlorostyrene Trichlorostyrene Tetrachlorostyrene Pentachlorostyrene Hexachlorostyrene Heptachlorostyrene Octachlorostyrene Chlorophenol Dichlorophenol Trichlorophenol Tetrachlorophenol Chloromethylphenol Dichloromethylphenol Trichloromethylphenol Chlorodimethylphenol Dihlorodimethylphenol Trichlorodimethylphenol Ditert-butyl-chlorophenol Tert-butyl-dichlorophenol Chlorobenzaldehyde + Dichlorobenzaldehyde Trichlorobenzaldehyde Chloroanisole Dichloroanisole + Trichloroanisole + + Tetrachloroanisole Pentachloroanisole Chloroveratrole Dichloroveratrole Trichloroveratrole Tetrachloroveratrole Chlorobenzoic acid + + + + + Trichlorobenzoic acid Chloroaniline Dichloroaniline Trichloroaniline Pentachloroaniline Dibromochloroaniline Bromodichloroaniline Tribromochloroaniline Dichloropyridinamine Dichloro-methoxy-pyridinamine Dichloro-ortho-phenylenediamine Chlorobenzenethiol Dichlorobenzenethiol
130
Released after BBr3 treatment
S2 S1 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorothioanisole Dichlorothioanisole Trichlorothioanisole Tetrachlorothioanisole Pentachlorothioanisole Chloromethyl phenyl sulfide Chloro-(chloromethyl)thio-benzene Diochloro-(chloromethyl)thio-benzene Chloronaphthalene Dichloronaphthalene Trichloronaphthalene Tetrachloronaphthalene Pentachloronaphthalene Hexachloronaphthalene Heptachloronaphthalene Octachloronaphthalene Chlorotrimethylnaphthalene Dichlorotrimethylnaphthalene Trichlorotrimethylnaphthalene
Chlorobiphenyl PCB-Cl1
Dichlorobiphenyl PCB-Cl2
Trichlorobiphenyl PCB-Cl3
Tetrachlorobiphenyl PCB-Cl4
Pentachlorobiphenyl PCB-Cl5
Hexachlorobiphenyl PCB-Cl6
Heptachlorobiphenyl PCB-Cl7 Chlorodiphenylmethane Trichlorodiphenylmethane Tetrachlorodiphenylmethane Pentachlorodiphenylmethane Chlorophenoxybenzene Dichlorophenoxybenzene Trichlorophenoxybenzene Tetrachlorophenoxybenzene Trichloromethoxydiphenyl ether Tetrachloromethoxydiphenyl ether Pentachloromethoxydiphenyl ether Dichlorodiphenylsulphide Trichlorodiphenylsulphide MDT MDD MDE MDMU
131
Released after BBr3 treatment
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL2 SOIL1 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW 1,1-dichloro-2-(chlorophenyl)-2- phenylethane 4-chloro-1-methoxy-2-(phenylmethyl)- benzene Dichlorodibenzo-p-dioxin Dichlorodiphenylsulfone Dichlorodiphenyl disulfone Chlorotriphenylmethane Dichlorotriphenylmethane Trichlorotriphenylmethane Trichloro-methyl-triphenylmethane Dibromochloronitromethane + + + + 1-(3-chlorophenyl)-4-(3- chloropropyl)piperazine Clobazam + + + + + + + + + Tris(2-chloroisopropyl)phosphate
132
Released after RuO4 oxidation
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Trichlorobutadiene Tetrachloroethane + + Tetrachlorobutadiene Pentachloroethane Pentachlorobutadiene Hexachloroethane + + + + Hexachlorobutadiene Hexachlorohexadiene Dichlorothiophene Trichlorothiophene Tetrachlorothiophene Tetrachlorocyclohexene TCCH Pentachlorocyclohexene PCCH Hexachlorocyclohexane HCH Dichlorobenzene + Trichlorobenzene + + Tetrachlorobenzene + Pentachlorobenzene Hexachlorobenzene Chlorotoluene + + + + + Dichlorotoluene Trichlorotoluene Tetrachlorotoluene Pentachlorotoluene Hexachlorotoluene Chloroxylene Dichloroxylene Trichloroxylene Pentachloroethylbenzene Bromochlorotoluene + + + Dibromochlorotoluene + + + + Tribromochlorotoluene + + + + Dibromochlorobenzene + + + + + Bromodichlorobenzene Bromo-trichlorobenzene + + + + Bromotetrachlorobenzene Bromopentachlorobenzene
133
Released after RuO4 oxidation
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorostyrene Dichlorostyrene Trichlorostyrene Tetrachlorostyrene Pentachlorostyrene Hexachlorostyrene Heptachlorostyrene Octachlorostyrene Chlorophenol Dichlorophenol Trichlorophenol Tetrachlorophenol Chloromethylphenol Dichloromethylphenol Trichloromethylphenol Chlorodimethylphenol Dihlorodimethylphenol Trichlorodimethylphenol Ditert-butyl-chlorophenol + + + + + Tert-butyl-dichlorophenol + + + + + Chlorobenzaldehyde Dichlorobenzaldehyde Trichlorobenzaldehyde Chloroanisole Dichloroanisole Trichloroanisole Tetrachloroanisole Pentachloroanisole Chloroveratrole Dichloroveratrole Trichloroveratrole Tetrachloroveratrole Chlorobenzoic acid Trichlorobenzoic acid Chloroaniline Dichloroaniline Trichloroaniline + Pentachloroaniline Dibromochloroaniline + + + + Bromodichloroaniline + + Tribromochloroaniline + Dichloropyridinamine Dichloro-methoxy-pyridinamine Dichloro-ortho-phenylenediamine Chlorobenzenethiol Dichlorobenzenethiol
134
Released after RuO4 oxidation
S2 S1 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorothioanisole Dichlorothioanisole Trichlorothioanisole Tetrachlorothioanisole Pentachlorothioanisole Chloromethyl phenyl sulfide Chloro-(chloromethyl)thio-benzene Diochloro-(chloromethyl)thio-benzene Chloronaphthalene Dichloronaphthalene Trichloronaphthalene Tetrachloronaphthalene Pentachloronaphthalene Hexachloronaphthalene Heptachloronaphthalene Octachloronaphthalene Chlorotrimethylnaphthalene Dichlorotrimethylnaphthalene Trichlorotrimethylnaphthalene
Chlorobiphenyl PCB-Cl1
Dichlorobiphenyl PCB-Cl2
Trichlorobiphenyl PCB-Cl3 + + + + + + + + +
Tetrachlorobiphenyl PCB-Cl4 + + + + + + + +
Pentachlorobiphenyl PCB-Cl5 + + + + + + + + + + +
Hexachlorobiphenyl PCB-Cl6 + + + + + + + + + + +
Heptachlorobiphenyl PCB-Cl7 + + + + + + + + Chlorodiphenylmethane Trichlorodiphenylmethane Tetrachlorodiphenylmethane Pentachlorodiphenylmethane Chlorophenoxybenzene Dichlorophenoxybenzene Trichlorophenoxybenzene Tetrachlorophenoxybenzene Trichloromethoxydiphenyl ether Tetrachloromethoxydiphenyl ether Pentachloromethoxydiphenyl ether Dichlorodiphenylsulphide + Trichlorodiphenylsulphide MDT MDD MDE MDMU
135
Released after RuO4 oxidation
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL2 SOIL1 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW 1,1-dichloro-2-(chlorophenyl)-2- phenylethane 4-chloro-1-methoxy-2-(phenylmethyl)- benzene Dichlorodibenzo-p-dioxin Dichlorodiphenylsulfone + Dichlorodiphenyl disulfone Chlorotriphenylmethane Dichlorotriphenylmethane Trichlorotriphenylmethane Trichloro-methyl-triphenylmethane Dibromochloronitromethane + + + 1-(3-chlorophenyl)-4-(3- chloropropyl)piperazine Clobazam + + + + + Tris(2-chloroisopropyl)phosphate
136
Released after TMAH thermochemolysis
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Trichlorobutadiene Tetrachloroethane + + + + + Tetrachlorobutadiene Pentachloroethane Pentachlorobutadiene Hexachloroethane Hexachlorobutadiene Hexachlorohexadiene Dichlorothiophene + Trichlorothiophene Tetrachlorothiophene + Tetrachlorocyclohexene TCCH Pentachlorocyclohexene PCCH Hexachlorocyclohexane HCH + Dichlorobenzene + + + + + + + + + + Trichlorobenzene + + + + + + + + + + Tetrachlorobenzene + + + + + + + + + Pentachlorobenzene + + + + Hexachlorobenzene + + + Chlorotoluene + + + + + + + + + + Dichlorotoluene + + + + + Trichlorotoluene + + + + + Tetrachlorotoluene Pentachlorotoluene + Hexachlorotoluene Chloroxylene + + + + + + Dichloroxylene + + Trichloroxylene Pentachloroethylbenzene Bromochlorotoluene Dibromochlorotoluene Tribromochlorotoluene Dibromochlorobenzene Bromodichlorobenzene Bromo-trichlorobenzene Bromotetrachlorobenzene Bromopentachlorobenzene
137
Released after TMAH thermochemolysis
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorostyrene + + + + + + + + Dichlorostyrene + + + + + Trichlorostyrene + + + + + + + Tetrachlorostyrene + Pentachlorostyrene Hexachlorostyrene Heptachlorostyrene Octachlorostyrene Chlorophenol Dichlorophenol Trichlorophenol Tetrachlorophenol Chloromethylphenol Dichloromethylphenol Trichloromethylphenol Chlorodimethylphenol + Dihlorodimethylphenol Trichlorodimethylphenol Ditert-butyl-chlorophenol Tert-butyl-dichlorophenol Chlorobenzaldehyde + Dichlorobenzaldehyde + + Trichlorobenzaldehyde + Chloroanisole + + + + + Dichloroanisole + + + + + + Trichloroanisole + + Tetrachloroanisole Pentachloroanisole Chloroveratrole Dichloroveratrole Trichloroveratrole Tetrachloroveratrole Chlorobenzoic acid + Trichlorobenzoic acid Chloroaniline Dichloroaniline + Trichloroaniline + + Pentachloroaniline Dibromochloroaniline Bromodichloroaniline Tribromochloroaniline Dichloropyridinamine + + + Dichloro-methoxy-pyridinamine Dichloro-ortho-phenylenediamine + + + Chlorobenzenethiol Dichlorobenzenethiol
138
Released after TMAH thermochemolysis
S2 S1 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL1 SOIL2 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW Chlorothioanisole + + + + + + + Dichlorothioanisole + + + + + + Trichlorothioanisole + + Tetrachlorothioanisole Pentachlorothioanisole Chloromethyl phenyl sulfide Chloro-(chloromethyl)thio-benzene Diochloro-(chloromethyl)thio-benzene Chloronaphthalene + + + + + Dichloronaphthalene + + + + + + Trichloronaphthalene Tetrachloronaphthalene Pentachloronaphthalene Hexachloronaphthalene Heptachloronaphthalene Octachloronaphthalene Chlorotrimethylnaphthalene + + + Dichlorotrimethylnaphthalene + + Trichlorotrimethylnaphthalene + +
Chlorobiphenyl PCB-Cl1 +
Dichlorobiphenyl PCB-Cl2 +
Trichlorobiphenyl PCB-Cl3
Tetrachlorobiphenyl PCB-Cl4
Pentachlorobiphenyl PCB-Cl5
Hexachlorobiphenyl PCB-Cl6
Heptachlorobiphenyl PCB-Cl7 Chlorodiphenylmethane + + + + + + Trichlorodiphenylmethane Tetrachlorodiphenylmethane Pentachlorodiphenylmethane Chlorophenoxybenzene + Dichlorophenoxybenzene + + Trichlorophenoxybenzene Tetrachlorophenoxybenzene Trichloromethoxydiphenyl ether Tetrachloromethoxydiphenyl ether Pentachloromethoxydiphenyl ether Dichlorodiphenylsulphide + Trichlorodiphenylsulphide MDT MDD MDE + MDMU
139
Released after TMAH thermochemolysis
S1 S2 - -
Compounds S1 S2 S3 S4 S5 S6 G1 G2 G3 G4 ------SOIL2 SOIL1 ------TC TC TC TC TC TC TC TC TC TC BFW BFW BFW BFW 1,1-dichloro-2-(chlorophenyl)-2- + phenylethane 4-chloro-1-methoxy-2-(phenylmethyl)- benzene Dichlorodibenzo-p-dioxin Dichlorodiphenylsulfone Dichlorodiphenyl disulfone Chlorotriphenylmethane Dichlorotriphenylmethane Trichlorotriphenylmethane Trichloro-methyl-triphenylmethane + + Dibromochloronitromethane 1-(3-chlorophenyl)-4-(3- + + chloropropyl)piperazine Clobazam Tris(2-chloroisopropyl)phosphate
140 Appendix B: The quantified data of AOHs (for chapter 3)
Free extractable
S2 S5 S1 S1 S3 S4 S6 S2 S3 G4 G1 G2 ------TC TC TC Compounds TC TC TC TC SOIL1 SOIL2 TC TC TC - - BFW BFW BFW BFW Dichlorothiophene ------< 0.01 - - 0.02 - - - < 0.01 Trichlorothiophene 0.02 - - - - - < 0.01 0.02 - - - - 0.06 < 0.01 Tetrachlorothiophene 0.10 - - - - - 0.01 0.07 < 0.01 < 0.01 < 0.01 0.02 0.52 0.01 Tetrachlorocyclohexene 1.8 < 0.01 < 0.01 ------TCCH Pentachlorocyclohexene 4.7 0.01 0.01 < 0.01 < 0.01 0.01 - - 0.03 - < 0.01 0.04 0.27 0.06 PCCH Hexachlorocyclohexane 1900 11 12 0.07 0.11 0.58 0.41 0.99 1.6 0.25 1.6 5.7 45 11 HCH Dichlorobenzene 3.2 0.79 8.7 0.24 0.44 0.16 0.31 3.1 0.51 0.28 0.31 0.74 3.1 0.09 Trichlorobenzene 2.2 0.08 0.27 0.05 0.05 0.05 0.04 1.0 1.3 0.89 0.90 7.1 5.8 0.15 Tetrachlorobenzene 0.88 < 0.01 - < 0.01 < 0.01 0.01 0.02 0.57 0.54 0.52 0.35 7.4 5.1 0.14 Pentachlorobenzene ------0.01 0.07 0.05 0.04 0.02 0.21 0.97 0.03 Hexachlorobenzene ------0.07 0.11 - - - - 2.1 0.13 Dibromochlorobenzene ------Chlorostyrene - - 0.13 - < 0.01 < 0.01 - - 0.13 0.27 0.42 0.53 - - Dichlorostyrene 0.21 1.3 5.0 0.16 0.19 0.38 0.05 0.03 0.31 0.38 0.52 2.5 - < 0.01 Trichlorostyrene 11 - 0.35 0.07 0.10 0.25 0.18 0.22 1.9 1.7 2.3 7.3 0.81 0.05 Tetrachlorostyrene 5.0 - - - - - 0.03 0.01 0.18 0.18 0.11 1.8 0.12 < 0.01 Pentachlorostyrene ------0.01 ------< 0.01 Hexachlorostyrene ------0.02 < 0.01 Heptachlorostyrene ------0.02 - - - - 0.16 0.01 Octachlorostyrene ------< 0.01 - - - - 0.06 0.01 Chlorophenol 0.18 0.01 0.18 - - - - - 0.09 0.19 0.23 - - - Dichlorophenol - 0.08 0.34 - - - - 0.09 0.04 0.31 0.27 0.35 0.91 0.01 Trichlorophenol 0.12 < 0.01 - - - - - 0.02 0.02 0.14 0.13 0.26 0.19 - Tetrachlorophenol - < 0.01 ------Ditert-butyl-chlorophenol ------Chlorobenzaldehyde 1.8 0.10 0.32 - 0.02 0.03 - - - 0.17 0.31 0.69 - - Dichlorobenzaldehyde ------Trichlorobenzaldehyde ------
141 Free extractable
S3 S1 S2 S3 S4 S5 S6 S1 S2 G1 G2 G4 ------TC TC TC Compounds TC TC TC TC SOIL1 SOIL2 TC TC TC - - BFW BFW BFW BFW Chloroanisole 0.73 0.05 0.04 0.16 0.06 0.87 - 0.41 - - - 0.14 0.18 - Dichloroanisole 0.28 - 0.13 0.04 0.02 0.24 0.02 0.11 0.04 0.05 0.10 0.53 15 2.4 Trichloroanisole 2.6 0.02 0.11 - - - < 0.01 0.60 0.08 0.12 0.07 0.73 7.1 1.3 Tetrachloroanisole 0.08 ------0.08 - - - 0.02 1.9 0.28 Pentachloroanisole ------0.10 - - - - 3.5 0.31 Chloroveratrole ------0.17 - - - - - 0.04 - Dichloroveratrole - - - 0.12 ------0.22 - Trichloroveratrole 0.03 ------0.07 < 0.01 Tetrachloroveratrole ------< 0.01 - - - - 0.11 0.01 Chlorobenzoic acid 4.0 0.05 0.21 - 0.02 - - 0.24 - 0.39 0.23 1.0 0.46 - Trichlorobenzoic acid ------Chloroaniline ------0.02 - - - - 0.07 0.02 < < Dichloroaniline - - - 0.02 0.03 0.16 - - - < 0.01 - - 0.01 0.01 Trichloroaniline ------Pentachloroaniline ------< 0.01 < 0.01 < Dichloropyridinamine ------0.34 - 0.14 - - - - 0.01 Dichloro-methoxy-pyridinamine ------0.02 0.32 0.17 0.18 - - Chlorobenzenethiol ------0.06 ------Dichlorobenzenethiol ------Chlorothioanisole ------0.04 0.06 1.1 1.2 0.12 - - Dichlorothioanisole ------< 0.01 - 1.1 2.6 3.2 1.3 - - Trichlorothioanisole ------0.18 0.82 0.33 0.41 - - Tetrachlorothioanisole ------0.11 0.07 0.10 0.25 - - Pentachlorothioanisole ------0.02 0.01 0.03 0.05 - - Chloromethyl phenyl sulfide ------Chloro-(chloromethyl)thio-benzene ------< 0.01 Diochloro-(chloromethyl)thio-benzene ------< 0.01 Chloronaphthalene - 0.24 0.54 0.12 0.21 0.48 ------Dichloronaphthalene - 0.13 0.28 0.09 0.14 0.33 0.03 0.25 < 0.01 - - 0.02 1.5 0.05 Trichloronaphthalene - - - - 0.03 0.02 - - 0.14 0.02 0.02 0.03 0.58 0.03 Tetrachloronaphthalene ------0.02 - - - - 0.53 0.04 Pentachloronaphthalene ------< 0.01 - - - - - 0.02 Hexachloronaphthalene ------< 0.01 Heptachloronaphthalene ------< 0.01 Octachloronaphthalene ------< 0.01 Chlorotrimethylnaphthalene ------Dichlorotrimethylnaphthalene ------< 0.01 - Trichlorotrimethylnaphthalene ------
142 Free extractable
S1 S2 S3 S4 S5 S6 S1 S2 S3 G1 G2 G4 ------TC TC TC TC TC TC TC SOIL1 SOIL2
Compounds TC TC TC - - BFW BFW BFW BFW
Chlorobiphenyl PCB-Cl1 0.03 0.01 0.02 0.10 < 0.01 0.16 0.15 0.18 0.14 0.07 0.46 0.04
Dichlorobiphenyl PCB-Cl2 - - - - - 0.03 0.05 0.27 0.03 0.02 0.02 0.05 0.88 0.05
Trichlorobiphenyl PCB-Cl3 - 0.11 0.55 < 0.01 0.03 0.15 0.04 0.08 < 0.01 < 0.01 < 0.01 0.01 0.18 0.01
Tetrachlorobiphenyl PCB-Cl4 - 0.06 0.22 < 0.01 0.01 0.04 - 0.04 < 0.01 - - - 0.20 0.01
Pentachlorobiphenyl PCB-Cl5 ------0.02 - - - - 0.08 1.10
Hexachlorobiphenyl PCB-Cl6 ------0.14 0.36
Heptachlorobiphenyl PCB-Cl7 ------< 0.01 Chlorophenoxybenzene - 0.01 0.02 - - 0.02 - - 0.03 0.02 0.02 0.02 - - Dichlorophenoxybenzene ------0.02 0.01 0.01 < 0.01 0.18 < 0.01 Trichlorophenoxybenzene ------0.02 < 0.01 Tetrachlorophenoxybenzene ------
Trichloromethoxydiphenyl ether - 0.02 0.02 0.06 0.03 0.59 - 2.2 - 0.01 < 0.01 0.01 0.15 0.01 Tetrachloromethoxydiphenyl ------1.7 - - - - 0.65 0.02 ether Pentachloromethoxydiphenyl ------0.35 - - - - 0.26 - ether Dichlorodiphenylsulphide 0.04 ------0.29 0.03 0.03 0.03 0.03 0.60 0.01 Trichlorodiphenylsulphide - - - - < 0.01 - - - 0.01 0.01 0.02 < 0.01 - - MDT 60 1.2 1.0 < 0.01 0.03 0.25 < 0.01 4.3 - - - - 4.2 0.03 MDD 0.97 0.05 0.06 < 0.01 < 0.01 0.01 - 0.73 - - - - 0.01 < 0.01 MDMU 0.24 ------
Chlorotriphenylmethane - - - 0.06 - 0.80 - 0.04 - - - - 0.16 - Dichlorotriphenylmethane ------0.04 - - - - 0.26 - Trichlorotriphenylmethane 11 0.90 2.1 0.01 0.02 0.26 - 0.12 0.17 0.16 0.24 0.15 0.14 - Trichloro-methyl------triphenylmethane 1,1-dichloro-2-(chlorophenyl)-2- 97 - 0.46 - - - - - 0.41 0.11 0.28 0.29 0.17 - phenylethane Dichlorodiphenylsulfone 15 0.56 1.3 ------2.0 0.13 Clobazam ------
143
Released after alkaline hydrolysis
S1 S2 S3 S4 S5 S6 S1 S2 G1 G2 G3 G4 ------Compounds TC TC TC TC TC TC SOIL2 SOIL1 TC TC TC TC - - BFW BFW BFW BFW Dichlorothiophene ------< 0.01 - - - Trichlorothiophene - < 0.01 ------Tetrachlorothiophene ------Tetrachlorocyclohexene ------TCCH Pentachlorocyclohexene ------PCCH Hexachlorocyclohexane ------0.37 - 1.7 - - - 0.16 0.05 HCH Dichlorobenzene 0.09 22 8.2 - 0.50 4.2 0.68 2.2 1.7 2.1 0.13 0.61 0.18 0.21 Trichlorobenzene 0.09 4.1 1.9 - 0.11 0.96 0.08 0.60 1.3 2.1 0.49 3.1 - 0.04 Tetrachlorobenzene - 0.03 0.01 - - - - - 0.04 0.02 0.01 0.08 - - Pentachlorobenzene ------Hexachlorobenzene ------Dibromochlorobenzene ------Chlorostyrene - 1.8 0.48 - - - - - 1.7 14 0.79 1.3 - - Dichlorostyrene - 1.3 0.46 ------0.05 < 0.01 0.06 - - Trichlorostyrene - 0.14 0.03 - - - - - 0.04 - - - - - Tetrachlorostyrene ------Pentachlorostyrene ------Hexachlorostyrene ------Heptachlorostyrene ------Octachlorostyrene ------Chlorophenol 0.30 0.42 ------3.6 - - - 0.23 - Dichlorophenol 0.37 0.49 - - - - 0.38 1.8 6.3 0.27 0.45 1.5 10 0.20 Trichlorophenol 0.12 ------4.4 - Tetrachlorophenol 0.01 ------0.02 0.50 - Ditert-butyl-chlorophenol ------Chlorobenzaldehyde 0.19 1.50 0.92 ------0.46 0.24 2.00 - - Dichlorobenzaldehyde ------Trichlorobenzaldehyde ------
144 Released after alkaline hydrolysis
- - S1 S2 S3 S4 S5 S6 S1 S2 G1 G2 G3 G4 ------Compounds - - BFW BFW TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC BFW BFW Chloroanisole - 1.1 ------Dichloroanisole 0.49 ------0.15 - - - - - Trichloroanisole 0.04 ------0.13 0.25 - Tetrachloroanisole ------Pentachloroanisole ------Chloroveratrole 0.04 3.6 7.1 12 6.8 46 ------0.28 - Dichloroveratrole - 1.0 0.23 0.36 0.27 0.48 ------Trichloroveratrole ------Tetrachloroveratrole ------Chlorobenzoic acid 4.6 1.6 1.1 - - - - 0.18 - 3.3 4.3 15 - - Trichlorobenzoic acid ------2.5 5.1 13 - - Chloroaniline ------0.47 - - - - 0.40 - Dichloroaniline ------0.15 0.03 0.45 0.10 0.13 Trichloroaniline ------0.10 0.09 0.01 0.06 1.6 - Pentachloroaniline ------Dichloropyridinamine - 6.0 1.4 ------2.1 - 1.1 - - Dichloro-methoxy-pyridinamine ------1.9 0.78 0.95 0.37 0.29 - Chlorobenzenethiol ------1.8 1.2 6.8 1.1 - Dichlorobenzenethiol ------1.7 - - Chlorothioanisole ------0.39 0.10 0.05 0.36 - - Dichlorothioanisole ------0.49 0.04 0.05 0.18 - - Trichlorothioanisole ------Tetrachlorothioanisole ------Pentachlorothioanisole ------Chloromethyl phenyl sulfide ------0.17 0.21 < 0.01 1.2 - - Chloro-(chloromethyl)thio-benzene ------0.31 - Diochloro-(chloromethyl)thio-benzene ------Chloronaphthalene - 1.2 0.76 ------Dichloronaphthalene - 0.61 0.40 - 0.33 0.32 ------Trichloronaphthalene ------Tetrachloronaphthalene ------Pentachloronaphthalene ------Hexachloronaphthalene ------Heptachloronaphthalene ------Octachloronaphthalene ------Chlorotrimethylnaphthalene ------Dichlorotrimethylnaphthalene ------Trichlorotrimethylnaphthalene ------
145 Released after alkaline hydrolysis
S1 S2 S3 S4 S5 S6 S1 S2 G2 G3 G1 G4 ------
Compounds TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW
Chlorobiphenyl PCB-Cl1 - 0.11 0.09 - - - - - 0.41 0.04 0.03 0.00 - -
Dichlorobiphenyl PCB-Cl2 - 0.01 0.06 ------0.02 0.02 0.15 0.06 -
Trichlorobiphenyl PCB-Cl3 - < 0.01 0.04 ------< 0.01 < 0.01 0.01 - -
Tetrachlorobiphenyl PCB-Cl4 ------< 0.01 < 0.01 0.06 - -
Pentachlorobiphenyl PCB-Cl5 ------
Hexachlorobiphenyl PCB-Cl6 ------
Heptachlorobiphenyl PCB-Cl7 ------Chlorophenoxybenzene ------0.39 - - - - - Dichlorophenoxybenzene ------0.44 0.01 < 0.01 0.03 - - Trichlorophenoxybenzene ------Tetrachlorophenoxybenzene - - 0.02 ------0.02 0.04 0.36 - - Trichloromethoxydiphenyl ether ------Tetrachloromethoxydiphenyl ether ------Pentachloromethoxydiphenyl ether ------Dichlorodiphenylsulphide ------Trichlorodiphenylsulphide ------MDT ------MDD ------MDMU 0.17 0.20 0.18 ------Chlorotriphenylmethane 0.84 ------Dichlorotriphenylmethane ------Trichlorotriphenylmethane 0.08 - 0.04 ------Trichloro-methyl-triphenylmethane ------1,1-dichloro-2-(chlorophenyl)-2------phenylethane Dichlorodiphenylsulfone ------Clobazam ------
146 Released after BBr3 treatment
S1 S2 S3 S4 S5 S6 S1 S2 G2 G1 G3 G4 ------
Compounds TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW Dichlorothiophene ------0.02 - - Trichlorothiophene ------Tetrachlorothiophene ------Tetrachlorocyclohexene ------TCCH Pentachlorocyclohexene ------PCCH Hexachlorocyclohexane ------HCH Dichlorobenzene < 0.01 0.32 ------Trichlorobenzene < 0.01 0.01 ------0.06 - Tetrachlorobenzene ------Pentachlorobenzene ------Hexachlorobenzene ------Dibromochlorobenzene - 0.02 1.1 < 0.01 0.02 0.63 0.01 - - 0.42 < 0.01 0.03 - < 0.01 Chlorostyrene ------Dichlorostyrene ------Trichlorostyrene ------Tetrachlorostyrene ------Pentachlorostyrene ------Hexachlorostyrene ------Heptachlorostyrene ------Octachlorostyrene ------Chlorophenol ------Dichlorophenol ------Trichlorophenol ------Tetrachlorophenol ------Ditert-butyl-chlorophenol ------Chlorobenzaldehyde ------4.8 - - Dichlorobenzaldehyde ------Trichlorobenzaldehyde ------
147 Released after BBr3 treatment
S1 S2 S3 S4 S5 S6 S1 S2 G1 G2 G3 G4 ------
Compounds TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW Chloroanisole ------Dichloroanisole ------Trichloroanisole - - - 0.28 ------0.40 - Tetrachloroanisole ------Pentachloroanisole ------Chloroveratrole ------Dichloroveratrole ------Trichloroveratrole ------Tetrachloroveratrole ------Chlorobenzoic acid - - - 0.07 - - - - - 1.8 1.2 3.8 0.91 - Trichlorobenzoic acid ------Chloroaniline ------Dichloroaniline ------Trichloroaniline ------Pentachloroaniline ------Dichloropyridinamine ------Dichloro-methoxy-pyridinamine ------Chlorobenzenethiol ------Dichlorobenzenethiol ------Chlorothioanisole ------Dichlorothioanisole ------Trichlorothioanisole ------Tetrachlorothioanisole ------Pentachlorothioanisole ------Chloromethyl phenyl sulfide ------Chloro-(chloromethyl)thio-benzene ------Diochloro-(chloromethyl)thio-benzene ------Chloronaphthalene ------Dichloronaphthalene ------Trichloronaphthalene ------Tetrachloronaphthalene ------Pentachloronaphthalene ------Hexachloronaphthalene ------Heptachloronaphthalene ------Octachloronaphthalene ------Chlorotrimethylnaphthalene ------Dichlorotrimethylnaphthalene ------Trichlorotrimethylnaphthalene ------
148 Released after BBr3 treatment
S3 S5 S6 S1 S2 S4 S1 S2 G1 G2 G3 G4 ------Compounds TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW
Chlorobiphenyl PCB-Cl1 ------
Dichlorobiphenyl PCB-Cl2 ------
Trichlorobiphenyl PCB-Cl3 ------
Tetrachlorobiphenyl PCB-Cl4 ------
Pentachlorobiphenyl PCB-Cl5 ------
Hexachlorobiphenyl PCB-Cl6 ------
Heptachlorobiphenyl PCB-Cl7 ------Chlorophenoxybenzene ------Dichlorophenoxybenzene ------Trichlorophenoxybenzene ------Tetrachlorophenoxybenzene ------Trichloromethoxydiphenyl ether ------Tetrachloromethoxydiphenyl ether ------Pentachloromethoxydiphenyl ether ------Dichlorodiphenylsulphide ------Trichlorodiphenylsulphide ------MDT ------MDD ------MDMU ------Chlorotriphenylmethane ------Dichlorotriphenylmethane ------Trichlorotriphenylmethane ------Trichloro-methyl-triphenylmethane ------1,1-dichloro-2-(chlorophenyl)-2------phenylethane Dichlorodiphenylsulfone ------Clobazam - 66 1.2 0.52 2.0 - - - 0.52 - 0.13 1.4 0.05 0.21
149 Released after RuO4 oxidation
S4 S5 S6 S1 S2 S1 S2 S3 G1 G2 G3 G4 ------Compounds TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW Dichlorothiophene ------Trichlorothiophene ------Tetrachlorothiophene ------Tetrachlorocyclohexene ------TCCH Pentachlorocyclohexene ------PCCH Hexachlorocyclohexane ------HCH Dichlorobenzene ------Trichlorobenzene ------0.06 0.02 - - Tetrachlorobenzene ------< 0.01 - - Pentachlorobenzene ------Hexachlorobenzene ------Dibromochlorobenzene - - - < 0.01 < 0.01 - - - < 0.01 < 0.01 - 0.11 < 0.01 - Chlorostyrene ------Dichlorostyrene ------Trichlorostyrene ------Tetrachlorostyrene ------Pentachlorostyrene ------Hexachlorostyrene ------Heptachlorostyrene ------Octachlorostyrene ------Chlorophenol ------Dichlorophenol ------Trichlorophenol ------Tetrachlorophenol ------Ditert-butyl-chlorophenol 3.9 - 0.45 0.05 0.73 ------0.01 - Chlorobenzaldehyde ------Dichlorobenzaldehyde ------Trichlorobenzaldehyde ------
150 Released after RuO4 oxidation
S1 S2 S3 S4 S5 S6 S1 S2 G2 G3 G4 G1 ------Compounds - - - - TC TC TC TC TC TC SOIL2 SOIL1 TC TC TC TC - - BFW BFW BFW BFW Chloroanisole ------Dichloroanisole ------Trichloroanisole ------Tetrachloroanisole ------Pentachloroanisole ------Chloroveratrole ------Dichloroveratrole ------Trichloroveratrole ------Tetrachloroveratrole ------Chlorobenzoic acid ------Trichlorobenzoic acid ------Chloroaniline ------Dichloroaniline ------Trichloroaniline ------0.05 - - Pentachloroaniline ------Dichloropyridinamine ------Dichloro-methoxy-pyridinamine ------Chlorobenzenethiol ------Dichlorobenzenethiol ------Chlorothioanisole ------Dichlorothioanisole ------Trichlorothioanisole ------Tetrachlorothioanisole ------Pentachlorothioanisole ------Chloromethyl phenyl sulfide ------Chloro-(chloromethyl)thio-benzene ------Diochloro-(chloromethyl)thio-benzene ------Chloronaphthalene ------Dichloronaphthalene ------Trichloronaphthalene ------Tetrachloronaphthalene ------Pentachloronaphthalene ------Hexachloronaphthalene ------Heptachloronaphthalene ------Octachloronaphthalene ------Chlorotrimethylnaphthalene ------Dichlorotrimethylnaphthalene ------Trichlorotrimethylnaphthalene ------
151 Released after RuO4 oxidation
S1 S2 S3 S4 S5 S6 S1 S2 G1 G3 G2 G4 ------
Compounds TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW
Chlorobiphenyl PCB-Cl1 ------
Dichlorobiphenyl PCB-Cl2 ------
Trichlorobiphenyl PCB-Cl3 0.06 0.03 - 0.03 0.02 0.04 - - 0.02 0.01 - 0.02 0.05 -
Tetrachlorobiphenyl PCB-Cl4 0.11 0.04 - 0.02 0.01 - - 0.01 0.01 - 0.03 0.10 -
Pentachlorobiphenyl PCB-Cl5 0.33 0.15 - 0.13 0.14 0.34 - - 0.07 0.17 0.02 0.16 0.36 0.08
Hexachlorobiphenyl PCB-Cl6 0.77 0.28 - 0.25 0.31 0.61 - - 0.14 0.18 0.03 0.44 0.76 0.67
Heptachlorobiphenyl PCB-Cl7 - - - 0.03 0.02 - - - 0.01 < 0.01 - 0.02 0.22 0.25 Chlorophenoxybenzene ------Dichlorophenoxybenzene ------Trichlorophenoxybenzene ------Tetrachlorophenoxybenzene ------Trichloromethoxydiphenyl ether ------Tetrachloromethoxydiphenyl ether ------Pentachloromethoxydiphenyl ether ------Dichlorodiphenylsulphide ------0.22 - Trichlorodiphenylsulphide ------MDT ------MDD ------MDMU ------Chlorotriphenylmethane ------Dichlorotriphenylmethane - - - - 0.61 - - - - 0.13 - - - - Trichlorotriphenylmethane ------Trichloro-methyl-triphenylmethane ------1,1-dichloro-2-(chlorophenyl)-2------phenylethane Dichlorodiphenylsulfone ------3.9 - Clobazam - - - - 1.9 - - - 0.35 0.03 - - 1.0 0.40
152 Released after TMAH thermochemolysis
S1 S2 S3 S4 S5 S6 S1 S2 G1 G2 G3 G4 ------Compounds - - - - TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW Dichlorothiophene - 0.97 ------Trichlorothiophene ------Tetrachlorothiophene ------0.06 - Tetrachlorocyclohexene ------TCCH Pentachlorocyclohexene ------PCCH Hexachlorocyclohexane 0.04 ------HCH Dichlorobenzene 0.27 25 34 0.26 0.17 0.04 - - 0.06 < 0.01 0.13 0.67 - - Trichlorobenzene 3.7 1.1 7.9 0.08 0.03 < 0.01 - - 0.04 - 0.03 0.18 0.08 - Tetrachlorobenzene 0.06 0.15 0.24 < 0.01 < 0.01 - - - - - < 0.01 0.03 0.09 < 0.01 Pentachlorobenzene ------0.01 - - - - 0.11 < 0.01 Hexachlorobenzene ------0.02 - - - - 0.41 0.01 Dibromochlorobenzene ------Chlorostyrene - 2.1 6.5 0.41 0.01 - - - 1.5 0.20 3.7 16 - - Dichlorostyrene - 1.3 3.0 0.05 0.02 - - - - - 0.04 - - - Trichlorostyrene 0.07 0.29 0.68 0.01 0.01 - - - - - 0.02 0.10 - - Tetrachlorostyrene 0.17 ------Pentachlorostyrene ------Hexachlorostyrene ------Heptachlorostyrene ------Octachlorostyrene ------Chlorophenol ------Dichlorophenol ------Trichlorophenol ------Tetrachlorophenol ------Ditert-butyl-chlorophenol ------Chlorobenzaldehyde ------0.38 ------Dichlorobenzaldehyde ------0.15 - - - - 0.10 - Trichlorobenzaldehyde ------0.13 -
153 Released after TMAH thermochemolysis
S1 S2 S3 S4 S5 S6 S1 S2 G1 G2 G3 G4 ------Compounds - - - - TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW Chloroanisole 0.38 - - 0.21 - - - - 0.88 - - 0.11 - - Dichloroanisole 0.93 - - 0.03 - - - - 1.5 - - 0.07 - 0.15 Trichloroanisole 0.23 ------0.03 - - - - - Tetrachloroanisole ------Pentachloroanisole ------Chloroveratrole ------Dichloroveratrole ------Trichloroveratrole ------Tetrachloroveratrole ------Chlorobenzoic acid 3.7 ------Trichlorobenzoic acid ------Chloroaniline ------Dichloroaniline ------1.5 - - Trichloroaniline ------0.15 0.17 - Pentachloroaniline ------Dichloropyridinamine - - 8.3 - - 0.68 - - - - 1.2 - - - Dichloro-methoxy-pyridinamine ------Chlorobenzenethiol ------Dichlorobenzenethiol ------Chlorothioanisole - 5.5 9.0 - - - - - 8.4 0.28 22 41 - 0.04 Dichlorothioanisole - 3.3 5.8 - - - - - 1.3 0.11 4.8 11 - - Trichlorothioanisole ------0.19 0.45 - - Tetrachlorothioanisole ------Pentachlorothioanisole ------Chloromethyl phenyl sulfide ------Chloro-(chloromethyl)thio-benzene ------Diochloro-(chloromethyl)thio-benzene ------Chloronaphthalene - 2.1 3.6 0.05 0.04 0.19 ------Dichloronaphthalene - 0.77 1.0 0.01 0.01 0.05 ------0.08 - Trichloronaphthalene ------Tetrachloronaphthalene ------Pentachloronaphthalene ------Hexachloronaphthalene ------Heptachloronaphthalene ------Octachloronaphthalene ------Chlorotrimethylnaphthalene ------0.96 2.0 0.01 - Dichlorotrimethylnaphthalene ------0.05 0.13 - - Trichlorotrimethylnaphthalene ------0.02 0.02 - -
154 Released after TMAH thermochemolysis
S1 S2 S3 S4 S5 S6 S1 S2 G1 G2 G3 G4 ------Compounds TC TC TC TC TC TC SOIL1 SOIL2 TC TC TC TC - - BFW BFW BFW BFW
Chlorobiphenyl PCB-Cl1 ------
Dichlorobiphenyl PCB-Cl2 ------0.05 -
Trichlorobiphenyl PCB-Cl3 ------
Tetrachlorobiphenyl PCB-Cl4 ------
Pentachlorobiphenyl PCB-Cl5 ------
Hexachlorobiphenyl PCB-Cl6 ------
Heptachlorobiphenyl PCB-Cl7 ------Chlorophenoxybenzene ------0.13 - - Dichlorophenoxybenzene ------0.06 0.07 - Trichlorophenoxybenzene ------Tetrachlorophenoxybenzene ------Trichloromethoxydiphenyl ether ------Tetrachloromethoxydiphenyl ether ------Pentachloromethoxydiphenyl ether ------Dichlorodiphenylsulphide ------0.04 ------Trichlorodiphenylsulphide ------MDT ------MDD ------MDMU ------Chlorotriphenylmethane ------Dichlorotriphenylmethane ------Trichlorotriphenylmethane ------Trichloro-methyl-triphenylmethane ------0.23 0.14 - - 1,1-dichloro-2-(chlorophenyl)-2------0.15 - - phenylethane Dichlorodiphenylsulfone ------Clobazam ------
155