Vrije Universiteit Brussel

Development of the Diffusive Gradient in Thin-Films (DGT) passive sampling technique for Platinum Group Elements (PGEs) and its application in urban rivers Abdulbur Alfakhory, Ehab

Publication date: 2021

Document Version: Final published version

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Citation for published version (APA): Abdulbur Alfakhory, E. (2021). Development of the Diffusive Gradient in Thin-Films (DGT) passive sampling technique for Platinum Group Elements (PGEs) and its application in urban rivers.

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Download date: 03. Oct. 2021

Faculteit wetenschappen en bio-ingenieurswetenschap Analytical, Environmental & Geo-chemistry department (AMGC)

Development of the Diffusive Gradient in Thin-Films (DGT) passive sampling technique for Platinum Group Elements (PGEs) and its application in urban rivers.

PhD dissertation presented to obtain the degree of Doctor of Science from the Vrije Universiteit Brussel by

Ehab Abdulbur-Alfakhoury Promotor: Prof. Dr. Martine Leermakers

Gabriel Billon Professor, Université de Lille Examiner Pavel Divis Professor, Brno University of Technology Examiner Steven Goderis Professor, Vrije Universiteit Brussel Chair Yue Gao Professor, Vrije Universiteit Brussel Secretaries Frederik Tielens Professor, Vrije Universiteit Brussel Examiner Willy Baeyens Professor, Vrije Universiteit Brussel Examiner Harry Olde Venterink Professor, Vrije Universiteit Brussel Examiner Martine Leermakers Professor, Vrije Universiteit Brussel Promoter

June 2021

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Abstract of the PhD research The Platinum Group elements (PGEs) Pt, Pd and Rh are increasingly used in modern day society in industrial and medical applications, resulting in increases in their concentrations in the environment. However, their environmental behaviour, fate and impact are still widely unknown. The accurate determination of PGEs at environmentally relevant concentrations is still a challenge for analytical chemists. Sensitive and interference-free analytical methods are required for measuring the very low levels of Pt, Pd, and Rh in complex matrices. This involves preconcentration of the elements from the matrix and separation from the interfering elements. The aim of this PhD study was to thus to develop, validate, and apply the in situ preconcentration technique Diffusive Gradients in Thin-Films (DGT) for Pt, Pd and Rh speciation in surface water.

In the first step, the performance of the DGT for PGEs was tested by using various chelating resins: Purolite S914, Purolite S920, Purolite S985, MPX-317 and MP-102 resins. The influence of parameters such as pH, ionic strength, effective capacity and organic matter was explored. The precise quantitation of these elements by High Resolution Inductively Coupled Plasma Mass Spectrometry (HR-ICP-MS) was achieved (Abdulbur-Alfakhoury E. et al., Talanta 203 (2019): 34-48). The method was further improved by developing selective leaching procedures to separate the PGEs from interfering elements accumulating on the resin gels (Abdulbur-Alfakhoury E. and M. Leermakers, Talanta 223 (2020): 121771).

In addition, the influence of ageing of solutions containing PGEs on the DGT response was evaluated showing that inert species of Rh can be formed and that organic matter plays an important role in forming DGT labile species from the inert form (Abdulbur-Alfakhoury E. and M. Leermakers, Journal of Analytical Atomic Spectrometry, 2021. DOI: 10.1039/D0JA00442A). Subsequently, the DGT technique was validated by several in situ DGT field trials which were carried out in both urban rivers and a hospital effluent. A first interlaboratory comparison was carried out between VUB_AMGC and LILLE_LASIRE showing a good agreement for Pt and demonstrating the need for evaluation of the DGT operational parameters in further studies (Abdulbur-Alfakhoury, E., et al. (2021). Science of the Total Environment 784). The combination of the laboratory experiments and field studies allows obtaining a better

2 understanding the new emerging pollutants of PGEs in environments heavily impacted by human activities, making it possible to trace their sources, monitor their evolution and evaluate their potential impact in the future. Nederlandstalige abstract De platina groep elemeten (PGEs) Pt, Pd en Rhodium worden recent veel gebruikt voor medische en technologische toepassingen en worden daarom vaak als contaminanten in het milieu aangetroffen, maar hun gedrag, lot en de impact op het milieu is nog grotendeels onbekend. De nauwkeurige bepaling van PGE's bij milieurelevante concentraties is nog steeds een uitdaging voor de analytische chemie. Gevoelige en interferentievrije analytische methoden zijn vereist voor het meten van de zeer lage concentraties van Pt, Pd en Rh in complexe matrices. Dit omvat preconcentratie van de elementen uit de matrix en scheiding van de interferende elementen. Het doel van deze doctoraatsstudie is de ontwikkeling, validatie en toepassing van de in-situ preconcentratietechniek Diffusive Gradients in Thin-Films (DGT) voor Pt-, Pd- en Rh in oppervlaktewater. In de eerste fase werd een uitgebreid laboratorium onderzoek uitgevoerd om de techniek te ontwikkelen waarbij verschillende chelaterende harsen vergeleken werden: Purolite S914, Purolite S920, Purolite S985, MPX-317 en MP-102. De invloed van parameters als pH, ionische sterkte, capaciteit en organisch materiaal werd onderzocht. De nauwkeurige kwantificering van deze elementen door middel van inductief gekoppelde plasmamassaspectrometrie met hoge resolutie (HR-ICP-MS) (Abdulbur-Alfakhoury E. et al., Talanta 203 (2019): 34-48). De methode werd nog verbeterd door de ontwikkeling van selectieve extracties te ontwikkelen om interfererende elementen te scheiden van de PGEs (Abdulbur-Alfakhoury E. and M. Leermakers, Talanta 223 (2020): 121771). Vervolgens werd het verouderingseffect van oplossingen bestudeerd waarbij werd aangetoond dat inerte Rh vormen kunnen ontstaan en de aanwezigheid van opgelost organisch materiaal een rol speelt bij het vormen van labiele DGT verbindingen uit deze inerte vorm (Abdulbur- Alfakhoury E. and M. Leermakers, Journal of Analytical Atomic Spectrometry, 2021. DOI: 10.1039/D0JA00442A).. Daarnaast werd de DGT-techniek gevalideerd door verschillende in situ DGT-veldstudies die zowel in stedelijke rivieren als in een ziekenhuiseffluent werden uitgevoerd. Een eerste interlaboratorium vergelijkingsstudie tussen VUB-AMGC en Lille-LASIRE werd uitgevoerd en toonde een goede overeenkomst voor Pt en toonde de noodzaak aan om de DGT parameters

3 te evalueren in verdere studies (Abdulbur-Alfakhoury, E., et al. (2021). Science of the Total Environment 784). De combinatie van laboratoriumexperimenten en veldstudies maakt het mogelijk om een beter inzicht te krijgen in de toepasbaarheid van de DGT-techniek als een milieumonitoringstool voor de nieuwe opkomende verontreinigende stoffen zoals PGE's in omgevingen die zwaar worden beïnvloed door menselijke activiteiten, waardoor het mogelijk wordt om hun bronnen te traceren, hun evolutie te monitoren hun mogelijke toekomstige impact te evalueren.

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Contents Abstract of the PhD research ...... 2 Nederlandstalige abstract ...... 3 Contents ...... 5 List of abbreviations ...... 9 List of tables ...... 17 Chapter 1: Background and Literature Study...... 19 1.1. Occurrence, properties and applications of platinum-group elements (PGEs) ...... 21 1.1.1. Application in automobile catalysts ...... 21 1.1.2. Medical applications ...... 23 1.2. Anthropogenic PGE emissions into the environment ...... 23 1.2.1. PGE emission by automotive catalytic converters ...... 23 1.2.2. PGE Emissions from Non-automobile Sources ...... 24 1.3. Transformations and transport of PGEs in the environment ...... 25 1.4. PGEs in environmental matrices ...... 28 1.4.1. PGEs in air, soil, dust and vegetation (terrestrial environment)...... 28 1.4.2. PGEs in rivers, coastal waters, and oceans (in the aquatic ecosystem)...... 31 1.5. Aquatic speciation of PGEs ...... 33 1.6. Toxicity and Health effects ...... 36 1.7. Analytical methods for the determination of PGEs ...... 37 1.7.1. Instrumental methods ...... 37 1.7.2. Analyte Pre-Concentration and Matrix Separation Methods ...... 38 1.7.3. Ion exchange and chelating resins for PGEs ...... 41 1.8. In situ monitoring and dynamic speciation measurements in solution using DGT ...... 52 1.8.1. Introduction ...... 52 1.8.2. Principles of the passive sampler ...... 54 1.8.3. Principles of the Diffusive Gradients in Thin-films (DGT) technique ...... 56 1.8.4. Components of the DGT device ...... 58 1.8.5. Parameters influencing DGT measurement ...... 62 1.8.6. Applications of the DGT technique ...... 66 1.9. Aims of the research and structure of the thesis ...... 67 Chapter 2: Methodology ...... 71 2.1 General Procedures ...... 72 2.2. Analysis of PGEs with ICPMS ...... 72

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2.2.1. Principles of ICP-MS measurements ...... 72 2.2.2. Specifications of the Thermo Fisher Element II ICP-SF-MS ...... 75 2.2.3. Stability of PGEs in solutions and instrumental memory effects ...... 76 2.2.4. Interferences in PGE analysis ...... 79 2.2.5. Practical considerations in ICPMS measurements ...... 88 2.3. DGT preparation and procedures ...... 90 2.3.1. Diffusive gel preparation ...... 90 2.3.2. Binding gels preparation investigation ...... 90 2.3.3. DGT assembly ...... 92 2.3.4. Diffusion cell experiments ...... 92 Chapter 3: Development of the Diffusive Gradients in Thin-Films Technique (DGT) for platinum (Pt), palladium (Pd), and rhodium (Rh) in Natural Waters ...... 94 Keywords ...... 95 3.1. Introduction ...... 96 3.2. Materials and methods ...... 98 3.2.1. General Procedures ...... 98 3.2.2. DGT preparation and assembly ...... 98 3.2.3. Preparation of the deployment solutions ...... 98 3.2.4. Sample analysis ...... 99 3.2.5. Characterization of DGT performance ...... 99 3.3. Results and Discussion ...... 101 3.3.1. Chemical interactions between diffusive gels and PGEs ...... 101 3.3.2. Uptake kinetics of PGEs by binding gels ...... 102 3.3.3. Elution factor ...... 103 3.3.4. Diffusion coefficients using the diffusive cell...... 104 3.3.5 Effective Diffusion coefficient measurements using time-series DGT deployments...... 105 3.3.6. The effects of pH, ionic strength, and DOM on uptake on DGT measurements ...... 110 3.3.7. The effect of interferences on ICPMS measurement and selectivity of binding gels ...... 111 3.3.8. Binding gel blanks and DGT detection limits...... 117 3.4. Conclusions ...... 118 Acknowledgements:...... 119 Chapter 4: Evaluation of the effect of solution ageing on the DGT speciation of Rh and Pt ...... 120 Abstract: ...... 121 4.1. Introduction: ...... 122

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4.2. Materials and methods ...... 123 4.2.1. DGT probe preparation ...... 123 4.2.2. Preparation of the deployment solutions ...... 123 4.2.3. Uptake kinetics on resin gels...... 123 4.2.4. Effective diffusion coefficients using time series DGT deployment ...... 123 4.2.5. Sample analysis ...... 124 4.3. Results and Discussion ...... 124 4.3.1. Uptake kinetics of PGEs by binding gels ...... 124 4.3.2. Effective Diffusion coefficient measurements using time-series DGT deployments .. 125 4.3.3. The effects of ionic strength and DOM on uptake on DGT measurements...... 129 4.4. Conclusions ...... 130 Chapter 5: Elimination of Interferences in the Determination of Platinum, Palladium and Rhodium by Diffusive Gradients in Thin-Films (DGT) and Inductively Coupled Plasma Mass Spectrometry (ICP MS) using selective elution ...... 131 Abstract ...... 132 Keywords ...... 132 5.1. Introduction ...... 133 5.2. Materials and methods ...... 136 5.2.1. General Procedures ...... 136 5.2.2. DGT preparation and assembly ...... 136 5.2.3. Metal uptake on resin gels: ...... 137 5.2.4. Removal the interferences from the resins gels ...... 137 5.2.5. Sample analysis ...... 138 5.3. Results and Discussion ...... 138 5.3.1. Uptake of metals on resin gels ...... 138 5.3.2. Removal the interferences from the resin gels ...... 141 5.4. Conclusion ...... 150 Chapter 6: Developing SPE to measure PGEs from surface water ...... 151 . 6.1 Introduction ...... 152 6.2. Experimental ...... 153 6.3. Results and discussion ...... 159 6.4. Conclusions: ...... 178 Chapter 7: Distribution of platinum (Pt), palladium (Pd), and rhodium (Rh) in urban tributaries of the Scheldt River assessed by Diffusive Gradients in Thin-Films Technique (DGT) ...... 179 Abstract: ...... 180

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7.1. Introduction ...... 181 7.2. Materials and methods ...... 182 7.2.1. Description of the study area ...... 182 7.2.2. DGT functioning ...... 186 7.2.3. DGT preparation, deployment, treatment and water sampling ...... 186 7.2.4. Sample analysis ...... 188 7.3. Results and Discussion ...... 188 7.3.1. Evaluation of DGT performance...... 188 7.3.2. PGE concentrations in rivers ...... 189 7.3.3. PGE concentrations in hospital effluents and in treated and untreated sewers ...... 193 7.3.4. Concentration of dissolved Pt and Pd in untreated and treated wastewaters ...... 194 7.4. Conclusion ...... 197 7.5. Acknowledgments ...... 198 7.6. Supplementary Information...... 198 Chapter 8: Conclusions and perspectives ...... 200 8.1. General Conclusions ...... 201 8.1.1. Method development ...... 201 8.1.2. Selective removal of interfering elements from resin gels ...... 202 8.1.3. Importance of ageing of solutions ...... 202 8.1.4. Field deployments and interlaboratory comparison ...... 202 8.1.5. Overall evaluation of the resins used ...... 204 8.2. Future perspectives ...... 204 Annex I ...... 206 Supplementary information for chapter 4: ...... 206 References: ...... 212 Curriculum Vitae: ...... 235 List of publications ...... 235

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List of abbreviations

A Exposure area of a DGT sampler AAS Atomic Adsorption Spectrometry AGA Pure agarose gel APA Agarose cross-linked polyacrylamide gel AR Aqua regia ASV Adsorptive Stripping Voltammetry B magnetic sector BE Nier-Johnson Geometry BG borosilicate glass C Celsius C Concentration of metal in a bulk solution

CDGT Time-averaged DGT labile concentration CPC cancerostatic platinum compounds D Diffusion coefficient of metal d the ionic diameter of the analyte DBL Diffusive boundary layer

Dcell Diffusion coefficient of metal in a agarose gel estimated by the diffusion cell

Dcell diffusive coefficient measured with the diffusion cell Effective diffusion coefficient of metal/ Diffusion coefficient of metal in a D DGT polyacrylamide Effective diffusion coefficient of metal DGT Diffusive Gradients in Thin-Films technique DOC Dissolved Organic Carbon E electrostatic sector EB reversed Nier-Johnson Geometry Eh Eh Electrochemical potential ETA– electro-thermal atomization atomic absorption spectrometry AAS F Flux of metal through a diffusive DGT layer FA Fulvic Acid fe Elution efficiency of metal fe Elution factor of metal FEP fluorinated ethylene propylene HA Humic acid HEDPA 1- hydroxyethane-1,1-diphosphonic acid HR High resolution mode HS Humic substances HSAB Hard and Soft Acids and Bases I Ionic Strength ICP– AES inductively coupled plasma atomic emission spectrometry ICP-MS Inductively coupled plasma sector field mass spectrometry

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ICP– inductively coupled plasma optical emission spectrometry OES IER Ion-Exchange resins IS internal standardization IUPAC IUPAC International Union of Pure and Applied Chemistry

Kb Boltzmann constant LOD LOD Limit of detection LR low resolution mode m Thickness of a filter membrane M M Mass of analyte M M Metal

MDGT mass of accumulated analyte on the passive sampler (DGT) MDL Method Detection Limit

Minorg mass of accumulated analyte contributed from inorganic species ML Metal ligand

Morg mass of accumulated analyte contributed from organic species MQ MilliQ water MR meduim resolution mode MΩ Mega-ohm NAA neutron activation analysis NAA neutron activation analysis NOM natural organic matter PAM Polyacrylamide PE polyethylene PGEs Platinum group elements PP polypropylene PTFE polytetrafluorethylene Q-ICP- quadrupole Inductively coupled plasma sector field mass spectrometry MS REE rare earth elements RSD Relative standard deviation SCR Strong cation exchange resins SF- high-resolution sector field -Inductively coupled plasma sector field mass ICPMS spectrometry SPE Solid phase extraction t Deployment time T temperature (K) TDS total dissolved solids TEM transmission electron microscopy TEMED N,N,N’,N’-tetramethylenediamine TOF- ICPMS time-of-flight Inductively coupled plasma sector field mass spectrometry TWAC time-weighted average concentration TWCs three way catalysts

10 v v/v Volume/volume Veluant Veluant Volume of an eluent WWTPs wastewater treatment plants δ Thickness of a diffusive boundary layer Δg Thickness of a diffusive layer ΔG° free energy Δgel Thickness of a diffusive gel μ dynamic viscosity

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List of figures Figure 1. 1. Simplified graph shows the chemical reaction in the car catalytic convertor [27]...... 22 Figure 1. 2. Main emission sources and distribution pathways of platinum group elements in the environment [17]...... 26 Figure 1. 3. represents distribution pathways and concentration ranges of platinum group elements in selected environmental compartments including values reported until 2013 [77]...... 31 Figure 1. 4. Literature data on PGE concentrations in estuaries, and marine environments. [101]...... 32 Figure 1. 5. Simplified chemical structure of Magpie Polymers - Italmatch Chemicals MP- 102, MP-101 polymers as well as MPX-317 and MPX-310 complex with Cd and phosphorous-containing resins from other producers (Diphonix, Lewatit VP OC 1026 and Lewatit TP-272)...... 49 Figure 1. 6. Passive sampling devices that operate in different accumulation régimes [263]. 55 Figure 1. 7. DGT assembly ...... 56 Figure 1. 8. Schematic representation of a concentration gradient through a DGT device at steady state...... 57 Figure 1. 9. Schematic view of labile and non-labile trace element compounds diffusing into a DGT probe. Pore-size is about 10 nm for Open-Pored and 1 nm for Restricted-Pored hydrogels [259]...... 60

Figure 2. 1. General scheme of an ICP-MS instrument [324] ...... 74 Figure 2. 2. Schematic setup of a magnetic sector field ICP-MS (in this example of the Element 2)...... 75 Figure 2. 3. Shows the repeated measurement (n=45) of 80 µg L-1 Pt, Pd and Rh using LR mode in 3% HCl matrix using 3% HCl carrier solution. Dilution factor is 10...... 78 Figure 2. 4. Shows the repeated measurement (n=13) of 10 µg L-1 Pt and Rh, 6 µg L-1 using LR mode in 3% HNO3 matrix using 3% HNO3 carrier solution. Dilution factor is 10...... 78 Figure 2. 5. Interference equivalent concentrations (IEC) in function of the concentration of interfering elements on 105Pd, 106Pd, 108Pd, 194Pt, 195Pt and 103Rh measured in low, medium and high resolutions by ICP-SF-MS...... 86

Figure 3. 1. The relative accumulated mass of Rh, Pd and Pt in diffusive gels in function of the time (n= 2)...... 102 Figure 3. 2. Uptake kinetics of Pt, Pd and Rh on different resin gels (n= 3)...... 103 Figure 3. 3. The accumulation of mass of PGEs in the receptor compartment of a diffusion cell over time. The experimental conditions for the data were as follows: pH of source and receiving solution = 8.3 ±0.2, salinity = 33.2, DOC ˂ 1 mg.L-1, source solution concentraion...... 105 Figure 3. 4. Time-series experiment (The accumulated mass in function of time). The diffusion coefficients D (cm2s-1), of Pt, Pd and Rh in the diffusive layers Δg = 0.092 cm (the agarose diffusive gel (0.75 mm) and the filter (0.17 mm)) using DGT units (n=2 at each time)

12 with AGA diffusive gel layers overlaid on binding gel of AGA with S914, S920, S985, MPX-317 and MP-102 and deployed in filtrated river water (FRW) spiked with Pt, Pd, and Rh 24h before the deployment. The experimental conditions for the data were as follows: pH~8.64, salinity = 0.5. Area of window (A) = 3.14 cm2, deployment solution concentration (C= ng mL-1) for each resin gel test (Pt, Pd, Rh) (average concentration ± standard deviation over the 72h): S914 (19.9±1.7,14.0±1.1, 14.7±0.8), S920 (19.87±1.7, 16.8±1.22, 23.4±1.1), S985 (19.9±1.7, 13.9±0.5, 23.4±1.0), MPX-317 (19.9±1.7, 14.2±0.9, 23.4±1.1), MP-102 (12.4±1.0, 9.1±0.6, 14.9±0.7); temperature 19°C; diffusive gel thickness and filter Δg = 0.092 cm; graph gradient of mass accumulation over time (α) = ng/h. Mean values and standard deviation (error bars) of duplicate measurements are given...... 107 Figure 3. 5. Effect of pH, ionic strength and DOM on the ratio of DGT Pt, Pd and Rh concentrations, CDGT, to their concentrations in the bulk solution, C sol. Mean values and standard deviation (error bars) of triplicate measurements are given (n = 3)...... 111 Figure 3. 6. The uptake (ng) of Pt, Pd, Rh, Zn, Cu, Sr, Rb, Y, Hf and Pb over time using DGT units with diffusive layers (agarose diffusive gel (0.75 mm) and filter (0.17 mm)) overlaid on binding resin gel of AGA with S914, S920, S985, MPX-317 and MP-102 and deployed in spiked filtrated river water. The experimental conditions for the deployment solution were as follows T = 21 ⁰C, pH = 8.6, salinity = 0.4, DOC = 4.3 mg L -1. Mean values and standard deviation (error bars) of duplicate measurements are given (n =2)...... 115 Figure 3. 7. The accumulated mass (ng) (mean ± the standard deviation of 5 replicates), of Rh, Pd, Pt, Rb and Sr using DGT units (n=10) of diffusive layers (agarose diffusive gel (0.75 mm) and filter (0.17 mm)) overlaid on binding resin gel of agarose (0.5 mm) with Purolite S914 and deployed in filtrated seawater spiked with Pt, Pd and Rh, 24h before the deployment. 5- Binding gels were washed for 12 h with 10 mL MQ-water before the elution (washed). 5- Binding gels were eluted immediately after the deployment without any wash (No-wash). Recovery factor is considered 1 for Sr and Rb...... 116 Figure 3. 8. The uptake (ng) of Sr and Rb over time using DGT units with AGA diffusive gel layers (0.75mm) overlaid on binding resin gel of AGA with S914, S920, MPX-317 and MP- 102 and deployed in spiked filtrated river water. Resin gels were washed with MQ water for 12h before the elution step. Data of Sr and Rb are normalized to 113In. Recovery for Sr and Rb considered 1 and the dilution factor is 10. The experimental conditions for the deployment solution were as follows T = 21 ⁰C, pH = 8.6, salinity = 0.4, DOC = 4.3 mg L-1. STD is the for duplicate pistons measurement at each time...... 117

Figure 4. 1. Uptake kinetics of Pt and Rh on different resin gels and different solution ageing (1and 17) days (n= 3) at the same conditions (T = 24⁰C, pH = 6.1±0.2)...... 125 Figure 4. 2. Accumulated mass of metals in function of time for deployments in 17 days aged solutions. Filtrated river water (FRW). Filtrated sea water (FSW)...... 126 Figure 4. 3. Effect of ionic strength and DOC on the ratio of DGT Pt, Pd and Rh concentrations, CDGT, to their concentrations in the bulk solution, Csol Mean values and standard deviation (error bars) of triplicate measurements are given (n = 3). DDGT for each binding gel is adopted from our previous work [221] taking into account temperature correction...... 129

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Figure 5. 1. The uptake (ng) of Pt, Pd, Rh, Zn, Zr, Cu, Mo, Cd, Sr, Rb, Y, Hf and Pb over time using DGT units with diffusive layers (agarose diffusive gel (0.75 mm) and filter (0.17 mm)) overlaid on binding resin gel of AGA with S914, S920 and MPX-317 and deployed in spiked filtrated river water...... 141 Figure 5. 2. Accumulated mass of metals (ng) on S920 after different elution procedures: ref. -1 is the reference method with only a deionized water rinse; 1) 0.5 mol L Na2SO4, 2) 0.5 mol -1 -1 -1 -1 L NaNO2, 3) 1 mol L NaCl, 4) 0.5 mol L CH3COONa, 5) 0.1 mol L CH3COOH, 6) -1 -1 -1 0.01 mol L HNO3, 7) 0.05 mol L H2SO4, 8) 0.1 mol L KOH, 9) 50% CH3OH, 10) 0.5 -1 -1 mol L NH4SCN, 11) 0.5 mol L CH4N2S, and 12) 50% C2H5OC2H5. A single 12h elution was performed in each treatment...... 143 Figure 5. 3. Accumulated mass of metals on S914 after different elution procedures: ref. is -1 -1 the reference method with only a deionized water rinse; 1) 0.5 mol L Na2SO4, 2) 0.5 mol L -1 -1 -1 NaNO2, 3) 1 mol L NaCl, 4) 0.5 mol L CH3COONa, 5) 0.1 mol L CH3COOH, 6) 0.01 -1 -1 -1 -1 mol L HNO3, 7) 0.05 mol L H2SO4, 8) 0.1 mol L KOH, 9) 50% CH3OH, 10) 0.5 mol L -1 NH4SCN, 11) 0.5 mol L CH4N2S, and 12) 50% C2H5OC2H5. A single 12h elution was performed in each treatment...... 146 Figure 5. 4. Coordination complex of the thione form with metals ion...... 147 Figure 5. 5. Possible binding modes for Cd with thiourea functional groups...... 147 Figure 5. 6. Accumulated mass of metals (ng) on MPX-317 after different elution procedures: -1 ref. is the reference method with only a deionized water rinse. 1) 0.5 mol L Na2SO4, 2) 0.5 -1 -1 -1 -1 mol L NaNO2, 3) 1 mol L NaCl, 4) 0.5 mol L CH3COONa, 5) 0.1 mol L CH3COOH, -1 -1 -1 6) 0.01 mol L HNO3, 7) 0.05 mol L H2SO4, 8) 0.1 mol L KOH, 9) 50% CH3OH, 10) 0.5 -1 -1 mol L NH4SCN, 11) 0.5 mol L CH4N2S, and 12) 50% C2H5OC2H5 .A single 12h extraction was performed in each treatment...... 149 Figure 5. 7. MPX-317 resin and Cd2+ Possible bonding modes of cadmium to the chelating resins [234]...... 149

Figure 6. 1. Recovery % (not retained by the column) (±S.D.%, n = 4) in model solutions (200 ml/ 4 fractions) of Rh (470 ng), Pd (480 ng), Pt (537 ng), Sr (51543 ng), Rb (15184 ng), Cu (2319 ng), Zn (9139 ng), Y (3.3 ng), Hf (5.9 ng), Mo (1685 ng), Cd (69 ng), Zr (4.2 ng), and Pb (680 ng) in 0.5M HCl on Dowex 50W X8 200-400 mesh cation-exchanger determined by ICPMS. The first graph: C1, C2, C3 and C4 are four replicates of the column separation (uncertainty is the standard deviation of the C1, C2, C3 and C4 in %). The high uncertainty represents the unrepeatable behaviour of the elution. The second figure represents the mean recovery of the columns C (1-4). Dilution factor is 2 for ICPMS measurements. ... 161 Figure 6. 2. Pre-concentration by heating procedure. Mean recovery R_ Mean%: Rh (100.13±3.05), Pd (98.83±9.75), and Pt (95.92±8.0). (±S.D. %, n =3; T2, T3 and T4). The * individual replicates considered are (T2, T3 and T4). Replicate (T1 ) is excluded from the calculation due to the high loss. Error bars represent the standard deviation in% of 3 replicates...... 162 Figure 6. 3. Recovery (Not sorbet fraction by the column) (±S.D.%, n = 4) in filtrated river water (200 ml/ 4 fractions) of Rh (470 ng), Pd (480 ng), Pt (537 ng), Sr (51543 ng), Rb (15184 ng), Cu (2319 ng), Zn (9139 ng), Y (3.3 ng), Hf (5.9 ng), Mo (1685 ng), Cd (69 ng), Zr (4.2 ng), and Pb (680 ng) on Dowex 50W X8 200-400 mesh cation-exchanger determined

14 by ICPMS. The first graph: C1, C2, C3 and C4 are four replicates of the column separation (uncertainty is the standard deviation of the C1, C2, C3 and C4 in %). Dilution factor is 2 for ICPMS measurements...... 164 Figure 6. 4. Uptake % (sorbet fraction by the resins using the static mode (Batch process) for 48h) (±S.D.%, n = 3) in filtrated river water using borosilicate glass bottles (200 ml, adjusted to pH 0.3 ± 0.1 with HCl) of Rh (6802 ng), Pd (3967ng), Pt (7847ng), on 1 gr of each resins of S920, S914, S924, MPX317 and MPX 310 determined by ICPMS...... 169 Figure 6. 5. Uptake % (retained fraction in the column) (±S.D.%, n = 4) in filtrated river water (200 ml) of Rh (1027 ng), Pd (1060 ng), Pt (1013 ng), Sr (42991 ng), Rb (13215 ng), Cu (142 ng), Zn (6125 ng), Y (5 ng), Hf (3.8 ng), Mo (1500ng), Cd (3 ng), Zr (12 ng), and Pb (12 ng) on Purolite S920 determined by ICPMS. Mass uptake (%) is calculated from the difference between the initial and effluent concentration in %. The first graph: C1, C2, C3 and C4 are four replicates of the column separation (uncertainty is the standard deviation of the C1, C2, C3 and C4 calculated in %). Dilution factor is 2 for ICPMS measurements. Mean Mass sorbet in % in pH = 0.3± 0.1 HCl, Rh = (26.1 ±3.3) %, Pd= (94.0±0.1) %, Pt= (99.9±0.1) %. Mass lost % during the wash the column with 3ml of 0.5M HCl are Rh = 10.0 ±0.744, Pd=0.248 ±0.035, Pt=0.024±0.010...... 171 Figure 6. 6. Recovery % of Rh, Pd and Pt from S920 resins under dynamic conditions using Thiourea (1M) in 2M HCl eluent. Mean recovery (Mean elution) % represent the average of the recovery of 4 replicates. Error bars represent the uncertainty (±S.D.%, n = 4) of 4 replicates. PGEs-loaded resins mass are as follows: Rh=267.02 ng, Pd= 966.4 ng, Pt = 1011.98 ng. Reproducibility of Rh, Pd and Pt results expressed as RSD of 4 replicate determinations of analytes in river water and they are 17.7, 15.4 and 15.7, respectively. .... 173 Figure 6. 7. Recovery % of Rh, Pd and Pt from S920 resins under static mode using hot aqua regia 60°C eluent for 24 h. Mean recovery (Mean elution) % represent the average of the recovery of 4 replicates. Error bars represent the uncertainty (±S.D.%, n = 4) of 4 replicates. PGEs-loaded resins mass are as follows: Rh=267.02 ng, Pd= 966.4 ng, Pt = 1011.98 ng. Reproducibility of Rh, Pd and Pt results expressed as RSD of 4 replicate determinations of analytes in river water and they are 17.51, 12.38 and 7.94, respectively...... 174 Figure 6. 8. Rh, Pd and Pt elution efficiency (ng/ disk) using different eluents, the reference is (AR) hot aqua regia at 60°C, Thiourea 1M in 2M HCl (TU_HCl) at room temperature T25°C and T60°C and Thiourea 1M in 2M HNO3 (TU_HNO3) at room temperature T25°C and T60°C. Error bars represent standard deviation of 3 replicates. Rh, Pd and Pt data are normalized to 115In, 209Bi and 185Re...... 177

Figure 7. 1. A map of the Scheldt River Basin and the sampling stations in the tributaries Zenne, Marque Rivers ...... 183 Figure 7. 2. Rainfall (mm) during sampling period of the Marque River in April-May 2019. Red bars indicate days when grab sampling was performed. Black line is the DGT deployment period...... 185 Figure 7. 3. DGT PGE concentrations obtained in the Marque River using different resins: inter-laboratory study. Uncertainty bars of DGT concentrations represent standard deviations of 4 replicates...... 190

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Figure 7. 4. DGT concentrations of Pt, Pd and Rh obtained in the Zenne River in September 2018...... 193 Figure 7. 5. DGT PGE concentrations in combined sewage drain of UZ Brussels Hospital, Jette using 3 different resins (S920, S914 and MPX317)...... 194 Figure 7. 6. Cumulated Pt and Pd loads at station Z5 (Zenne River downstream WWTP Brussels South), at station Z7 (Zenne river upstream WWTP Brussels North) and at station Z9 (Zenne river downstream WWTP Brussels North). At Z9 the loads in the Zenne river are compared to the loads in the effluents of WWTP Brussels North ...... 196

16

List of tables

Table 1. 1. Heavy metal and Rare earth elements Concentrations ...... 29 Table 1. 2. Ratios of PGE in sediment samples...... 35

Table 2. 1. Specifications of the ICP-SF-MS instrument used in this thesis ...... 76 Table 2. 2. Isotopes and first ionization potentials of the analytes and internal standards (ISs) ...... 80 Table 2. 3. Potential spectral interferences on the masses of Pd, Pt and Rh[178, 343] ...... 81 Table 2. 4. Physical-chemical characteristics of ion exchangers investigated ...... 90

Table 3. 1. Elution factor fe ...... 104 Table 3. 2.Diffusion coefficients (mean ± standard deviation of 3 replicates, ×10-6 cm2 s-1, at 25⁰C) obtained in the diffusion cell experiments in seawater: artificial seawater (ASW), filtrated seawater (FSW) and in fresh water: 0.01M NaCl and filtrated river water (FRW). 105 Table 3. 3. Shows the Diffusion coefficients (×10-6 cm2 s-1, at 25⁰C) with different binding gel obtained in time series experiments in artificial seawater, filtrated seawater, 0.03 M sea salt and filtrated river water. The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution ...... 108 Table 3.4. DGT blank values (n=4) and method detection limits (MDL) for the different resin gels. MDLs are calculated based on a deployment time of 14 days, ∆g =0.092 cm, A=3.14 cm2, D river water at 25°C, pH=8±0.2...... 118

Table 4. 1. Diffusion coefficients (×10-6 cm2 s-1, at 25⁰C) with different binding gel obtained in time series experiments in filtrated seawater and filtrated river water fresh obtained from our previous work [221] and aged for 17days after spiking with Pt and Rh. The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution ...... 126

Table 6. 1. Blank and MDL (ng L-1) ...... 159 Table 6. 2. PGE species found in aqueous chloride media[424] ...... 166

Table 7. 1. DGT preparation, characteristics and handling procedures...... 187 Table 7. 2. Method detection limits (MDL) for the resins used by the AMGC and LASIRE research groups. Abbreviation: N.D.: not determined...... 188 Table 7. 3. Concentrations of dissolved Cu, Zn, Pb, Fe and Gd in surface waters in the Marque River system...... 192 Table 7. 4. Yearly average dissolved Pt (Pt-diss) and Pd (Pd-diss) and total Pt (Pt-tot) concentrations in the treated (outflow) and untreated (inflow) wastewaters of WWTP Brussels North and South, and % reduction after treatment (n=12)...... 195

17

Table 7. 5. Water discharge (Q), concentrations and daily loads (L) of dissolved Pt and Pd in Zenne stations Z5, Z7, Z9 and in the treated waters (outflow) of WWTP-North. Note that at Z7, the water discharge is not exactly known because of unknown contributions of untreated sewage, which should, however, be small (SBGE pers. Comm.)...... 195 Table 7. 6. Water discharge (Q), number of hospital beds (bds), concentrations and loads (L) of Pt and Pd in the untreated sewage of WWTP Brussels North (WWTP-N), South (WWTP- S), and all Brussels WWTP; and in the effluents of hospitals UZ-Brussels (UZ) and all hospitals in Brussels. Contribution of Pt and Pd loads in all hospital effluents to their total loads in Brussels sewage (H/Swg)...... 197 Table 7. 7. Total dissolved Pt concentrations (0.45 µm filter) and DGT Pt concentrations with S914, S920 and MPX-317 binding phases for 3 deployment times (5, 7 and 14 days) at station Z5 in Zenne river in September 2018. The measurements named “0.45 µm” represent the direct total dissolved Pt determination. Uncertainty bars of DGT concentrations represent standard deviation of 4 replicates. Uncertainty bars of total dissolved concentrations (< 0.45 µm) represent standard deviations of duplicate measurements at the beginning and at the end of the DGT deployment...... 198

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Chapter 1: Background and Literature Study

19

The six metallic elements platinum (Pt), palladium (Pd), rhodium (Rh), ruthenium (Ru), iridium (Ir) and osmium (Os) are known as Platinum group elements, abbreviated as (PGEs). The high melting point, the excellent mechanical, catalytic, as well as corrosion and oxidation- resistant properties of the PGEs are particularly pronounced at high temperatures. This has led to a number of applications for PGEs (especially Pt, Pd and Rh) and their alloys: (1) the introduction (1975 (USA), 1976 (Japan), and 1986 (Europe)) of PGEs as the components of automobile catalysts for cleaning exhaust gases and (2) the introduction (1978) of cisplatin into worldwide chemotherapy as an effective anticancer agent. These applications were found to be anthropogenic sources of ultra-traces of PGEs in the environment due to thermal and mechanical abrasion of the catalysts during vehicle operation and a release of pharmaceuticals used from urban and hospital effluents. Additional trace amounts of PGEs can also occur in the environment as a result of their wide industrial (chemical, electrical, glass, and jewellery) applications. It was previously believed that emitted PGEs into the environment are relatively inert [1], but the discrepancies between expected and observed PGE concentrations has now been shown that these metals undergo environmental transformations into more reactive species [2-4], which become bioavailable and contribute to PGE fluxes in different environmental compartments [5], posing a great threat to human health through the inhalation of fine dusts, transfer, and transformation in food chain [6-9]. Increased PGE concentrations have been found in various water samples (rainwater, groundwater, surface water, sea water, sea sediments, sludge, etc.)[10, 11]. Although the level of platinum metals in aquatic ecosystems is relatively low (pg L-1 to ng L-1) as compared with their concentrations in other components of the environment, these pollutants can be expected to appreciably affect aquatic animals due to their ability to bio-accumulate. Therefore, it is necessary to monitor PGEs in various environment matrixes to investigate their bioavailability, biogeochemical behaviour, and toxicity [12]. The method with a great potential in this regard is the Diffusive Gradients in Thin-Films (DGT). This in situ technique relies on the diffusive mass transport from the bulk solution to a binding phase layer and was invented by Zhang and Davison over 30 years ago [13]. DGT provides time-weighted average concentrations of labile metal species, thus proving to be useful as a long term monitoring tool and as a metal bioavailability predictor [14-16].

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1.1. Occurrence, properties and applications of platinum-group elements (PGEs)

PGEs is a group of six elements—platinum (Pt), palladium (Pd), rhodium (Rh), iridium (Ir), osmium (Os), and ruthenium, (Ru)—that are among the least abundant elements in the Earth’s continental crust. The natural occurrence of platinum group elements (PGEs) in the upper earth crust is typically in the sub μg kg−1 range and higher in distinct sites where PGEs are part of noble metal-rich ores[17]. Current estimates of average PGE abundances in the upper continental crust (UCC) are Rh = 0.018 ppb, Pt = 0.599 ppb, and Pd = 0.526 ppb [18]. These values have been used as background values for PGE in several environmental studies [19, 20]. Due to the widespread use of Pt, Pd and Rh as catalytic converters in cars since the ‘70s, and the use of Pt anticancer drugs, the emissions of these three elements have increased worldwide and their potential environmental impact of concern. PGEs are mainly mined in South Africa, Siberia, the United States, Canada, Colombia, Zimbabwe [21, 22]. Worldwide production of PGEs has been growing steadily since 1970. In 2011, the global supply of these metals amounted to183 tons of Pt, 209 tons of Pd, and 22 tons of Rh [23]. Therefore, in this thesis, we will focus our attention on the elements Pt, Pd and Rh.

The unique properties of PGEs include their chemical inertness (high oxidation potential), catalytic activity, high mechanical strength, electric conductivity and biocompatibility and their catalytic properties are used in the automobile industry, petrochemistry, chemical industry and as anti-cancer drugs and are further found in jewelry, glass and textile industry and electronics. 1.1.1. Application in automobile catalysts A catalytic converter is a unit that fits into the front part of exhaust system of a vehicle, close to the engine, to reduce the emission of gaseous pollutants, such as carbon monoxide (CO), nitrogen oxides (NOx) and hydrocarbons (HCs). It is usually fitted with a heat-shield to limit internal temperature drops, i.e. the heat loss. Removal of the pollutant gases from the exhaust of a properly tuned combustion engine takes place by either reduction (for NOx) or oxidization (for CO and HCs). The catalytic role of PGEs in the automobile industry (and chemical industries) is the result of two key properties of PGEs. First, their resistance to oxidation, and second, their capacity to adsorb gaseous molecules on their surface, allowing a fast surface reaction. While two-way catalysts can still be found on diesel engines, the most common ones in present day vehicles are three way catalysts (TWCs) which catalyse the transformations of

HC, CO and NOx to H2O, CO2 and N2 in three consecutive stages: i) reduction of NOx to N2; ii) oxidation of unburnt HC to CO2 and H2O; and iii) oxidation of CO to CO2 (Figure 1.1)

21

Combination of heat with the precious metal catalysts facilitates these heterogeneous reactions in the converter. Modern vehicle exhaust catalysts convert over 90% of CO, HCs and NOx into carbon dioxide (CO2), water and nitrogen. Catalytic converters are manufactured with a ceramic or metal substrate, covered in an alumina wash coat containing metal additives such as Ce, Zr, etc. On the surface of the wash coat, varying proportions of PGE (mainly Pt, Pd, and Rh) are highly dispersed. Three metals of platinum group metals (PGEs), i.e. platinum (Pt), palladium (Pd) and rhodium (Rh) are employed as active materials in three-way catalysts (TWC). The gasoline engine is supported by the use of TWC. This oxidizes CO and HC to

CO2 and H2O, while at the same time reducing NOx to N2. Pt and Pd are involved in the oxidation of HC and CO, while Rh is used in the reduction of NOx [24-26] . In a typical diesel oxidation catalyst, Pt or Pd is used [24-26]. The noble metals, Pt, Pd and Rh, are fixed on the surface of the washcoat usually by impregnation of hexachloroplatinic (IV) acid

(H2PtCl6.6H2O), palladium chloride (PdCl2) and rhodium chloride (RhCl3) precursor salts, respectively[20]. The proportion of Pt, Pd and Rh in a catalytic converter depends on a number of factors, such as the manufacturer, the characteristics of the vehicle including the engine power, the type of fuel consumed by vehicle (gasoline or diesel), as well as the weight of vehicle and the required catalytic functions. Although Pt is an efficient oxidation catalyst, it is more costly and sensitive to catalyst poisons than Pd [20]. For economic reasons Pd is the preferred metal for petrol-engines, and for technical reasons Pt is the main constituent for diesel-engines [21].

Figure 1. 1. Simplified graph shows the chemical reaction in the car catalytic convertor [27].

22

1.1.2. Medical applications Platinum group metals are exploited in medicine for cancer treatment and the preparation of dental fillings [20, 23]. Pt complexes are of great medical importance. Dental alloys can contain up to 18 % Pt and up 80 % Pd [28]. For the past 30 years Pt compounds have been used to treat numerous types of tumours (testicular, ovarian, bladder, and many other kinds of tumours). Two Pt complexes, namely cisplatin (cis-diammine-dichloro-platinum[II]) and carboplatin (diamine[1,1-cyclobutanedicarboxylato] platinum[II]) have been successfully applied for the treatment of human carcinomas [29, 30]. The new generation of anticancer drugs introduce the palladium, ruthenium, gold, silver, rhodium, iridium, osmium, rhenium complexes as anticancer agents [31, 32]. Evidence based on the unique isotope signature of industrial Os led to argue that this element is released from biomedical facilities [27]. So far, Ru(II) and Ru(III) complexes have shown very promising properties while the Ru(III) compound, [ImH] [trans-

Cl4(Me2SO)(Im)Ru(III)] (Im=imidazole, NAMI-A), is the first ruthenium compound that successfully entered phase I clinical trials. There have been also tested rhodium based compounds but found to be less effective as anticancer agents mainly due to their toxic effects. Dimeric μ-Acetato dimers of Rh(II) as well as monomeric square planar Rh(I) and octahedral Rh(III) complexes have shown interesting antitumor properties[33].

1.2. Anthropogenic PGE emissions into the environment

1.2.1. PGE emission by automotive catalytic converters During the release of the exhaust gases from the engine, the surface of the washcoat is chemically and physically stressed by fast changing oxidative and reductive conditions, high temperature and mechanical abrasion, thus, producing the emission of PGE containing particulate matter into the environment [34]. Emission rates are significantly higher for diesel catalysts than for three-way catalysts used with gasoline engines[35], and increase at higher speeds[36]. The amount and rate of PGE emission are also affected by age of the catalyst and the type of fuel additives [37]. PGE emission rates were estimated at 6–8 ng km−1 for Pt, 12– 16 ng km−1 for Pd and 4–12 ng km−1 for Rh for aged gasoline catalysts and 108–150 ng Pt km−1 from aged diesel catalysts [7, 38]. Catalytic converters generally have a service life of between 50 000–100 000 miles; which is reduced by catalysator poisons (Pb, Zn, S and P contamination), engine ignition/fueling faults, usage of the vehicle and short journey use, preventing the converter from reaching the optimal working temperature (approx.400°C) [20]. The principal cause of converter failures is carbon pollution, leading to a partial or sometimes, a total blockage of the catalyst, and the internal fracture of the catalyst surface, usually induced

23 by external/internal physical damage [20]. It is important to note that estimates of PGE emission from automobile catalysts remain quite uncertain despite nearly 30 years of research [27]. Assuming that 500 million vehicles are equipped with PGE catalysts, that the average mileage is 15,000 km/yr per vehicle, and that the Pt emission rate is in the range of 100-800 ng/km, a global emission of 800–6000 kg/yr can be estimated for Pt from automobile catalysts Similarly, applying an emission rate of 5-100 ng/km for Pd (Melber et al., 2002) and 1-10 ng/km for Rh the estimated global emission rate would be 40-800 kg/yr for Pd and 8-80 kg/yr for Rh [39]. 1.2.2. PGE Emissions from Non-automobile Sources Discrepancies between expected and observed PGE concentrations or abundance ratios suggest that a number of sources contribute to PGE fluxes in urban areas[27]. PGE concentrations do not necessarily correlate with traffic intensities. For instance, relatively high PGE concentrations at a site with low traffic intensity in Mexico City were attributed to an industrial source [40]. In addition, no significant difference was found between airborne PGE concentrations in samples collected on weekdays and weekends in Boston, USA, although traffic intensity is expected to be lower on weekends [41]. 1.2.2.1. PGE Emissions from Medical Treatment Applications After treatment, chemotherapeutic agents are excreted by patients. For instance, cisplatin administered to patients is eliminated from the body in the form of highly active and cytotoxic mono-aqua-cisplatin [23]. Emissions of administered drugs and their derivatives are expected to be in form of soluble compounds and excreted via urine [23, 42, 43]. Pharmaceuticals may also be introduced to soil through the application of contaminated sewage sludge or manure as fertilizers or the use of contaminated wastewater for irrigation[44]. In 1999, the concentration of Pt was determined in the sewage from 5 European hospitals, in Austria, Belgium, Germany, Italy and The Netherlands, to provide reliable data on the Pt emission from hospitals with Pt-drug therapy into the aquatic environment [43]. It was shown that 70% of the Pt, administered in the form of either cisplatin or carboplatin, was excreted, and therefore, it ended up in the hospital effluents. The Pt concentration in the total effluents of hospitals ranged widely from <10 ng L-1 for the Belgian and Italian hospitals to 3.5 µg L-1 for the Austrian and German hospitals. In 2014, Platinum-based anticancer drugs in the wastewater effluent from oncology in UK hospital [45], max result measured during the 15 days sampling was 137.8 µg L-1. Since effluents from hospitals are not treated in any special way, these compounds are released directly to municipal wastewater systems [23]. Furthermore, it has been estimated that the majority of drugs containing Pt complexes are

24 excreted with urine at patients’ homes for a period of over 8 years (about 70%) and also enter wastewater[23, 42, 46]. Consequently, hospital effluents are no longer the main expected entry route of anticancer drugs into the aquatic environment [47]. A lack of appropriate treatment methods for the purification of such active compounds due to incomplete removal in wastewater treatment plants (WWTPs) [48] contributes to environmental contamination with PGEs [23].It was reported that total platinum emissions into the public sewage systems via hospitals were approx. 14.3 kg of Pt in 1996 in Germany, compared to approx. 187.2 kg of total Pt from cars[49]. Thus, traces of pharmaceuticals have been detected in hospital effluents, influent/effluent wastewaters and surface and groundwater [44, 45, 48, 50]. The use of sewage sludge in agriculture has been identified as a source of PGE in soils [51]. PGE enrichments have also been found in incinerator ash, reflecting the occurrence of PGE in municipal waste [52]. 1.2.2.2. Other sources Emissions from mining activities and associated activities (e.g. processing plants, smelters) [53].[54], biomass burning [55] and the coal combustion [27] and industrial emissions[40] may result in increased concentrations in environmental compartments nearby the emission sources.

1.3. Transformations and transport of PGEs in the environment

It was previously believed that PGE and their compounds are relatively inert [1], but discrepancies between expected and observed PGE concentrations or abundance ratios has now been shown that these metals undergo chemical or biological transformations into more reactive species mainly as chloro- or organic complexes [2-4], which may be bioavailable and contribute to PGE fluxes in urban areas [5], posing a great threat to human health through the inhalation of fine dusts, transformations and transport to water, soils and sediments, from where they can enter plants, animals, and humans Figure 1.2[6-9].

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Figure 1. 2. Main emission sources and distribution pathways of platinum group elements in the environment [17].

PGEs are principally emitted from automobile catalysts in metallic form or as oxides, in particle sizes ranging from micro-meter to sub-micro-meter (<1 μm to >63 μm) and deposited on the road surface or in the roadside environment [25, 26]. Due to the size of dust particles in exhaust gasses[56] and the number of chemical processes which PGEs undergo, significant quantities of PGEs are converted into bioavailable forms [2, 3]. In addition, chemical transformation is suggested by the occurrence of soluble PGE in automobile exhaust. Whereas soluble Pt represents less than 10% of total Pt emissions, soluble Pd and Rh fractions might be greater than 50% [57] or 100% [25] of total emissions depending on the type and the age (new or old) of the catalyst convertor. - 2- PGEs can be mobilized by simple inorganic anions (e.g. NO3 , SO4 ) [58] as well as compounds that can complex or chelate PGE cations, such as methionine [59], citric acid [59], EDTA [60], and simulated intestinal fluid components [61]. - - 2- 3- Pd is the most soluble in the presence of anionic species (Cl , NO3 , SO4 and PO4 ), followed by rhodium (Rh) and platinum (Pt) [58, 62]. The solubility is enhanced under the combined conditions of lower pH, higher Eh and higher Cl-[63, 64]. PGE can be efficiently mobilized in rain water (pH 4–5) rich by natural organic matter NOM, contributing to their redistribution in the environment[65]. The relative solubility of Rh is higher than that of Pt and that the ratio Pt / Rh is between 1-4 While environmental materials always exhibit Pt / Rh ratios close to the initial value of 5 [66]. Rainwater usually provides a significant medium for the transportation of PGE particles. In run-off water up to 1 mg L-1 Pt was found [67, 68]. The highest solubility

26 of Pt and Rh in ground catalytic converter material in rainwater was reached at pH 1 (0.35– 0.5% Pt and 1.0% Rh), and decreased rapidly to a constant low level between pH 3 and 9 (0.01 – 0.025% Pt and 0.05% Rh) [25]. The relative solubility is thus higher for Rh compared to Pt [25]. In a mixture of rainwater and soil, Pt solubility is lower than in rainwater alone (0.001% Pt), indicating a retention capability of the soil due to adsorption by clay and humic substances [25] [66]. Road de-icing salt contains two ligands chloride (Cl-) from NaCl and cyanide (CN-) from Ferro- 4- cyanide (Fe(CN)6 ) anti-caking agents has shown to increase Pd mobility with increasing the Ferro-cyanide concertation but the (Cl-) did not affect the Pd mobility while the mix of both shows elevated Pd mobility[69]. The author concluded that chloride can play an important role in this mobilization by increasing reaction rates [69], and similar results are found for Pt and + Pd in [70] using Ammonia (NH3) and ammonium (NH4 ) and the chloride. In both studies [69] [70]found that Pd release is greater than Pt release. This result indicates that no seasonal change on the solubility of Pt and Rh should occur in nature [20]. Increasing sulphur concentration in soil affects Pt solubility positively and it was also enhanced with time [25] [66]. No effect was observed for Rh. Thirty-five percent of the total Pd in road dust was shown to dissolve in a solution at pH 3, simulating rain, and a solubility gradient of Pd>Rh>Pt was observed[25]. In a solubility study on vehicle catalysts, was found the solubility rate was also faster for Pd than for Pt and Rh. Because of the degree of solubility observed for Pd, it was suggested that it is unlikely that Pd is in the metallic form, but may be present as chloride species [25]. The residence time for the road dust was less than 1 week; thus, the chemical conversion from a metallic state to a chloride species must be a relatively rapid process. The high solubility of Pd might imply a more effective transport in the aqueous environment[65].The solubility of PGE also depends on particle size. The smaller diameters of the platinum particles are more easily oxidized by oxygen present in the aqueous solution, because the activation energy necessary for oxidation is smaller for fine particles [25]. After a longer exposure the Pt particles on the surface of the aluminium oxide particles are dissolved in the organism and become bioavailable [25]. Microorganisms do not influence the dissolution and transformation of metallic catalyst- emitted Pt into bioavailable species for shorter exposure periods (up to 60 days). On the other hand, it was suggested that Pt in urban gully pot sediments occurring in the organic fraction is a result of bacterial action. Theoretically, Pt could be transformed into a methylated species in soil, although methylation has only been demonstrated under laboratory conditions. Methylation of Pt requires the presence of both the Pt(II) and Pt(IV) oxidation states[25]. These

27 platinum compounds can be methylated by methylcobalamin in vitro, forming a platinum chloride-methylcobalamin complex[25].

1.4. PGEs in environmental matrices

1.4.1. PGEs in air, soil, dust and vegetation (terrestrial environment) Figure 1.3. represents distribution pathways and concentration ranges of platinum group elements in selected environmental compartments. The PGE ratio (Pt/Pd; Pt/Rh; Pd/Rh) is an important indicator of the catalytic converter source [26, 71, 72]. Previously defined ranges are Pt/Rh 5–16 and Pt/Pd 1–2.5 and Pd/Rh 4–9 [73, 74]. There has been a gradual shift of Pt based catalysts to Pd usage [75] [26, 76] [77] in the late 1990s - since it is cheaper - and an introduction of only Pt, Pd–Rh, and Pd only or Pt–Pd–Rh catalysts. As a result, Pt/Pd and Pd/Rh ratios, derived from environmental samples analyzed, are non-uniform today [26, 76]. In road dust samples collected at increasing distance (0–10 m) from the road, the Pt/Rh ratio is relatively constant at ~ 7.1, while Pt/Pd ratios are more variable with a mean of 6.6 and a range between 2.0 and 26.6. This suggests that there is a significant difference in chemical behaviour between Pd and Rh. The Pt/Rh ratio varied between 4.6 and 5.6 in soil, pointing to automobile catalysts as the source, which indicates a relatively inert and immobile behaviour of Pt and Rh [25]. The Pt/Pd ratio in surface soil decreased with distance from the highway surface, which is further evidence of a higher environmental mobility of Pd compared to Pt [25]. Vertical profiles of PGEs in soils showed that Pd was found even in a depth of about 12–16 cm below the surface soil near a German highway, while Pt was only found in the surface layer. Comparing soil PGE concentration with plant PGE concentrations significant correlations could be found only for Pt and Rh, whereas grass Pd concentrations showed no relationship to the respective soil concentrations [77]. Traffic-related metals (Cr, Ni, Cu, Zn, Cd, Pb, As, Rb and Ba) were also enriched in the upper 20–25 cm layer[78]. Similarly to soil samples [79], the Pt/Rh ratio in run-off reservoir sediments remained relatively constant (4.6), and a strong correlation was found between Pt and Rh. These results are indicative for a common source, the catalytic converter, during the time interval 1987–1995 (correlated with 0–25 cm depth). Surface waters from run-off reservoirs along different highways gave the same coherence. A strong co-variance between Pd and Ni in seawater implies similar biogeochemical pathways[65]. Scanning laser ablation of road and river sediments revealed a coincidence of PGE and Ce peaks, indicating direct transport of PGE containing catalyst particles into the river. In river sediments, PGE remain associated to Ce particles, but part of the particulate PGE might be released from the Ce particle through formation of soluble PGE species or breakdown of the particle [25].

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Traffic derived trace elements (Ba, Ce, Cu, La, Mo, Ni, Pb, and Zn) are usually associated with high traffic densities, originating from exhaust emissions, tires, braking, vehicle and engine wear, and/or the re-suspension of road dusts [80]. The correlation of the PGE and Pb, Cu and Zn can be explained by the fact that Pb, Cu and Zn are metals known to be associated with motor vehicle pollution[65]. Copper is added as an antioxidant and may be present in the coarse fraction of the particles originating from the wearing down of engine bearings or other components. Zn is associated with the use of additives and lubricants [81]. The strong association between Mo and Ni may be due to the fact that these elements are used to coat cylinders of diesel vehicle motors [65]. A positive correlation between Pt and Pb (r = 0.436) but not very high, while no correlation between Pd and Pb was demonstrated[65]. Ba, La, and Ce are present as additives in catalytic converters. La and Zr oxides are added to the wash coat as stabilizers [81]. Ce, employed as a promoter in catalytic converters, is used in a Ce/Pt ratio of about 20– 100[25, 82]. Nevertheless, no correlation was observed between PGE and Ba, La and Ce [65, 74]. Reported concentrations of traffic derived trace elements and rare earth elements of interest in sediments and surface waters are available in the [83-89].

Table 1. 1. Heavy metal and Rare earth elements Concentrations Eleme referen nt La Ce Nd Eu Gd Dy Er Tm Yb Lu ce

Spain ;

river Odiel and 0.1- 0.03- 0.03- 0.06- 0.04- 0.03- 0.01- 0.04- 0.01- [90] µg water Tinto 0.1-38 /l 24.67 65.12 2.41 11.46 6.92 4.39 0.57 3.33 0.67 rivers

Suspen Spain ; ded Odiel and 1.36- 3.03- 1.53- 0.19- 0.44- 0.59- 0.77- 0.05- 0.06- [90] particu pg Tinto 0.09-8.6 late /g 96.31 65.61 40.18 2.38 12.37 10.54 11.78 1.55 1.34 rivers matter

Eleme nt Hf Cu Zn Pb Y Sr Rb Cd Pb Zn

MoÈlndal river Not µg saÊn in 1 upto Not sedime g- Gothenbu 0.05-1 10-500 50-1000 10-500 1-100 5 upto 50 0.01-5 reporte [91] nts 50 reported 1 rg d

24.86- river pp Not 1.84- Not Not Not 2.43- Not 1.92- China 0.1-2.64 [87] water b reported 95.41 reported reported reported 202.01 51.44 reported 706.25

Eleme nt Zr Mo Ga Ni Sc Cd Ge As Se river pp 0.09- 0.61- 0.15- 0.81- 1.67- 0.01- 0.04- 0.01- China 0.09-0.3 [87] water b 2.601 8.13 0.51 178.89 42.4 1.34 0.17 0.82

In soils, dusts and plants exposed to high-traffic density, the concentration of PGEs far exceeds the natural background levels. Hence, most of the studies have been involved in the sampling of soil and vegetation adjacent to the heavily travelled highways, and dust swept from the surface of roads [92]. An overview for PGE concentrations in soil and road dust is available in [20] in that study the highest concentration is Pt roadside soil in USA in 2001 was (144-339 ng/g) and Pt and Rh road dust in Madrid in 2003 was (144-339 ng/g) and (44-64 ng/g) , respectively. And in the recent studies in São Paulo top soils expressed in ng/g as Pd(3–378 ), Pt (1–208)and Rh(0.2–45) [93], Soils along a major highway in Germany Pd( 20-191 ), Pt (41-254 ) and Rh (7-36 )[94].

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Pt and Rh concentrations in externally contaminated grass growing 20 cm from a highway increased 3.5 and 9 times, respectively, from 1992 to 1997[25]. During the same time, the percentage of cars equipped with automobile catalysts increased from 30% to 75%[95]. In urban soil, Pt concentrations were six times higher in 2001 compared to 1992, correlating with the increased use of automobile catalysts [96]. Pt and Rh levels were close to four times higher in the upper soil profile close to a highway in 1996 compared to 1994[97].Pt levels were approximately three times higher in road dust sampled in 1991 compared to 1984, also correlating well with the increased use of automobile catalysts during that time [98]. Pt and Rh levels in urban road dust < 63 µm were approximately 3 and 1.5 times higher, respectively, in 1998 compared to 1991[98]. The Rh concentration, in the form of externally attached Rh containing particles, was significantly higher in feathers of raptors living after the introduction of automobile catalysts in 1986 compared to those living before[99], correlating with the increased Rh level in the environment since automobile catalysts were introduced [25]. Information on species-specific uptake, accumulation and effects on plants can be obtained after exposure of plants with selected PGE species. For instance, barley plants were grown in nutrient solutions containing either Pd(II), Pd-nanoparticles (Pd-NPs) or µm-sizes silica-particles decorated with Pd-NPs . Pd-NPs was visualized by transmission electron microscopy (TEM) in pollen tubes and barley plant sap, respectively[77]. However, TEM visualization cannot be studied in collected outdoor plants, because considerably lower PGE concentrations occur in the plants and so the detection by electron microscopy techniques is almost impossible[65, 77]. A PGE concentration gradient decreasing from roots > shoots > leaves > seeds was observed[77]. Anyway, plants are able to take up metals by their root system from soil or water, but also from their leaves’ surface. However, the mechanism of the latter uptake type is still unknown[77]. PGE levels in tree bark from urban sites in the U.K. were in the order of 1–3 ng g -1, which is of the same order of magnitude as in grass samples exposed to traffic. The Pt/Rh ratio was 1–3 and the Pt/Pd ratio was 1–1.7 [100]. These ratios are similar to ratios in airborne particles [99]. Platinum was taken up by grass growing adjacent to a U.S. highway. The Pt concentration was 1.2 ng g-1 in washed leaves and shoots, and Pd and Rh levels were 1.0 and 0.1 ng g-1, respectively [25]. The Pt/Pd ratio was 3.5 times lower in the grass than in the soil it was growing on, indicating a higher environmental mobility of Pd [25]. The highest concentrations of Pd (10.2 µg g−1) and Pt (14.6 µg g−1) were reported in wild carrots collected along a country highway with heavy traffic. For Rh, however, the highest value reported is about ten times lower (1.10 µg g−1) and was detected in reed collected along a riverside affected by massive urbanization. Rh was found to be the most soluble PGE in soil sorption studies, providing evidence for potentially high eco-toxicity of this element. Comparing the concentrations of Pd, Pt and Rh in all plant material collected within the last 5 years in the environment, Pd concentrations were generally highest. Pd compounds are known to be more soluble and more mobile in the environment than Pt compounds and therefore show a presumably higher bioavailability.

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The examination of food samples purchased from retail outlets in selected towns of the UK, revealed Pt and Rh concentrations below the limit of detection (LOD) of 0.5 ng.g−1 in all food groups, but palladium concentrations ranged between 0.23 ng.g−1 in green vegetables and 1.9 ng.g−1 in nuts. However, Pd contents in fresh fruits were below the LOD of 0.03 ng.g−1. Typical concentration ranges of platinum, palladium and rhodium in plant material collected in the environment is available in the [77].

Figure 1. 3. represents distribution pathways and concentration ranges of platinum group elements in selected environmental compartments including values reported until 2013 [77].

1.4.2. PGEs in rivers, coastal waters, and oceans (in the aquatic ecosystem). Aquatic ecosystems can be considered as an important sink of PGEs. Different sources, like road runoff or industrial effluents are directly discharged into aquatic ecosystems. A number of review articles on concentrations, routes and distribution of PGEs in aquatic systems are available [20] [101]. PGE concentrations in water samples most often near or below the analytical detection limits. The reported concentrations of Pd in river water samples vary between 0.4 and 13.1 ng L-1, from 0.006 to 2.6 ng L-1 for Pt and below the detection limit for Rh [102, 103] .

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Figure 1. 4. Literature data on PGE concentrations in estuaries, and marine environments. [101].

Adsorption on aquatic particles and estuarine behaviour There is little information on the adsorption and desorption behaviour of PGEs in the aquatic environment, though dissolved PGEs are not regarded as particularly particle reactive 24]. The affinity for particles follows the order Rh>Pt>Pd. In estuaries, Pt shows a non-conservative mixing which has also been demonstrated in laboratory experiments [65, 104]. Mixing of fresh water with saline water results in the reduction of the adsorption of Pt on particles and an increase in the fraction of dissolved Pt, similar to what is observed for Cd and Hg, which also form strong chloride complexes [104]. Only one publication reports the dissolved particulate distribution of Pt in estuaries (Gironde estuary) based on field data [105]. Removal of dissolved Pt was observed in the maximum turbidity zone under high water flow conditions, whereas in low water flow conditions a mid-estuary maximum may be related to local inputs. In laboratory studies, Turner et al (2015) [44] investigated the dissolved-particulate behaviour of Pt-based anticancer drugs and found an decrease in adsorption in the order cisplatin> carboplatin > oxaliplatin in river water predicted that during estuarine transport 35-45 % of cisplatin and carboplatin are removed in the estuary, but no more than 7% oxaliplatin is retained.

PGEs in seawater Pt concentrations decrease from coastal to open ocean areas. This cannot be observed for other PGE, as the data basis for the other PGE is too small. Platinum is markedly enriched over

32 palladium in seawater compared to cosmic crustal values [106]. This can be explained by the softer acidic character of Pt(II) relative to Pd(II) and the consequential stronger stabilization in seawater through complex formation of platinum with such ligands as chloride and bromide [106]. In addition, iridium and platinum are both enriched in some ferromanganese minerals compared to palladium. Platinum and iridium are probably oxidized during mineral formation from the divalent and trivalent states, respectively, to the tetravalent ones. By this mechanism, platinum and iridium can be accumulated over palladium, which only exists in the divalent state in the marine environment [106]. An increase of aqueous Pt and Pd concentrations with water depth in open oceans has been observed, indicating a nutrient like behaviour in open ocean [101]. In marine sediments, PGE concentrations were found in a wide concentration range (Figure 1.4.)[101]. For estuary and continental shelf samples, the concentration range in sediment samples resemble that of river systems. Similar to lake sediments depth profiles of Pt concentrations were taken for estuary sediments showing higher Pt concentration at the sediment surface compared to deeper areas indicating a relatively recent Pt discharge into estuarine systems [107].

1.5. Aquatic speciation of PGEs

PGEs can occur in various oxidation states, and in each oxidation state different complexes with inorganic and organic ligands can be formed depending on pH and ligand concentrations [108]. In natural waters, Pt can occur as Pt(II) and Pt(IV) and generally Pt(II) is the dominant Pt species. However, the stability of Pt (IV) increases and can be the dominant species in highly oxidized and highly saline waters under acidic to slightly alkaline pH (e.g., surface seawater). The most common oxidation state of Rh is (III) and (II) for Pd. As soft metal ions, Pt2+, Pt4+, Pd2+ and Rh3+ + form stronger complexes with soft ligands such as chloride, hydroxides, ammonia, sulfides, polysulfides, cyanide, and natural organic acids. The hard ligands such as 2 3- CO3 - and PO4 form only very weak PGE complexes. Numerous studies have been carried out to determine or estimate the thermodynamic constants for PGE complexes in aqueous solutions, however, few were determined under conditions that are relevant to natural waters (e.g., low temperature and low PGE concentrations), particularly to low-ionic strength freshwaters. Although, PGEs are known to complex with natural organic ligands, the stoichiometry and formation constants of these complexes have been poorly established. Furthermore, few data are available for Rh species in the aqueous solution.

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In freshwater, the thermodynamic data available for typical freshwater pH indicates that the − dominant form of Pt is Pt(OH)2 for Pt(II) over Pt(OH)5 for Pt(IV). At typical seawater pH 2− 2- (7.5–8.4) the dominant inorganic forms are Pt (IV) PtCl5(OH) and PtCl6 over Pt(II) 2− 2- PtCl4 and PtCl3(OH) [109, 110]. The predominant inorganic form of Pd in freshwater is expected to be the neutral hydroxide 2- 2- species Pd(OH)2. In seawater, the predominant forms are PdCl4 and PdCl3(OH) [25]. Stability constants of Pd(II) complexes are strongly correlated with those of Pt(II) and Pt(IV) complexes, but are generally smaller than the Pt(II) and Pt(IV) complexes [214]. Although there is some similarities in the equilibrium chemistries of Pd(II) and Pt(II) in seawater, there kinetic behaviors have been shown to be very different: Pt(II) and Pt(IV) complexes are kinetically inert relative to Pd(II) complexes and thus the complexation and adsorption reactions involving Pt(II) and Pt(IV) occur at a much slower rate than those involving Pd(II), making Pt (II) and Pt(IV) appearing less reactive than Pd(II). This may have a major implication in dispersion and biological uptake of PGEs in the aquatic environment [166]. There is a lack of reliable kinetic and thermodynamic data on the interactions of Rh(III) with inorganic and organic ligands. The available thermodynamic data predict the dominance of aqua-hydroxy, aqua-chloro-, mixed aqua-chloro complexes in the form of 3 – x – y [RhClx(OH)y(H2O)z] , where x + y ≤ 6 and z = 6 – x – y ; [111-114]. Rh(III) shows extremely slow reaction kinetics; equilibrium is often only achieved after several weeks [150]. - In freshwater at typical pH values of 6-8, both Rh(OH)3° and Rh(OH)4 are thermodynamically - stable; however, chromatographic data suggest that principally Rh(OH)4 is present [109, 110]. 2+ + - 2- 3- Positively [Rh(OH) , Rh(OH)2 ] and negatively,[ Rh(OH)4 , Rh(OH)5 and Rh(OH)6 charged species predominated under pH 4 and above pH 7, respectively[109, 110]. In seawater, -3 RhCl6 is expected but tentative speciation calculations predict the dominance of hydroxylated - 2- species Rh(OH)4 (H2O)2 and Rh(OH)5 (H2O) [109, 110]. The slow complexation kinetics of Pt and Rh result in a so-called “ageing” of solutions spiked with Pt and Rh, which means that over time the speciation of the PGEs in a solution will change from its original composition due to aquation and hydrolysis in which different aqua chloro- and aqua hydroxo-complexes are formed and these complexes are kinetically inert [111-114]. This is especially pronounced for Rh [109, 110], and not observed for Pd. As a result, the environmental mobility of the PGEs varies in the order Pd>Pt>>Rh.

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The kinetically inert complexes can be “activated” by addition of complexing ligands such as chlorides or chelating agents from natural organic matter. Dissolved natural organic matter, in particular fulvic acids (FA) and humic acids (HA) show a strong tendancy to bind PGEs, increasing their environmental mobility by orders of magnitude [115, 116]. The investigation of PGE interaction with model organic compounds and FA extracted from river waters showed that the solubility of platinum and palladium hydroxides increases markedly in the presence of organic matter [117, 118]. Based on the obtained data on the stability constants of platinum and palladium-hydroxy and fulvate complexes, thermodynamic calculations supported the dominance of hydroxy–fulvate species in weakly acidic and neutral media [117]. Koshcheeva et al. [115] observed a sharp increase (by two orders of magnitude) in rhodium hydroxide solubility at pH 7 in the presence of natural FA. In contrast to Pt in the pH range of 5 to 8, rhodium(III) is completely complexed with humic acid [116]. Furthermore, Pt [119, 120]and Rh [110, 116] are characterized by their slow kinetics of complexation and dissociation, so that the equilibrium speciation of Rh or Pt is not rapidly attained. As metals accumulate in lake sediments, sediment cores can be used as pollutant archives, and together with geological dating techniques can tell us the history of metal input into the system. Sediment cores of the Upper Mystik Lake in Boston concentrations layers dated prior to 1975 ranged between 0.5 and 2.3 ng/g for Pt, approx. 2 ng/g for Pd and 1 ng/g for Rh[121]. In sediment layers starting in the 1980s, concentrations of Pt, Pd and Rh increased to 29, 21 and 3 ng g-1, respectively. Levels of the three metals were positively correlated and ratios of Pt/Rh and Pd/Rh resembled those ratios which are found in automobile catalyst converters [122].

Table 1. 2. Ratios of PGE in sediment samples. Pt/Pd ratio Pt/Rh ratio Pd/Rh ratio Country Reference 0.5-2.8 2-17.5 2-12.5 UK [123] 1.38 - - UK [124] 1.9 - - Australia [125] 0.6-4.3 6.2-18.4 24-Feb UK [126] 1.2-1.6 6 3.7-4.5 [127]

Concentrations of PGEs in urban river sediments exceed rural levels and levels above 20ng/g can be considered as highly polluted [101] (Figure 1.3.).

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1.6. Toxicity and Health effects

There are few data on the toxicity of platinum metals to animals under natural conditions since most papers addressing this issue describe research performed in the laboratory. Basic knowledge of PGE levels in animal tissues and their toxicity is essential to assessing the potential risk of environmental contamination with PGEs to human health [23]. The toxicity of PGEs depends on their chemical form electronic structure and increases in the following order Rh (III) << Os (IV) ∼ Pt (IV) < Ir (III) ∼ Ru (II) ∼ Ru (III) < Pd(II) ∼ Pt (II) [23, 128, 129], which is corresponding to same mobility order [4]. It has been shown that some species of Pt, Rh, and Pd have high levels of biochemical activity, for example cytotoxicity and mutagenicity, and, because they are emitted in the form of fine suspended particulate matter from converters, they are be inhaled and accumulated in human lungs, leading to the risk of significant negative effects on human health after chronic exposure even at very low levels [130, 131]. Pt-based drugs and their metabolisms are mainly excreted via urine in their soluble form and have a significantly higher toxicological and cancerogenic impact than catalyst-born inorganic platinum [48] and the recent study on platinum compound drug could have the potential toxicity effect on aquatic organisms after dilution in urban sewerage and receiving water body [50]. Platinum drug compounds are classified as probably carcinogenic to humans (group 2A) by the International Agency for Research on Cancer (1987). The platinum of cytostatic agents is excreted by the patients and reaches the municipal sewer system. The environmental distributions and impacts of these drugs and their metabolites are, however, particularly poorly defined [132]. Bioaccumulation and transformation of Pt occur not only in the environment but also higher organisms, e.g., in the bowel of higher animals [20]. This can lead to enrichment along the food chain and elevated PGE concentrations in human and other mammalian tissues. Studies have shown that Pt compounds are highly cytotoxic, have mutagenic potential in microbial and mammalian cells and may trigger toxic and allergic reactions in humans [133- 135]. Among PGEs, platinum has the greatest detrimental effects on living organisms (its excess in the body is bound by proteins) [136]. Some Pt compounds can bind to N- and S- containing groups in proteins producing a reduction in enzymatic activity of enzymes [137]. While the impact of Pt compounds on ecosystem functioning of complex ecosystems at environmentally relevant concentration is lacking, studies suggest that Pt enrichment in urban environments may adversely affect ecosystems experiencing high Pt loads [20]. So have bioaccumulation experiments with plants and animals shown adverse effects on some of the tested species [20, 135]. Results on the uptake of Pt by Sphagnum showed significant decreases

36 in plant length and biomass as Pt concentration increased [135]. Histological and pathological analysis of the rats revealed vacuolization, eosinophil inclusion bodies in adrenal glands, shrinkage of glomeruli in the kidney, and enlargement of white pulp in the spleen [135]. In addition, in both species (Sphagnum and the rats) DNA damage was detected. This suggests that population-wide health issues in the human population may arise from elevated Pt concentrations accumulating along the food chain [35, 56, 138]. To date, toxicological effects of Pt in humans are generally confined to Pt-halide and -ammonium complexes, and to the anti- tumor agent cisplatin and its analogues [139]. They include sensitization, irritation of the respiratory tract, asthmatic symptoms and contact dermatitis. Chemotherapy treatment using cisplatin has given rise to severe nausea and vomiting, nephrotoxicity, with both tubular and glomerular lesions, ototoxicity, with tinnitus and hearing loss, as well as sensory peripheral neuropathy[140]. Toxicological data on Pd and Rh are scarce. For both PdCl2 and RhCl3 , increased tumour incidences were observed in a lifetime drinking-water carcinogenicity study with mice [38]. In contrast, metallic Pt, unless mobilized in bodily fluids, is considered non- toxic to humans [141].

1.7. Analytical methods for the determination of PGEs

1.7.1. Instrumental methods Investigating PGE content in environmental samples poses a big challenge because of the very low concentration levels of analytes, difficulties with quantitative dissolution of samples, lack of proper reference materials and unsatisfactory metrological characteristics of the analytical techniques used in the final determination step [142]. The inadequate detection power of the majority of instrumental analytical techniques puts severe constraints on the determination of these elements in environmental samples [143]. Separation and pre-concentration steps often precede the final detection of the metals [144]. Numerous interfering effects from matrix elements, in particular when examining complex environmental samples of extremely low PGE concentrations, may limit direct application of particular techniques [144]. The approaches quite often used are based on Adsorptive Stripping Voltammetry (ASV) [145, 146], electro- thermal atomization atomic absorption spectrometry (ETA–AAS)[147], neutron activation analysis (NAA) [148], inductively coupled plasma atomic emission spectrometry (ICP–AES) [149, 150] and inductively coupled plasma mass spectrometry (ICP–MS) [88, 143, 151-157], often in combination with demanding pre-concentration procedures [158-161]. However, Adsorptive Stripping voltametric methods for PGEs are extremely sensitive to organic matrices, requiring destroying the organic matrix to limit carbon content to below 0.1%

37 is required [162], in addition interferences and blank value problems in the determination of Pt by ASV limits this approach [163, 164]. Atomic absorption spectrometry employing flame atomizers does not cover the concentration ranges of interest, in particular when examining environmental materials. High matrix effects and small linear concentration ranges are additional disadvantages of the AAS technique [144]. High sensitivities (pg g−1 and sub-pg g−1 levels) available by neutron activation analysis (NAA) and a practical lack of blanks make the technique particularly suitable for the determination of ultra-traces of the elements in complex materials [165] but access to a radiation source to produce activated analyte nuclides and safety of analytical procedures used restrict the accessibility of NAA methods. Due to the very low detection limit, wide dynamic range, analysis of a large number of samples, and multi-element capability of ICP-MS, this method is considered as the most appropriate for direct determination of PGEs in various environmental samples [157]. Commercially available instrumentation fall into three groups, namely quadrupole (Q-ICPMS) based instruments, high- resolution sector field instruments (SF-ICPMS) and time-of-flight (TOF-ICPMS) systems[166], but the Q-ICPMS and SF-ICPMS are the most commonly used for environmental applications. SF-ICPMS will be used in this work and is described in detail in Chapter 2. 1.7.2. Analyte Pre-Concentration and Matrix Separation Methods Different pre-concentration/separation procedures have been developed and applied to the determination of PGEs by AAS, ICP-AES and ICP-MS techniques and are summarized in [150]. Generally, these procedures are based on either precipitation and co-precipitation [167]; liquid-liquid extraction [108] [167] [168] or solid phase extraction, either on anion- or cation- exchange resins or on chelating resins [169-173]. Only the solid phase extraction will be discussed here. Strong cation exchange resins + Containing sulfonic acid groups or the corresponding salts (Sulphonic acid: —SO3–H ). The chemical behavior of these resins is similar to that of a strong acid[174]. These resins are highly ionized in both the acid (R-SO3H) and salt (RSO3Na) form of the sulfonic acid group (-SO3H) [174]. They can convert a metal salt to the corresponding acid by the reaction:

(R-SO3H) + NiCl2 → (R-SO4) Ni + 2HCl [174]. The hydrogen and sodium forms of strong acid resins are highly dissociated, and the exchangeable Na+ and H+ are readily available for exchange over the entire pH range[174]. Consequently, the exchange capacity of strong acid resins is independent of the solution

38 pH[174]. SCRs remain ionized over the entire pH-scale and, thus, the effective capacity of the ion exchanger is practically pH independent [175]. Weak cation exchange resins Containing carboxylic acid groups or the corresponding salts ( —COOH) . These resins behave similarly to weak organic acids that are weakly dissociated [174]. The degree of dissociation of a weak acid resin is strongly influenced by the solution pH. Consequently, resin capacity depends in part on the solution pH [174]. A typical weak acid resin has limited capacity below a pH of 6.0, making it unsuitable for deionizing acidic metal finishing wastewater[174].Weak cation exchange resins are ionized only under high pH conditions. In hydrogen form, they remove preferentially divalent ions (e.g. calcium and magnesium) from solutions containing alkalinity. Strong anion exchange resins + – Containing quaternary ammonium groups (Quaternary ammonium:—N(CH3)3 OH ) [176, 177]. Strong base resins are highly ionized and can be used over the entire pH range [174]; remain ionized over the entire pH-scale and, thus, the effective capacity of the ion exchanger is practically pH independent[175]. Weak anion exchange resins + + + Containing ammonium chloride or hydroxide -N(CH3)2, -NH3 , -NH2 and -NH . Weak anion exchange resins are ionized only under low pH conditions, and the effective capacity of compound binding by these groups is progressively decreased by an increase in pH [174, 176, 178]. Hence, weak base resins exhibit minimum exchange capacity above a pH of 7.0. In the case of weak base exchangers, the process can be reversed by simply reducing the acid concentration, thereby deprotonating the base[177]. Chelating Resins A chelating resin is broadly classified as one type of ion exchange resins, but with an enhanced selectivity towards certain metals. The bonding of one metal ion with more than one electron donors within a functional group produces a coordination bonding with chelating effect [179, 180]. Generally the electron donor atom comes from the element groups V, VI, and VII in the periodic table, the most common ones being N, P, O, and S [181]. Higher bonding energy and stability are achieved with higher dentation number favouring ion adsorption[179]. The preference of specific metal–ligand bonding and stability depends on physicochemical properties such as stability/solubility constant of the formed complex, ion charge of the donor atom types and metal ion, electronegativity, geometry, ionic diameter [180] the number of electron donor atoms in a functional group [179] and the matching the radius of metal ion to

39 the size of the macrocycle cavity [182]. These properties are summarized under the Pearson Hard and Soft Acids and Bases (HSAB) theory. Thus, the ligands of the functional groups in resins behave as bases (i.e., electron donors) and the metal ions as acids (i.e., electron acceptors) following the Lewis theory of acids and bases. Hard acids are ions with low electronegativity, non-polarizable, have high charge to ionic radius ratio, and are attracted to hard bases of high electronegativity. In contrast, soft acids are polarizable ions with low charge to ionic radius ratio and with relatively high electronegativity that matches the relatively low electronegativity of soft bases[180, 181]. Hard acids favour the interaction with hard bases, soft acids favor soft bases, and intermediate or borderline acids can interact with either soft or hard bases complying with HSAB theory [180, 181]. The precious metals for instance palladium(II), gold(III), silver(I) ions in their low oxidation state belong to the “soft acids” and has a strong affinity for the “soft base” donor atoms in the order of S > N > O [182-194] . On the other hand, hard metal ions, for instance iron (III), copper (II) and aluminum (III), show affinity for hard bases with the donor atoms: O > N > S. Moreover, amino resins including N donor atoms as chelating resins possessing ionic interaction/ion exchange properties due to protonated amines. Hence, resins containing functional groups with soft base characteristics are used for the adsorption of metal ions with soft acid characteristics. For example, ion- exchange resins containing thiol groups (soft base) will present higher selectivity than resins with carboxylic groups (hard base) in the adsorption of gold ions (soft acids) from solutions containing iron species [195]. Likewise, to selectively extract hard acids (e.g., uranium), a resin with a hard base functional group (e.g., amino group) would be used [196, 197]. When high selectivity between two hard or soft acids (i.e., metal ions) is required, a base group containing the number of donor atoms similar to the oxidation state of the target acid ion is preferred (e.g., carboxylic groups with two electron donor atoms favour the interaction with metal ions with the valence of two (II) over three (III) [198]). Additional selectivity could be obtained depending on the special distribution of the functional group and the geometric configuration of the target species [199]. The partition coefficients of many metal ions are much higher on complexing resins than on the ion exchange ones so that a much lower free metal ion concentration can be determined using complexing resins. This allows to evaluate free metal concentration at extremely low levels in natural waters [200]. To improve selectivity, precipitation/coprecipitation or masking agents can be used, or pH or other parameters must be manipulated[168]. In general, the extraction capacity of chelating resin depends on the amount of chelating groups on the resin and readily affects the extraction conditions such as solution pH, flow rate, and concentration of coexisting ions.

40

The principal advantage of chelating ligands is their high selectivity for certain precious metal ions or a group of ions, whereas, their main drawback (for resins with very high affinities towards the metals) lies in the difficulty of recovering the retained metals. Elution is achieved through complex solutions containing organic compounds (e.g. Thiourea), at elevated temperature [201], or even by total destruction of the sorbent [108, 167, 172, 174, 176, 178, 192, 202, 203]. 1.7.3. Ion exchange and chelating resins for PGEs The most common procedure used is transforming the PGEs to their anionic chloride form in HCl; removal of bulk metals in cationic form with a cation exchange column, followed by pre- concentration of the PGEs on an anion-exchange column: + - n- - nB Cl + [MCl6] ↔[B]n[MCl6] + nCl [150, 178]. In this case, an excessively large column containing cation exchanger sufficient to completely adsorb large amounts of base metals is required [108]. The main obstacles in the application of cation exchange are: a). Relatively large amounts of resin necessary to absorb non-PGE metals [204]. b) Relatively large eluent volume necessary for quantitative elution of PGE from the column, high recovery of PGE is achieved when a mix of acids is used as eluent [204]. c) Low separation efficiency for Hf and Zr [204], Mo [108] which may give rise to many problems during the PGE determination step via ICP-MS. d) Hydrofluoric acid used for sample digestion may cause problems with separation of some interfering matrix elements by ion chromatography, e.g., Hf, which forms anionic fluoride complex behaving similarly to PGE complexes during the separation step [144]. e) Difficulties in separating Platinum group elements due to the limited amount of eluent used in order to avoid concurrent washing out of other substances from sorbent [204]. Complexing ligands (e.g. amines) may be added to the sample solution to improve the retention of positively charged complexes on such columns[108]. Alternatively, the ammine complex of palladium and platinum behaves as a cation and is quantitatively adsorbed by cation exchange resins but not by anion resins [204]. Anionic exchangers appear to be preferable to the cationic exchange resins, because they demand a column of a smaller size and smaller volumes of the eluates [204]. Selectivity of anion exchangers is better because of the formation of stable ion pairs between chloro complexes and a sorbent’s active groups [204]. The tendency for the metal-chloro complexes to form ion pairs 2- 2- 3- with anion-exchangers is: [MCl6] > [MCl4] >> [MCl6] > aquo species, where M is a metal [205, 206]. Various oxidizing agents were evaluated for conversion of PGEs from their lower oxidation states to higher oxidation states [206, 207]. The order is selectivity determined by the charge to size ratio, or charge density of the species[177]. Species with low charge density

41 are more easily paired than species with higher charge density[177]. Formation of hydroxy- complexes of a smaller charge leads to a decrease in the sorption efficiency on such resins[208]. Rhodium is retained to a much lesser extent .The reason is the higher kinetic inertness of rhodium chloride complexes as compared to platinum complexes, especially in weakly acid media, where neutral and cationic rhodium aqua complexes are present [Rh(H2O)3Cl3], + 2+ [Rh(H2O)4Cl2] , and [Rh(H2O)5Cl] [112]. For strong basic anion exchangers, their sorption ability towards Pt and Rh increases with the decrease in acidity of contacting solution, although their functional groups of quaternary ammonia base maintain equal sorption ability in a broad range of pH values. Probably it can be explained by competition between platinum and rhodium complex anions and chloride ions during ion exchange process [112, 178]. Correspondingly, the lesser the concentration of chloride ions is in the solution, the weaker is their competing effect [112, 178]. The problems with strong anion exchangers are: 1- resins do not selectively adsorb negatively charged Pt and Pd chloro-complexes. The elements that form anionic fluoride complexes during the digestion procedure, i.e. Ti, V, and Al, can also be enriched on such resins. 2- conversion of PGEs to higher oxidation states prior to addition of the sample to the anion exchange column, e.g. by chlorination [207], was proposed [108]. Formation of hydroxy complexes with a smaller charge at low pH leads to a decrease in the sorption efficiency on ion-exchange resins. In addition, the presence of different forms of analyte in the solution (e.g. at different oxidation states or hydrolysis products) may cause difficulties concerning complete separation of metals, as the higher oxidation states of the PGEs are more strongly sorbed by anion resins, whereas lower oxidation states are weakly bound. 3- Incomplete separation of Pt from geological samples was reported [108]. On the other hand, some of the Pt complexes are adsorbed too strongly to be eluted from the stationary phase[108]. 4- Weak sorbents were used to discriminate between strongly weakly bound metal species [9]. However, its disadvantage is that recovery is non-repeatable (it depends on sample type) and the method cannot be used for concurrent separation of all platinum group metals [204]. Quantitative recovery of noble metals has been achieved by some strong eluents [209] or mixtures of different solvents [204]. The problem of low recoveries of PGEs from strong anion-exchange resin was overcome using several circulations of thiourea (TU) solution (10 times, 60⁰C) [150] (thiourea in HBr >HCl >

HNO3 [144, 210]), or NH3 solution at 65 ⁰C or hot concentrated mineral acids [108], per- chloric [204], ammonium chloride and hydroxide [203], thiocarbamide [204] or potassium thiocyanate [211], ammonium thiocyanate [112] were proposed. Thiourea has long been known to form extractable complexes with both Pd [205] and Pt(IV)[212]. The thiourea ligand can

42 coordinate with either the ‘hard’ nitrogen atom, or ‘soft’ sulphur atom[212]. This indicates that acidic thiourea has stronger affinity for Pt and Pd than for Rh, leading to slower rate and lower overall elution recovery of Rh from all three resins [205]. The problem of low recoveries of

PGEs with HNO3 from strong anion-exchange resin was overcome by using the isotope dilution technique in ICP-MS determination [108]. However, hot nitric acid also removes refractory metals from the resin and partly destroys the resin, which may cause additional unidentified interferences (formed by organic fragments of the resin) concerning the determination of small amounts of Pd and Pt [108]. 2.1.3.1. Resins with active functional groups based on sulphur used for noble metals Strong cation-exchanger The most common type of functional group in commercial resins based on sulphur is the strong – + + cation exchanger with sulfonic group (R-SO3) , having either H or Na as the mobile ion [174]. This group of resins shows the strongest adsorption capacity among cation exchangers, but their high strength results in poor metal selectivity and difficulties for elution of the adsorbed metals. Chelating resins with thiol- groups (R-SH)

The chelating ion exchangers of functional thiol and methylene thiol groups with the commercial names: Chelite S[186], Duolite GT-73 [213, 214], Imac GT-73, Duolite GT-74, Purolite S924, Spheron Thiol 1000 and Tulsion CH-97 are widely applied in sorption and separation of noble metal ions [169, 215, 216]. Thiol groups (R-SH), also known as mercaptans, have special affinity for Hg [217]. Resins with thiol groups are classified as strong cationic resin with chelating properties at low solution pH, and as soft Lewis bases showing high adsorption preferences toward soft acid (e.g., Hg, Cu+, Ag+, Au+, and some PGEs). The metal ion selectivity of thiol groups reported with Ambersep GT74 resin follows the order Hg2+ > Ag+ > Cu2+ > Pb2+ > Cd2+ > Ni2+ > Co2+ > Fe3+ > Ca2+ > Na+, as suggested by its manufacturer[181]. Employing the low capacity of thiol groups for base-metal adsorption. Duolite GT73 used to separate Pd (II) and Au (III) from solutions containing Cu (II) and Ni (II) in chloride and nitrate systems at low pH (pH < 2)[195]. The results obtained the extraction selectivity of Au3+ > Pd2+ > Cu2+ > Ni2+ [186]. A potential method to concentrate and recover palladium from chloride solutions with low Pd2+ concentrations (i.e., 0.0011 M of Pd2+) was proposed using the resins Duolite GT73 and Chellite S [186]. In other study described the relationship between the extent of base-metal extraction and pH values, where mild acidic conditions are preferred [218]. The high affinity of thiol groups with precious metals also

43 represents its main drawback as the adsorbed metal is difficult to recuperate during the elution process [182]. Thus, large amounts of organic eluents are required to achieve metal recuperation [195] or hot aqua regia [219] or acidified Thiourea [220]. Purolite S924 is a chelating resin with thiol group, polystyrene based and designed for the selective removal of mercury. Even so the high selectivity for metals such as silver, copper, lead, cadmium, nickel and cobalt, makes this resin useful in waste treatment and hydrometallurgical processes. The high selectivity for mercury is largely unaffected by high chloride or sulphate content of the effluent and operate at wide pH range (1-11). Hence, Thiol functional group of S924 allows higher selectivity for soft metal acids including Hg, Pd and Au over the base metals, makes this functional group an important choice for possible application for the DGT. Chelating resins with thiourea and isothiouria functional groups The ion exchangers of the functional thiourea groups, commercially known as Tulsion CH-95, D405-II, QuadraPure TU, Srafion NMRR, Lewatit TP-214 [186] and Purolite S914 [221] and isothiourea groups, commercially known as XUS 43600.00 [205], Purolite S920[186, 220], are widely used for the concentration of PGEs [190]. The two N atoms a long with S of the thione form coordination complexes with metals ions such as Cd, Zn, Hg and PGEs. Depending on medium’s pH and a form of the functional group, the isothiourea groups occur in form (free base), they create coordination bonds with metal ions [186]. When the groups occur in form (acid conjugated), platinum metal anion complexes 2– are bonded according to the anion exchange mechanism, e.g. [PdCl4] . In the in-between media both mechanisms are assumed to compete with each other[215].

SH S

R N NH R N NH2

H H Thione form Thiol form

Thiourea resin R NH + R NH H /X 2 - S S C + +X

NH 2 NH2 acid conjugated form Free base form isothioureum resin

44

Similar to thiol groups, resins based on thiourea and its conjugate isothiourea are classified as soft Lewis bases and strong cationic resins with chelating properties at low pH. As suggested by the manufacturer, thiourea groups contained in Lewatit TP 214 show a metal selectivity in the order of Hg2+ > Ag+ > Au+ = Au3+ > Pt2+ = Pt4+ > Cu2+ > Pb2+ = Pb4+ > Bi2+ > Sn2+ = Zn2+ > Cd2+ > Ni2+. However, Thiourea and isothiourea groups show higher adsorption capacities than thiol groups when adsorbing mercury, gold, silver, and PGE ions [181, 186]. Comparison study between thiol and Thiourea observed higher adsorption of PGE ions using Thiourea based resins (i.e., Lewatit TP 214 and Purolite S 920) than using resins containing thiol groups (i.e., Chelite S and Duolite GT 73)[186]. Similarly, the adsorption of PGM ions was observed to be higher in resins with isothiourea groups (i.e., XUS 43600) than in resins containing quaternary ammonium groups (Lewatit MonoPlus M600) or polyamine groups (i.e., Purolite S958;[205]). Alike resins containing thiol groups, neither resins containing thiourea nor resins containing isothiourea have satisfactory performance in adsorbing base metal ions[222]. Palladium (II) sorption on the chelating resins (Lewatit TP 214 and Purolite S 920) may be described as [170, 186, 205]: • Ionic interaction mechanism between the protonated amines and the chloro palladium(II) complexes: + − (푅1푅2)NH + H퐶푙푎푞 → (푅1푅2)푁퐻2 퐶푙푠표푟푏 + − 2− + 2− − 2(푅1푅2)N퐻2 퐶푙푠표푟푏 + [Pd퐶푙4]푎푞 → [(푅1푅2)푁퐻2 ]2[푃푑퐶푙4 ]푠표푟푏 + 2퐶푙푎푞 • Chelation (coordination) mechanism: + − 2− − + − (푅1푅2)N퐻2 퐶푙푠표푟푏 + [Pd퐶푙4]푎푞 → (푅1푅2)푁퐻 → [푃푑퐶푙3 ]푠표푟푏 + 퐻푎푞 + 2퐶푙푎푞 2− − − (푅1푅2)C=S + [Pd퐶푙4]푎푞 → (푅1푅2)퐶 = 푠 → [푃푑퐶푙3 ]푠표푟푏 + 퐶푙푎푞 • Or anion-exchange mechanism:

R R NH2 NH2 2- - 2- S C + [PdCl ] S C + X + [PdCl4] 4

NH NH2 2 2 2− • In the sorption reaction the resin sorbs the anionic Pd(II) complex (Pd퐶푙4 ) by the anion-exchange mechanism whereas during the elution, palladium is released from the resin when the negatively charged chlorides in the Pd(II) complexes are replaced by neutral Thiourea ligands and the Pd(II) complex is positively charged

45

R NH2 R NH2 Ligand substitution 2- + [PdCl ] - S C 4 2 S C + Cl + palladium (II) Thiourea (NH ) CS/ HCl complex NH 2 2 2 2 NH2

Industrial applications of Purolite S914 include mercury removal from brine and effluent in chloralkali process, mercury removal from flue gas scrubber effluent, recovery of platinum group metals from effluents amongst others. Purolite S920 is designed for the selective removal of mercury and for the recovery of precious metals from the industrial effluents. These properties are largely unaffected by high chloride (or sulphate) content of the effluent. Hence, Thiourea functional group in S914 and its derivative in S920 allow higher selectivity for precious metals, makes this functional group an important choice for possible application for the DGT. 2.1.3.2. Resins with active functional groups based on phosphorous used for noble metals Phosphorous-containing groups are less available in the market comparing with nitrogen- or sulfur-containing groups resins. Resins based on phosphorous-containing groups operate by chelating interactions in a wide range of pH. However, phosphorous-containing groups show hydrophobic characteristics that affect the adsorption kinetics. To enhance the reaction kinetics, other functional groups are included, e.g., sulfonic groups in the commercial resin Diphonix [197], to reduce hydrophobicity. Phosphonic and aminophosphonic groups are the most common phosphorous groups present in commercial resins. Phosphonic groups have strong, stable, and specific chelating interaction with polyvalent metals due to four possible donor atoms with ligand potential. Special preference for actinides in tri-, tetra-, and hexavalent oxidation states was reported in [223] resins containing diphosphonic groups (i.e., actinide resin from Eichrom Technologies; a.k.a. AC resin or DIPEX). The combination of sulfonic–phosphonic groups also shows preference for actinides, and an additional adsorption potential for transition metals (e.g., Cu2+, Co2+, and Zn2+) at pH ranging from 5 to 8 without any specific preference[224]. When compared with amidoxime (Purolite S910), aminophosphonic (Purolite S940), and quaternary ammonium groups (Purolite A520E), the combination of phosphonic and sulfonic groups in Purolite S957 helped extract higher amounts of uranium from seawaters entering desalination plants [225]. Owing to the strong

46 bonding generated, phosphonic-sulphonic resins (e.g., Diphonix) are used to control water pollution near uranium mining operations [197]. The strong interaction between phosphonic groups and metal ions results in difficult elution processes, requiring expensive elutants and highly specific and costly equipment with a general reduced metal recovery rate[226]. Commercial resins with phosphinic groups (e.g., Daion CRP-200 and Lewatit TP 272) show metal selectivity in the order of Fe3+ > V4+ > Zn2+ > Al3+ > Cu2+ > Mn2+ > Co2+ > Mg2+ > Ca2+ > Ni2+[226]. Phosphoric groups selectively chelate bivalent metal ions (e.g., Zn2+, Cd2+, Cu2+, Co2+, and Ni2+) [227]. Phosphine groups are classified as soft Lewis bases with adsorption preferences for soft acids (e.g., Hg, Ag, Au, and PGEs)[228]. Phosphines and its derivatives are used as ligands in coordination chemistry, and as scavengers in solvent extraction of metals. For instance, tri-n-octylphosphine oxide (TOPO) is not only used in solvent extraction procedures in the recovery of PGE [229, 230], but also for the selective removal of PGE from acidic media [231]. Due to the essentially identical functional groups for binding, MPX-310 could be roughly described as an enhanced version of “TOPO on a solid support” [201]. MPX-310 [201, 232, 233] and MPX-317 are two macroporous resins with a polyacrylic backbone and phosphine oxide and phosphine oxide thiourea coordinating functionalities, prepared and provided by ItalmatchMagpie Polymers (starting with mid-2017, Magpie Polymers is a part of Italmatch Chemicals, Italy) [201, 234]. In strongly acidic and oxidizing solutions, containing high concentrations of metals and other ions that prohibit sustained application of ion exchangers or of oxidation-sensitive materials, MPX-310 displays excellent selectivity for Pt, Pd and Rh [201] over base metals including Cu. In the same study, the phosphine oxide functional group of MPX-310 shows lower Cu accumulation compared to Thiourea functional group in TP-214 resin, which might be explained by the sulphur containing functional group which has high affinity for Cu. MPX-317 show high selectivity for Pt and Pd [221] and shows high adsorption capacity of Cd compared to MPX- 310, which can be explained by the grafting of the thiourea groups, which have a strong affinity for Cd(II), on the polyacrylic backbone [234]. MP-317 is polymer holding the same functional group of the resin MPX-317, which is the chelate phosphine Thiourea [221]. Commercially, the micrometric powder polymer (0.1–10 µm diameter) polymer containing amino-phosphine oxide groups as complexing group MP-102 developed by Italmatch (Magpie Polymers) for selective palladium sorption [189, 221, 235] by ligand exchange mechanisms on amine or on phosphine oxide groups [189] from acidic effluents containing base metals such as nickel and copper. Total Pd elution was achieved via acidified Thiourea (1M TU, 0.01M HNO3) at condition(T = 333 K and 18 h of contact time) [189], in other study the total elution of the

47 accumulated PGEs achieved via hot aqua regia procedure at 70°C [221]. MP-101 is polymers containing stable alkylphosphines [235], provide a unique functionality for the capture and recycling of transition metals. In all studies [189, 201, 232] used MPX-317, MPX-310 and MP-102, the commercial resins were tested at mg L-1 level concentration but they were not evaluated at ng L-1. The efficacy of ion exchange resins mainly depends upon their physical properties including their particle size [174], and for some extent to the concentration of the target element [112, 205]. Hence, the sorption selectivity of these resins at low centration was not tested. MP101 is a polymer containing stable alkylphosphines [235] produced by Magpie Polymers - Italmatch Chemicals and the polymer readily captured a variety of metals including cadmium, chromium, nickel, lead, palladium and uranium, at different metal concentrations and pH.

Phosphine oxides MP-102 structure [189]

Alkylphosphines MP-101 structure [235]

MPX-310 and Cd+2 complex [234]

MPX-317 and Cd+2 complex [234]

Phosphonic and sulfonic groups in polymeric resin Diphonix

48

Phosphoric group of D2EHPA present in Lewatit VP OC 1026

Phosphinic group of DTMPPA present in Lewatit TP 272

Phosphine group of TOPO Figure 1. 5. Simplified chemical structure of Magpie Polymers - Italmatch Chemicals MP-102, MP-101 polymers as well as MPX-317 and MPX-310 complex with Cd and phosphorous-containing resins from other producers (Diphonix, Lewatit VP OC 1026 and Lewatit TP-272).

In principle, phosphorous-containing groups show high selectivity for soft acids including PGEs but they are very few studies using P-type functional group for PGEs. Hence, it is very interesting to investigate their possibility use in DGT for PGEs uptake.

2.1.3.3. Resins with active functional groups based on nitrogen used for noble metals Resins with primary, secondary, and tertiary amino groups are behaving as weakly basic (weak anion exchangers). Sorbents with quaternary ammonium groups exhibit intermediate basic properties, while sorbents with nitrogen heterocycle functional groups are strongly basic anion exchangers [236]. Some N-type resins for Au, Pd, and Pt are working on anion exchange mechanism. Several brands of sorbents (Duolite, Amberlite, Dowex, etc.) with primary, secondary, tertiary, and quaternary amine functional groups are available [228, 236]. A study [176] using strong basic anion-exchange resins Type 1(commercial name: Varion ATM with functional group of trialkyl ammonium chloride ) and Type 2 (commercial name: Varion ADM with functional group of dialkyl 2-hydroxyethyl ammonium chloride), as well as weakly basic anion-exchange resins with Tertiary amine functional group for selective Pd(II) uptake in acidic conditions (0.1- 6M HCl). The study concluded that Varion resin are suitable for palladium(II) metal recovery from acidic solution, and the highest selectivity towards Pd(II) showed the weak basic anion exchange Varion ADAM.

49

Study [237] focused on sorption of Pd(II) complexes from acidic solutions with sodium chloride addition using strong base anion exchanges resins (Quaternary ammonium, type 1) with the polyacrylic skeleton (Amberlite IRA-458 gel and Amberlite IRA-958) are not suitable for removal of Pd(II) ions. It is a resultant of the aliphatic skeleton as well as the sieve effect. In another study [238], focused on sorption of Pd(II) complexes from acidic solutions with sodium chloride addition using the resins Varion ADAM (Tertiary amine weakly basic anion exchanger; polyacrylic matrix) for Pd(II) retention from the chloride solutions (HCl–NaCl) in comparison with Amberlite IRA-92 (Secondary amine weakly basic anion exchanger; polystyrene), Amberlite IRA-95 (Polyamine weakly basic anion exchanger; polystyrene- divinylbenzene), Amberlite IRA-96 (Tertiary amine weakly basic anion exchanger; polystyrene-divinylbenzene), Dowex 66 ( Tertiary amine weakly basic anion exchanger; polystyrene-divinylbenzene) and Lewatit MP-62 (Tertiary amine weakly basic anion exchanger ; polystyrene-divinylbenzene) was evaluated. Under the batch experimental conditions based on the decrease of capacities, the resins can be presented in the following order: IRA-96 > Dowex 66 > IRA-95 > IRA-92 = ADAM > MP-62 (solutions with the 1.0 M NaCl addition) and Dowex 66 > IRA-96 > IRA-95 > ADAM > MP-62 (solutions with the 2.0 M NaCl addition). When acidity increased from 0.1 – 2M HCl at same level of NaCl, the capacity followed Varion ADAM > Lewatit MP-62 ≈ Amberlite IRA-92 > Dowex 66 > Amberlite IRA-95 > Amberlite IRA-96. The matrix effect in the case of the Varion ADAM resin plays a less significant role in sorption properties than the mechanism of binding ; i.e. Varion ADAM resin are its polyacrylic matrix, more hydrophilic (weaker van der Waals type attraction) matrix than that of the other weak base anion resins – styrenedivinylbenzene (stronger van der Waals type attraction) and higher bead size span. This article [238] is one of few articles dealing with the effect of the resin matrix (polyacrylic, polysteren…., etc.) on the uptake. The authors, the highest working anion exchange resins are the resultant of: (1) resin’s matrix, (2) size of resin beads, and (3) binding mechanism of Pd(II). Another article discussing the matrix effect is in [182]. A group of new functional sorbents on poly-vinylbenzyl chloride–acrylonitrile–divinylbenzene matrix functionalized with aliphatic mono-, di-, and polyamines have been synthesized[239]. Recoveries over 95% were achieved for noble metals (Au(III), Pt(IV), Pd(II)) from multicomponent solutions. Purolite S985 and Purolite A500 contain polyamine and quaternary ammonium bases functional groups, respectively, are evaluated for sorption recovery of PGEs in the presence of some non-ferrous metal ions and iron (III) [113, 114, 240, 241]. The Purolite A 500 shows lower selectivity for Rh and higher adsorption for the base

50 metals in comparison to S985[113]. In another study[112], different N-type resins namely the Purolite S985 (polyamine), Purolite S 500 (quaternary ammonia base), Purolite A 530 (quaternary ammonia base), Av-17-8 (quaternary ammonia base), AN-251 (tertiary amino- groups), AN-82-10P (hexamethyleneimine) and AM-2B (secondary amino-groups, tertiary amino-groups, quaternary ammonia base) were fully compared for sorption Rh and Pt sorption at different HCl concentration (0.001 – 4 M). The anion exchangers investigated can be arranged according to their sorption ability in the following orders: a) for Pt in the presence of Rh — AM-2B > Purolite A 500 > Purolite S985 > AN-251 > AV-17-8 > Purolite A 530 > AN- 82-10P; b) for Rh in the presence of Pt — Purolite S985 > AM-2B > AN-251 > Purolite A 500 > AV-17-8 > AN-82-10P > Purolite A 530. In the same study when the test solutions are stored for 3 months “ageing effect”, the sorption of Pt and Rh is dropped with the order S985>A 500> AM-2B. The comparison of the polyamine functional group in Purolite S985 and the isothiourea functional group in XUS 43600.00 for Pt(IV), Pd(II) and Rh(III) present in chloride solution collected from leaching platinum group metal containing spent automotive catalyst [205]. The weak-base Thiouronium functionalised XUS 43600.00 chelating resin exhibits significantly greater adsorption selectivity for chloride complexes of Pt (IV) and Pd (II) in acidic chloride solutions followed by the Lewatit M+ MP 600 (Quaternary ammonium, type 2) while the Purolite S985 shows greater adsorption for Rh (III) of both resin. Other resins based on the nitrogen for selective adsorption of PGEs are fully discussed in [236]. Purolite S985 is a high capacity, macroporous, weak base anion exchange resin with a polyacrylic matrix supporting functional groups of the polyamine type for the removal of trace heavy metals from waste water streams and the special polyamine functionality produces very interesting operating capacities and makes the uptake of metallic cations possible even when they are present in the waste stream as organic anionic complexes. Hence, polyamine functional group in S985 over other N-type functional group shows interesting sorption ability for PGEs over other N-type functional groups, makes this functional group an important choice for possible application for the DGT.

51

1.8. In situ monitoring and dynamic speciation measurements in solution using DGT

1.8.1. Introduction

Basically, there are few approaches for the environmental monitoring and measurement of water. The first is grab sampling involve the collection of discrete samples using survey vessels with subsequent storage and laboratory analysis. Though widely used, this approach is associated with a number of commonly acknowledged drawbacks, including high cost and does not provide large data sets for spatial distributions or temporal evolutions studies and the fact that grab sampling gives only a snapshot of the water status in the investigated water. Often environmental water samples are difficult to access (e.g., sediment pore-waters, groundwater), and contaminations (chemical changes) or perturbations (physical changes) may occur during sampling. Problems often encountered are oxygen contamination after exposition to the atmosphere, metal contaminations, pressure changes affecting the gaseous species when samples are collected in depth. Temperature changes may modify the chemical speciation, especially when collecting samples from hydrothermal systems. This may result in systematic errors, in the species determination, which may lead to misinterpretation of the studied system. Consequently, high frequency or even continuous in situ monitoring is required, necessitating the use of reliable, accurate, and sensitive field-deployable instrumentation [200, 242, 243]. Another approach used to monitor environmental contaminants is biomonitoring. Biomonitoring is an integrative technique, and a number of species can be used depending on the environment being investigated. Analysis of contaminants that accumulate in organisms’ tissues can give an indication of the environmental concentration levels. However, the use of living organisms as samplers has some limitations. Firstly, there are several physiological, anatomical, and behavioral factors such as metabolism, depuration rates, excretion, stress, and viability that can affect the rate of bioaccumulation in organisms’ tissues and, consequently, the determined levels. In addition, only species that do not migrate are useful in biomonitoring in order to reflect the contaminant concentrations at a specific location and their variation during the sampling period. Finally, the extraction procedure from animal tissues for instrumental analysis is particularly complex [242]. An alternative approach in water monitoring is to measure the analyte immediately after the sampling (on-site but off line), which eliminates most of the issues associated with sample storage. If the analysis is done on-line, continuously or sequentially, this allows for close to real time mapping of spatial and temporal variations of the analyte. In-field type analysis is

52 often performed using traditional laboratory techniques, though sometimes modified and adapted to conditions in the field. The third approach is sampling by in situ measurements [242], which refers to analyses performed directly in the environmental compartment of interest (i.e. at the desired time, depth and location). This avoids most of the issues with the sampling, where changes in light, temperature, pressure and redox conditions may compromise the sample. In situ techniques can be divided into three distinct groups, one of which is continuous in situ sampling. This group of in situ techniques comprises electrodes that provide a continuous response to analyte concentrations in the water; examples include pH and ion selective electrodes. The second group contains techniques that provide series of in situ discrete measurements, including voltammetric and flow injection analysis techniques. In the last group, fractionation and accumulation of the analyte occurs in situ, but the analysis of the accumulated fraction is carried out in a subsequent step at the laboratory. Passive sampling techniques have been used for the determination of a wide range of analytes in various applications in air, water and soil for almost three decades[244], where the analytes are trapped in a medium within the sampler i.e., the receiving phase which can be a solvent, chemical reagent, or a porous adsorbent [242]. Thus, passive sampling methods are based on either adsorption or absorption of pollutants from water. Adsorptive methods are based on the physical or chemical sorption of analytes on surfaces and depend on surface binding and surface area. Absorptive methods include analyte distribution on the receiving phase and depend mainly on the polarity of the pollutant and the solvent phase and the solubility of the pollutant in water and the solvent [242]. Passive sampling uses the principles of mass transport across a diffusion layer. It was an attempt to identify and codify the factors controlling uptake rate, in the application of Fick’s law of diffusion [245]. These fundamental laws were stated by the German physiologist Adolf Eugene Fick in 1855 [246]. In the aquatic environment passive sampling has been used to determine concentrations, fluxes and lability of metals[247-249], anionic species[250-252], a wide range of organic pollutants[253, 254] (including pharmaceuticals [255] and endocrine disruptors [256]), as well as organo-metallic compounds [257]. Passive sampling has many advantages. First of all, passive samplers are relatively small and easy to use, so it is practical to carry them to sampling sites located far from the laboratory [242]. Another advantage is that passive sampling combines several methodological steps such as sampling, isolation, and pre-concentration. In this case, analytical cost decreases since sample pretreatment is not necessary and decomposition of analytes declines since they are

53 isolated in the passive sampler. As the analytes are pre - concentrated (on a resin gel layer in DGT), this results in an increased sensitivity. Moreover, at every sampling site, only one device is needed as long as the sampling lasts, and therefore, there is no need for large number of samples[242]. Continuous deployment can provide a long term in situ monitoring tool. One major advantage of passive sampling as a technique is its inherent specificity towards the analyte of interest [258]. Generally, a passive sampler device will only sample a fraction of the total analyte present; freely dissolved species and labile complexes as well as conjugated species [259]. More specifically, this means those species that would dissociate within the timescale of transport across the diffusion pathway of the sampling device, and that have a stability constant lower than the stability constant of the compound formed as a result of the binding to the samplers receiving phase [259]. The fraction accumulated by passive sampling reflects the analyte’s behavior in the investigated environment, yielding valuable information not only on its content but also on its chemical status (the different species present, speciation), thereby contributing to the more accurate assessment of the environmental impact of the analyte [260, 261] (e.g. the metal concentrations assessed with a passive sampler correlates to the biologically relevant fraction of the metal in the studied environment) [259, 261]. 1.8.2. Principles of the passive sampler The term “passive sampler” covers several distinct subgroups. These can be classified according to the sampling medium (gaseous or aqueous), operating in different accumulation regimes (equilibrium or kinetic) [242, 261, 262] and the target class of analyte (organic or inorganic)[244]. In equilibrium passive sampling (Figure 1.6) [242, 262], as the name suggests, the analyte(s) are accumulated in the device until the concentration in the sampler is in equilibrium with the bulk concentration. The method requirements are that the response time has to be known in order to reach stable concentrations and must be shorter than the analyte’s variations in the concentration being determined [242]. In addition, the environmental conditions cannot affect the analyte concentrations in the sampler. Equilibrium passive samplers are characterized by a fast attainment of equilibrium between analyte concentrations in sampled medium and passive sampler. This type of passive sampler is typically used to provide a snap shot of the labile analyte concentration at the moment of sampling, although in practice there is a response lag time before equilibrium is reached if there is a change in concentration. For this reason, these samplers cannot estimate time-weight-average concentrations [242, 262].

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Figure 1. 6. Passive sampling devices that operate in different accumulation régimes [263].

Equilibrium models are sufficiently well established and parametrized that they are used routinely for describing the distribution of chemical species in waters. They are essential to understanding and predicting solubility and sorption processes. They also form the basis for the free ion activity model (FIAM) and biotic ligand model (BLM) approaches to predicting uptake of metals by biota. However, this equilibrium condition is often not achieved in nature and the rate of dissociation of metal complexes may be important. For example, if membrane uptake of metal is fast, biological uptake will not depend simply on the free ion activity. The free metal ion is still presumed to be transferred across the membrane and there is a dynamic equilibration with the complexes, with association and dissociation continually occurring. However, depletion of the concentration of the free ion at the membrane surface diminishes the rate of complex association, resulting in a net dissociation of free metal ions from complexes. This component of complexed metal also contributes to the accumulated metal. To appreciate fully the uptake of metals in natural waters it is necessary to have information on the kinetics of dissociation of the metal complexes present [264] . The kinetic passive sampling (time-integrated sampling) [242, 262] never reach equilibrium during sampling period, and the analyte continues to be transferred constantly in the receiving phase; thus, the concentration between the sampled medium and the receiving phase follows a linear relationship [242]. Kinetic passive samplers are in some ways a special case of equilibrium passive samplers where the sampling medium has been chosen so that the water- sampler partition coefficient is large, and/or by assuring a large capacity of the receiving medium. Another difference is that it is generally desirable that the mass flux between the sampler and the bulk water compartment is slowed down, so that the time to equilibrium

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(saturation) is sufficiently long to allow extended sampling in the kinetic regime [258]. The techniques based on kinetic passive sampling are conceptually similar, even though there are some exceptions. Examples of kinetic passive samplers used for inorganic/ organic analytes includes the diffusive gradients in thin-films (DGT) [265, 266] , etc. 1.8.3. Principles of the Diffusive Gradients in Thin-films (DGT) technique The Diffusive Gradients in Thin-films (DGT) technique was invented by the researchers Hao Zhang and Bill Davison at Lancaster University in 1994 [267, 268]. DGT relies on the quantitative diffusive transport of solutes across a well-defined gradient in concentration, typically established within a layer of hydrogel and outer filter membrane [266]. Once diffusing through these outer layers, solutes are irreversibly removed at the back side of the diffusive gel by a selective binding agent with a strong affinity for the analyte and a large capacity, thereby effectively creating a perfect planar sink[269]. Typically, Chelex 100 resin is used as binding resin for most divalent and trivalent metal ions., which is immobilized in a second layer of hydrogel. The hydrogels used in DGT are commonly made of polyacrylamide, which can be fabricated with a range of properties, including almost unimpeded diffusion due to the gel having water content as high as 95% [270]. The GT assembly is shown in Figure 1.7.

DGT cap

0.45µm membrane Diffusive gel Binding gel

DGT base

Figure 1. 7. DGT assembly

Normally it is assumed that a) there are no interactions between the diffusing species and the medium of the diffusive layer, b) the receiving phase maintains the concentration at the interface at effectively zero [248, 271, 272] and c) the adsorption of the analyte species occurs in a plane sheet. The assumptions made in a, b and c have been shown to hold for the most common condition encountered [269, 273, 274]. The accumulation curve for a device in the kinetic phase consists of a linear section and a non-linear section, where the accumulation rate decrease, to eventually reach zero when equilibrium/ saturation is reached (Figure1.8).

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Figure 1. 8. Schematic representation of a concentration gradient through a DGT device at steady state.

When a steady-state gradient (F does not depend on time) is established that can be exploited mathematically to measure in-situ concentrations, and the principles of how the DGT operates: the flux of ions through the hydrogel is controlled by Fick’s first law of diffusion (equation 1.1): dc F = D Eq.2.1 dx

The diffusion coefficient measured at infinite dilution and a reference temperature of 25⁰C, D0, can be corrected to any in situ temperature, Dt, by applying the Stokes–Einstein equation, where

T0 and Tt are in Kelvin (K) equation 1.1.: 퐷 ɳ 퐷 ɳ 0 0 = 푡 푡 Eq. 1.2 푇0 푇푡

The viscosity of water can be expressed by the equation 1.3[275] where ɳ0 is the viscosity of water at the reference temperature of 25⁰C and ɳt is at the in situ temperature t(⁰C):

1.37023(푇−25)+8.36×10−4(푇−25)2 273+푇 퐿표𝑔퐷 = + log (퐷 ( )) Eq. 1.3 푇 109+푇 25 298

At steady-state, the concentration gradient, dC/dx, is the difference between the concentration in the bulk solution and the concentration at the interface between the diffusive and resin gel layers,퐶́, which is 0 if the resin-gel layer is a rapid and effective sink. The distance x is the diffusional path length, which is the combined thickness of the diffusive gel layer, the

57 protective membrane filter and the diffusive boundary layer in solution, DBL. Here we simplify the system by assuming that the thickness of the DBL is negligible and the diffusion coefficient for the other two layers, with a combined thickness of g, is the same:

(C−퐶́) F = D Eq. 1.4 ∆g

The flux (F) is equal to the mass (M, in g) of metal through an area (A, in cm2) per unit time (t, in s): 푀 퐹 = Eq. 1.5 퐴푡

The mass of metal, M, accumulated by the resin-gel layer is calculated after placing it into a known volume of elution acid (Ve). VGel, the volume of the gel layer, is generally taken to be

0.16 mL for an 8 mm thick gel and fe is the elution efficiency: C (V +V ) M = D 푒 푔 푔 Eq. 1.6 f푒

Substituting for F in equation 1.5, the bulk solution concentration as measured by DGT (CDGT) can be calculated by the following equation 1.6: ∆g M C = Eq. 1.7 퐷퐺푡 DAt

1.8.4. Components of the DGT device 1.8.4.1. Membrane filter A membrane filter protects both gels from colloids and particulate matter and holds all underlying layers packed. Its main role is to protect the DGT device from particulate matter and more generally from potential physical harm. Depending on the manufacturer, the thickness of these filters typically varies from 0.13 to 0.15 mm [276, 277]. The pore size of the filters is generally 0.45 µm, preventing any particulate matter to enter the probe. If the underlying layers were to be disrupted during the measurement, the DGT results would be void of any meaning. The filter is usually made of cellulose nitrate and cellulose phosphate, although polyethersulphone membranes have also been suggested if deployment times exceed 24 hours [276, 277]. Durapore membrane filter (HVLP pore size 0.45 µm, diameter 2.5 cm, thickness 120µm) used during this research. 1.8.4.2. Diffusive gel layer and interactions of solutes In kinetic passive sampling, it is generally desirable to have an ion-permeable hydrogel, often referred to as the diffusive gel or layer of well-defined thickness to lessen the impact of variations in water turbulence and to regulate and slow down the diffusion or supply of ions

58 to the resin gel layer, by creating a diffusive gradient [259]. The diffusion limiting layer can also have other functions, such as to exclude analyte species that are too large to pass through the pores, and to reduce the sampler sensitivity to variations in turbulence [259] _ figure 2.9_. The diffusive layer has ranging thickness (0.16-2.0 mm)[264]. A stable dimension is reached for a given ionic strength, pH and temperature [277]. Three types of gel have commonly been used in DGT, agarose, polyacrylamide cross-linked with an agarose derivative (APA) and polyacrylamide cross-linked with bis-acrylamide also called restricted gel as it has a pore size of 1nm compared to 10nm for APA gel[277]. In all cases, transport of solutes occurs solely by diffusion[277]. The common used diffusive gel in DGT is APA, Agarose is a linear polysaccharide obtained from marine red algae, by removing the agaropectin constituent. Although agarose mainly consists of repeating units of agarobiose (d-galactose and 3,6- anhydro-lgalactopyranose), The agarose gel comprises sites bearing a resemblance to pyruvic and sulfonic[278], ester sulphate, ketal pyruvate and carboxyl [277] functional groups, which may dissociate to provide a negative charge. The purity of the material determines the extent of these groups and the precise properties of the gel. When a hot solution of agarose in water is cooled, the molecular strands associate to form double helixes. The resulting three- dimensional network entraps water to form a weakly ionic gel. There is a wide-ranging distribution of pore sizes from 1 to 900 nm, depending on the agarose concentration. Agarose gels used for DGT measurements have been prepared from solutions containing 1.5% agarose. This produces a gel with pore sizes ranging from 1 to 480 nm, with average values of 35–47 nm [278]. The large pore size and dominance of free water allows relatively free diffusion of simple solutes. Ions in solution can in principle interact with gels and membranes in two ways. They can specifically bind to sites such as the amides within polyacrylamide. Like any chemical binding, the affinity can vary greatly between ions. Ions can also interact electrostatically with fixed charge groups in gels and membranes[277]. Within agarose gel there are sites of agarobiose substituted pyruvic acid, which may dissociate to provide a negative charge[278]. Cations from solution will electrostatically associate with these negative charges, raising cgel/csoln above 1. Consequently the flux to the binding layer is increased, causing cDGT to be more than it would be if the gel was uncharged. However, at high ionic strength, the attraction of the locally available major cations for the negatively charged sites minimizes the effect of the electrostatic interaction for the analyte ions means that cgel/csoln is dependent on ionic strength, and approaches 1 at high ionic strengths[277]. There is no evidence for adsorption of fulvic acid to agarose. In its presence, positive values of cgel/csoln for

59 metals, associated with specific binding, decrease, due to the competitive binding of the metal ions with fulvic acid in solution[277]. There are two features of diffusive layer that are considered influential for the DGT investigation. Firstly, the knowledge about overall effective rates of dissociation of metals from their complexes could be obtained by varying the thickness of the diffusive layer used in DGT samplers. Secondly, by manipulating the diffusional properties (i.e. the size of pores) of a diffusive gel, it is possible to assess the distribution of the metal between inorganic and organic complexes (assuming that both inorganic and organic metal complexes are fully labile). The thickness of the diffusive layer, Δg, affects the analyte flux and as a consequence, it has strong impact on the lability information of the measurements. The DGT technique allows the measurement of metal complexes that are labile enough to dissociate within the DGT timescale. Therefore, the information about the kinetic constants of metal complexes can be obtained by the means of the kinetic window of the DGT sampler (i.e. the thickness of the diffusion layer). A DGT sampler with a thicker diffusive layer allows for more dissociation of the complexes, thus their contribution to the mass of the metal accumulated on the binding phase will be significant. The DGT sampler with a thinner diffusive layer will tend towards the marginalization of a less labile species in the DGT measurement (Figure 1.9) [264, 279, 280]. .

Figure 1. 9. Schematic view of labile and non-labile trace element compounds diffusing into a DGT probe. Pore-size is about 10 nm for Open-Pored and 1 nm for Restricted-Pored hydrogels [259].

1.8.4.3. Resin layer Also known as binding layer; consisted of a hydrogel (agarose gel or polyacrylamide) impregnated with the resins or sorbents selective for desired analyte. The role of the binding layer is a rapid and irreversible accumulation of metals and their pre-concentration. There are several features and considerations of the binding phase that need to be recognized and

60 determined in terms of the accuracy of the DGT technique. One of the most important features of a binding phase the resin layer acts as a perfect planar sink, implying that the binding to the resin is strong, irreversible, and almost instantaneous and that the accumulated metal amount is well below the capacity of the resin[269]. The second important is its effective capacity for metal binding. In general, the effective capacity of the binding phase gel disc is determined in a single element solution, however it is important to acknowledge that the co-adsorbing analytes affect the capacity of the binding phase. This in turn, will have significant impact on the estimation of an optimal DGT field deployment time. Other parameters such as pH or ionic strength also contribute to the saturation of the binding phase due to the competition towards the binding sites of an adsorbent [281, 282]. To some extent the capacity of the binding phase gel might be extended by increasing the amount of adsorbent encapsulated within the hydrogel. Nonetheless, attention needs to be paid as the additional amount of adsorbent might negatively affect the properties and the robustness of the binding layer [221, 283]. The most commonly used resin gel incorporates Chelex®-100 resin with functional groups of iminodiacetate acid (IDA), which act as chelators for polyvalent with polyvalent cations and show a high affinity for metal ions, especially with divalent metal ions, but Chelex resins are unsuitable for elements forming oxyanion species (e.g. As, Mo), organometal species (e.g. methylmercury) and they have a low affinity for alkali- and earth alkali elements. Many other resins and sorbents have been used such as Diphonix for U[284, 285], MnO2 for Ra [286], 3- mercaptopropyl functionalized silica gel for Hg [287], Metsorb, ferrihydrite or ZrO for elements such as As, P, Mo, V[288]. For the PGEs the DGT technique has not yet been developed. Several resins are available which have a high selectivity towards PGEs. These include ion-exchange resins with polyamine [112-114, 205, 240, 241, 289], thiol [186], thiourea [186, 201, 215], isothioureum [186, 220], phosphine oxide [189, 201, 232] functional groups. Nevertheless, until today there is no investigation on the selectivity order of any DGT binding phases that includes Pt, Pd and Rh as the analyte of interest. 1.8.4.4. Diffusion boundary layer When a solid and a liquid phase are brought in contact with each other, a thin layer of laminar flow will be generated between the two phases, in which transport is entirely diffusion controlled. This layer is called the diffusive boundary layer (DBL) and will be formed whenever a DGT piston is plunged into solution, the DGT device acting as the solid phase and the aqueous environment as the liquid phase. Investigations into this approach have shown that in reasonably well stirred conditions (laminar velocity

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> 2 cm s-1) if the diffusion limiting layer is sufficiently thick (0.8 mm), the DBL will be less than 0.2 mm thick, and δ can be disregarded without significant loss of accuracy [197, 258, 274].

1.8.5. Parameters influencing DGT measurement 1.8.5.1. Temperature As the analyte is adsorbed on the receiving phase the local analyte concentration is lowered and a concentration gradient is established. The accumulation rate is limited by the speed of the analyte diffusion through the diffusion pathway, which is described by the diffusion coefficient, D (m2 s-1), and by the total length of the diffusion pathway. The diffusion coefficient is described theoretically by the Stokes-Einstein equation: 퐾 푇 퐷 = 푏 3휋µ푑

-23 -1 where Kb is the Boltzmann constant (1.38 x 10 J K ), T is the temperature (K), μ is the dynamic viscosity (g s-1 m-1) and d is the ionic diameter of the analyte (m). The direct relation between temperature T and diffusion coefficient D allows the following interpretation: as the temperature gradually rises the viscosity of the solution will drop, allowing the ions to move more freely in the solution, hence augmenting the diffusion rate and diffusion coefficient. Zhang and Davison verified that this equation holds truth in the range of 5 to 35°C[281], and has been extended to 60°C [267]. Therefore, other effects associated with possible size changes, such as expansion or contraction of the moulding, which could affect the area of the window, and the gel and filter thickness, appear to be negligible [267]. Another study conducted by Larner et al.[290] suggested that diffusion coefficient measurements should be redone if the temperature is around the freezing point, since it led to a 12% deviation from the supposed value for many metals, compared to their value at 20°C. If the temperature range does not exceed 5°C it is probably acceptable to use the mean temperature for field application, as the error in CDGT will most probably still be within 10% [267]. Daily and seasonal temperature fluctuations can have significant effect on the uptake of the analytes as temperature can also increase solubility and desorption of contaminants from suspended particles, thus increasing the dissolved fraction of the analytes[242]. 1.8.5.2. Ionic strength The application of the DGT technique in aquatic environments that can be characterized by a low[291] and high [292-294] ionic strength has been also explored. There has been much contention in the literature over the use of DGT in low ionic strength solutions of < 1 mmol L-

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1, with some studies reporting enhanced solute diffusion rates or poor DGT measurement precision[266]. This can be explained by diffusive- and binding phase-related way. Diffusive related way: Enhanced metal accumulation on the binding phase layer has been reported due to the interactions between the functional groups of a diffusive layer and metal cations [266, 274, 291]. If the polyacrylamide diffusive gel is not exhaustively washed, it has negatively charged functional groups that might bind metal cations, which in consequence leads to increased metal accumulation. Binding phase-related way: The co-adsorbing analytes might compete for the binding sites of the binding phase layer and for the binding with the analyte of interest. This in consequence will lead to the rearrangement of the selectivity order of the binding phase under those specific conditions and will negatively influence the binding phase strength. It might also reduce the effective capacity of the binding phase. The low ionic strength will also influence the accumulation of the partially labile metal complexes on Chelex®-100 in similar fashion as in case of pH. At high ionic strength, the electrostatic effects of negatively charged IDA functional groups of Chelex®-100 resin will be counterbalanced by the background electrolyte, but at low ionic strength, the metal accumulation on the binding phase gel will be affected by those interactions. The partially labile metal complexes that continue to dissociate within a binding phase layer, might be attracted or repulsed due to electrostatic effects of the binding phase gel [274, 295] 1.8.5.3. pH The pH limitation of the DGT technique is interrelated to both, diffusive and binding phase gel layer. Some authors reported that at pH below 1 and over 11, the hydrogel starts to swell, what might affect its properties [270, 281] . pH changes can influence the surface charge on adsorbent particles as well as the ionization potential of chemicals, leading to a change in the performance of DGT resins [296]. Although the range of the DGT technique is relatively wide, the use of a correct binding agent is necessary. Every resin has a specific range for every chemical element and will become more ill-suited the further they are used outside of this range. Diffusive gels and resin gels will be impacted by pH, some more than others, depending on the chemical components used in their fabrication and thus the functional groups present on the resin. This can be illustrated with Chelex since the IDA functional groups will become protonated at lower pH values, impairing its chelating efficiency. The exact pH at which decreased adsorption appears varies between systems and depends on the selectivity of IDA[297, 298]. The iminodiacetate (IDA) functional groups of the Chelex®-100 resin are pH- dependent, thus their form will change over the pH

63 range. At pH range 2-4, the cationic form of the resin prevails, at pH range 4-7.4 the iminodiacetate functional groups are in zwitterionic form and at pH over 7.4 the IDA functional groups are in anionic form[299]. The accumulation of free metal ion and fully labile complexes, irrespective of their charge, will not be affected by the form of the functional groups of a binding phase, because they are fully dissociated at the binding phase layer interface. However, in case of the partially labile metal complexes, which do not fully dissociate within the diffusive layer and can penetrate the resin layer, their accumulation might be obstructed by the opposite charge of the IDA groups of Chelex®-100 [295, 300, 301] . Moreover, at low pH, a proton competition for the binding sites of Chelex®-100 resin might occur and by the same token, it might lead to rearranging the selectivity order and reducing the effective strength of the resin [281, 302]. Furthermore, increasing or decreasing the pH above critical values will cause the gels to swell and the direct change in physical characteristics may cause a misinterpretation of results. 1.8.5.4. Competition of other ions and ligands Competition effects arise when the accumulation of a metal on a resin is dependent on the concentration of other metals present. This has been observed for Mn in the presence of Mg (decrease in the binding rate of Mn), as well as Mn at higher ionic strength (decrease in stability constant). Other competition effects include bicarbonate competition with the measurement of phosphate on ferrihydrite and competition is observed when the maximum capacity of the binding layer approaches, such as the competition between Fe and Mn, and Sr and Ca on Chelex. At high ionic strength, capacity issues becomes a problem for major elements such as Ca and Mg. 1.8.5.5. Uptake capacity The pre-requested of the binding layer is to act as a perfect planar sink, implying that the binding to the resin is strong, irreversible, almost instantaneous and that the accumulated metal amount is well below the capacity of the resin [269]. Using a non-selective resin will cause the resin gel to attain saturation much faster and is unwanted, pointing out once more the need for a very selective resin during measurements [303]. 1.8.5.6. Flow velocity A water flow rate above the 0.02 m s-1 threshold has no influence whatsoever on measurements; provided the thickness of the diffusive gel is > than 0.7mm. Decreasing flow to the point of stagnant water may severely impact results, causing up to a 50% decline of ion accumulation [304]. In this case the DBL must be taken into account and quantified using DGT devices with different diffusive gel thickness.

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1.8.5.7. Dissolved Organic Matter on DGT Measurements Natural organic matter (NOM), such as humic and fulvic acids, proteins and polysaccharides, − − −2 − − − and inorganic ligands such as F , C1 , SO4 , HCO3 , CO3 , and HPO3 are among the potentially important anions that may reduce free metal ion concentrations via complexation hence they affect the DGT measurements. This also explains, at least in part, how aquatic alkalinity, pH, and salinity can modulate toxicity: decreases in pH and alkalinity favors reduced complexation and higher amounts of free metal ions and therefore toxicity is increased. Increased salinity on the other hand results in greater Cl− complexation of free metal ions and therefore mitigation of toxicity. Dissolved natural organic matter (NOM) and other organic compounds can also complex free metal ions. Of these ligands, fulvic acids (FA) are generally complexed first, followed by proteins, oxides, polysaccharides and finally simple inorganic ligands [258]. NOMs are complex and heterogeneous compounds that arise from the degradation of terrestrial or aquatic biota and can present a number of different negatively charged ligands. For example, metal binding can be via carboxyl, phenolic, sulphidic, and protein groups on the NOM. NOM complexation capacity can vary with both the concentration of NOM (measured as dissolved organic carbon) as well as the relative composition of binding groups (i.e., its quality) [261]. The order in which sites in complexing agents are saturated can be understood by considering the free energy of complex formation, expressed through the standard equation for Gibbs free energy: [푀퐿] ∆퐺° = −푅푇 ln K=−푅푇 ln ( ) ([푀][퐿]) where M is the metal and L is representing any ligand complexing site. On the continuous scale of free energy, sites with the lowest ΔG° will be saturated first, i.e. strongly complexing fulvic acid sites, followed by weaker fulvic acid sites, and so on [305, 306]. 1.8.5.8. Biofouling The biofouling process refers to the progressive accumulation of the flora and/or fauna adhesion or growth on the active sample surface [295, 304, 307]. The biofouling on a DGT sampler has been observed after 21 days [282, 308] and 10 days [309-311], and in some cases even after 5 days [312, 313] of the DGT deployment. Biofilm is composed of bacteria, algae, fungi, and extracellular polymers resulting from cell metabolism. These various elements interact with trace metals through physical, chemical, and biological processes in the water. Biofilm growth at the surface of the samplers may affect the thickness of the diffusion layer and/or modify the diffusion coefficient [314]. Firstly, the algal films might be considered as an extension of the DBL, which limit molecular diffusion of the analytes. It is also possible that

65 the biofilm may bind the analytes, thus effectively removes them from the bulk solution, which leads to the underestimation of the DGT-labile concentration [283, 304, 307]. Similar effect might have a precipitation of minerals (i.e. HFO) or aggregation of colloidal material on the DGT filter membrane [315, 316]. The effect of the biofilm on the accuracy of labile concentrations measurement is element dependent and depend on the role of physicochemical interactions between metals and the biofilm[314]. Alternative filter types (Nucleopore membrane or silver-based filters,…etc) could be a solution to prevent the biofouling [307]. 1.8.6. Applications of the DGT technique 1.8.6.1. DGT as a monitoring tool DGT is a low cost and user-friendly device, provides in situ pre-concentrated samples and allows access to time-weighted average concentration (TWAC). The cost of deployment and retrieval is similar to a typical water quality analysis. The simplicity of DGT relies on 2 aspects: it does not require specifically trained personnel and it combines several methodological steps such as sampling, isolation, and pre-concentration [242]. In this case, analytical cost decreases since sample pre-treatment is not necessary and decomposition of analytes declines since they are isolated in the passive sampler. An explanation of the basic principles suffices for any future user. It is also non-mechanical, requires thus no power and little maintenance[317]. Moreover, at every sampling site, only one device is needed as long as the sampling lasts, and therefore, there is no need for large number of samples[242]. The TWAC responds to all concentration changes during deployment and integrates temporal environmental variability, it provides a more-representative measurement of solute concentrations compared to grab sampling. Moreover, as analyte species are continuously accumulated within the passive sampling device and eluted into a small volume, the concentration in the measured extract will usually be higher than grab-sample concentrations, reducing the uncertainty of analytical determination and improving detection limits [243, 318]. Field and experimental data from Munksgaard and Parry (2003) [319] showed that DGT devices provided adequate detection limits, accuracy and precision for monitoring of near-pristine levels of labile Mn, Co, Cu, Cd and Pb in north Australian coastal seawater when deployed for periods of 3 days. Ryan et al. (2003) [309] evaluated the use of DGT as a monitoring tool for Ni, Cu, Pb, and Zn concentrations at various locations in the Gold Coast Broadwater, a subtropical, coastal lagoon estuary in southeast Queensland, Australia. DGT measurements were compared with measurements made on conventional sampling system. Commonly the conventional sampling approach to characterise the trace metal concentrations in estuarine waters is time-consuming and logistically expensive, requiring hourly sampling across tidal and diurnal cycles at least one occasion in each major season [320]. Here DGT

66 has shown a great potential as a monitoring tool, particularly for dynamic estuarine waters, because of its ability to continually accumulate analytes and to provide a genuine, time-integrated average measurement. Recently, a large European Interreg project was launched, the MONITOOL Project [321] (https://www.monitoolproject.eu/), aiming at performing a large scale interlaboratory comparison on the use of DGT passive samplers in coastal waters of the Northern Atlantic for the measurement of priority pollutants including cadmium, nickel, and lead, within the framework of Water Framework Directive 2013/39/EU. The goal is to establish Environmental Quality Standards (EQS) based on the DGT labile fraction.

1.8.6.2. DGT as a predictor of bioavailability of elements

The DGT technique has been promoted as a surrogate of metal bioavailability. This has been supported by a vast number of investigations, that correlated metal concentration in biota with the DGT measured concentration. The empirical evidence that supports this relationship includes plant, invertebrates and vertebrates in water, sediment and soil. Lately, Zhang and Davison [15] elaborated the use of DGT as a predictor of metal bioavailability. Briefly, DGT mimics metal uptake by plant under diffusion-limited conditions, because of a striking parallel between a modus operandi of DGT and the diffusion-limited plant uptake, which was also extrapolated to (micro)organisms (i.e. invertebrates). During the diffusion-limited uptake, metal accumulation by biota occurs very rapid in comparison to the rate of supply of metal from the aquatic environment. Just as within the diffusive layer of DGT, the concentration gradient is established and the free metal concentration at the cell membrane-aquatic environment is effectively zero. Therefore, metal species that are labile enough to dissociate within the cell membrane, will contribute to the metal flux to biota.

1.9. Aims of the research and structure of the thesis

As we have shown in the literature overview, due to the problems associated with the analysis, there a very few data on the distribution of PGEs in aquatic systems, leading to a gap in our knowledge on their environmental fate, transformations and potential toxicity to aquatic organisms as well as their bioaccumulation and potential human health risk. For Belgium, no data exist on the concentrations of PGEs in rivers, hospital effluents and effluents of WWTPs. To understand their riverine transport, estuarine behavior and inputs to coastal waters, we first need to develop simple and reliable analytical methods to make their determination possible. Despite the numerous developments in the DGT technique and the increased use of PGEs in the past decades; the DGT method has never been developed for PGEs. The complexity of the chemistry of PGEs in solution; instability of PGE solution; speciation changes in stored PGE

67 solutions (ageing of solutions); together with the complexity in analytical procedures, interferences in ICPMS measurements and the lack of readily available resins produced for analytical purposes and variability in the blanks of industrial resins, are some of the factors which complicate the development. The aim of this work was thus to develop the Diffusive Gradients in Thin-Films (DGT) technique for the measurement of platinum (Pt), palladium (Pd), and rhodium (Rh); investigate its applicability as a monitoring tool to determine time weighted average concentrations of PGEs in urban rivers and hospital effluents and increase our knowledge on the sources and behaviour of PGEs in urban rivers. The thesis is divided in 8 Chapters Chapter 1. Background and Literature Study: This chapter describes the chemistry, physical properties and the occurrence of PGEs in the nature. It discusses the applications of the PGEs that led to spread the PGEs into the environment, the environmental transport, distribution and transformation and temporal trends of PGEs and health effects. This chapter provides an extended theoretical description of the passive sampling techniques and its advantages over the traditional methods. A detailed literature study was performed to select appropriate resins for the DGT development of PGEs. The research aims and structure e of the thesis is explained. Chapter 2. Methodology This chapter describes the PGE analysis by ICP-SFMS as well as the DGT protocols used. The determination of PGEs species in environmental samples is highly challenging due to problems associated with chemistry of PGEs in solution during sample handling and storage and problems related to the analysis itself for their ultra-trace concentration and the extremely high concentrations of the interfering elements. This chapter describes all the aspects related to analysis by ICPMS and interferences, stability etc. Results of the stability tests and interference tests are presented. The DGT procedures for making diffusive and resin gels, preparation of deployment solutions, use of diffusive gel etc is provided. Chapter 3. Development of DGT for PGEs. The DGT development for PGE using the chelating resins Purolite S920, S914, S985, Ionoquest MPX 317 and MP102 is described. Gel preparation and elution were optimized; the linear accumulation in function of time was evaluated, effective diffusion coefficients in spiked fresh water and seawater were determined as well as the evaluation of the influence of pH, salinity and organic on the DGT response. The results of the study were published in Talanta 203 (2019): 34-48. Additional resins which were investigated but not included in the manuscript are added in the Annex.

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Chapter 4. Influence of ageing of solutions of PGEs on DGT performance. Over time the speciation of the Pt and Rh in a spiked solution will change from its original spike composition due to hydrolysis an aquation and the equilibrium speciation change of Rh or Pt is not rapidly attained. The influence of the ageing of solutions of Pt and Rh on the affinity towards the resins and DGT response was evaluated showing that inert Rh species Rh(OH)3(s) could be formed that are not DGT available; but that the inert species can be solubilized by addition of organic matter. The results of the study were published in Journal of Analytical Atomic Spectrometry 36(4), DOI: 10.1039/D0JA00442A. Chapter 5. Elimination of Interferences in the Determination of Platinum, Palladium and Rhodium by Diffusive Gradients in Thin-Films (DGT) and Inductively Coupled Plasma Mass Spectrometry (ICP MS) using selective elution. The chelating may also accumulate other elements such (Cu, Zn, Pb, etc.) which will simultaneously be extracted by the hot aqua regia extraction and interfere with the ICPMS analysis. As the affinity of the resin for the interfering elements is lower than for the PGEs, selective extractions were investigated to release the interfering elements without loss of PGEs. Knowledge of the HSAB behavior of the PGEs helped us to develop DGT with high specificity for PGEs and develop selective elution procedures for the elimination of interferences. The results of the study were published in Talanta Volume 223, Part 2, 1 February 2021, 121771 doi.org/10.1016/j.talanta.2020.121771 Chapter 6. Pre-concentration of PGEs from solution on ion -exchange and chelating resins. In this section, we compared the Column and Batch process for separation of PGEs using resins with different functional group. In this section, we show the affinity of these resins for selective uptake of the PGEs in high acidic condition of around 1M HCl which also ensures the stability of PGEs during the surface water sampling. Chapter 7. Distribution of platinum (Pt), palladium (Pd), and rhodium (Rh) in urban tributaries of the Scheldt River assessed by Diffusive Gradients in Thin-Films Technique (DGT). This chapter describes the distribution of PGEs in two urban rivers with varying anthropogenic inputs: The Zenne River and the Marque River, both tributaries of the Scheldt River. In addition, a first interlaboratory intercomparison for the measurement of PGEs by DGT was performed. The data show that these elements can be useful tools to trace pollution sources and dispersion. The results of the study were published in Science of the Total Environment Volume 784, 2021, 147075 Chapter 8. General conclusions and future perspectives

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The Annex contain other resins evaluated for DGT but not included in the publications

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Chapter 2: Methodology

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2.1 General Procedures

All chemicals were of analytical reagent grade or greater. High purity nitric acid (HNO3) and hydrochloric acid (HCl) were used in all experiments (Fisher Scientific, Trace metal grade). Standard solutions used were Rhodium ICP Standard 1000 mg L-1 in 6% HCl (SIGMA -1 -1 ALDRICH), Palladium 1000 mg L in 2% HNO3 and Platinum 1000 mg L in 2% HCl ICP- MS ULTRA grade(TM) Standard (Ultra Scientific, North Kingstown, England). Deionized water (Milli-Q Advantage with Element Pod, Merck Millipore, USA), named Milli-Q hereafter, was used for the preparation of the solutions, gels and cleaning glassware and containers.

As natural water samples filtered seawater (North Sea, Belgian coast) and filtered river water (Zenne River, Belgium) was used. Artificial seawater was prepared using following composition: 27.5g NaCl, 6.78g MgSO4 H2O, 5.38g MgCl2.6H2O, 1.4g CaCl2.2H2O, 0.72g

KCl and 0.2g NaHCO3 per litre. Commercially available sea salt (Instant Ocean, Aquarium Systems, France) was used in a later stage and was also used to prepare to low salinity solutions (0.03 M sea salt). The humic acid stock solution was prepared by dissolving humic acid sodium salt (Sigma Aldrich) in 2 M NaOH, dilution with Milli-Q water and adjusting the pH between (6.5-7) with 0.5M HCl.

All experiments were carried out in a clean room. All equipment used was acid washed in 5%

(v/v) distilled HNO3 and rinsed with Milli-Q water and then dried in a laminar flow hood (class- 100) in a clean room before storage in cleaned polythene bags. All handling of equipment and samples was with polythene gloves. Temperature and pH measurements were performed using pH probe (WTW GmbH, Germany) and monitored during the experiments. The DGT deployment solutions were well-mixed using a mechanic stirring system, so the diffusive boundary layer (DBL) of DGT was considered negligible [197].

2.2. Analysis of PGEs with ICPMS

2.2.1. Principles of ICP-MS measurements Inductively coupled mass spectrometry (ICP-MS) is the most widely used technique for the analysis of PGEs in environmental samples, either using a quadrupole mass spectrometer (ICP- Q-MS) [153] or sector field mass spectrometer (ICP-SF-MS) also known as high resolution ICP-MS [143, 155, 322]. Compared to the routinely used quadrupole-ICP-MS, a major advantage of the sector field mass spectrometer (ICP-SF-MS) is the possibility to use higher

72 resolutions, which allows increased separation of interferences. In addition, ICP-SF-MS is more sensitive and precise, especially at low resolution mode [323].

The basic components of an ICP-MS are shown in Figure 2.1. and consist of a sample introduction system, the plasma torch, the interface, the ion optics, the mass analyser and the detector. The sample introduction system consists of a peristatic pump, a nebulizer and spray chamber, transforming the sample into an aerosol and transferring the small droplets to the plasma torch. The plasma torch consists of three concentric tubes. The central tube is the injector tube, transporting the sample to the plasma, the central tube is the auxiliary gas, which can change the position of the plasma and the outer tube an argon carrier gas, which transports the generated ions into the interface and serves as cool gas for the quartz torch. A radio-frequent (RF) coil, positioned at the end of the torch, is connected to an RF generator. The plasma is induced with a high voltage spark, generating electrons which are accelerated and collide with other argon atoms, creating more electrons and ions. This chain reaction results in an inductively coupled plasma (ICP), consisting of argon ions and electrons. The ICP is sustained as long as the RF field remains, and when sample droplets are introduced into this hot plasma, the sample is desolvated, atomized and ionized. The generated ions are then transported to the interface. The interface consists of a sampler and a skimmer, placed behind each other, which a pressure drop between the cones. The role of the interface is to transport ions from the plasma to the mass spectrometer efficiently and consistently. After passing the skimmer, ions are transported to the ion optics. The ion optics are a sequence of electrostatic lens components which direct ions from the interface to the mass analyser and separate ions from non-ionic species neutral species and photons.

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Figure 2. 1. General scheme of an ICP-MS instrument [324]

The mass separation device is either a quadrupole MS (in ICP-QMS) or a sector field mass MS were the ion beam passes through a magnetic and electric sector. Only the latter MS will be described here.

Double-focusing electric/magnetic sector instruments, are based on measurements of ion deflection [325-327]. The term double-focusing refers to the fact that the combination of electrostatic and magnetic sectors focuses ions according to both direction and energy to provide higher resolution. The magnetic sector (B) exerts a force perpendicular to the ion motion to deflect ions according to their momentum. Higher mass ions are deflected less than lower mass ions, so ions are separated only by their masses[325-327]. To obtain a spectrum of good resolution, the electrostatic sector (E) creates an electric field that exerts a force perpendicular to the ion motion to deflect ions according to their kinetic energy. Two geometries exist from these mass analyzers, depending on the order of the different sectors: (reversed Nier-Johnson Geometry) EB or BE (Nier-Johnson Geometry). The former is used in the Thermo Fisher Element II (Figure 2.2). The combined result of both fields is that ions are only dispersed according to mass, regardless of ion energy and angle of entry. There are a number of different ways to acquire accurate mass measurements: peak matching, dynamic voltage scanning, and magnet scanning. Briefly, the different scan modes differ whether the magnetic field is held constant (peak matching and dynamic voltage scanning) or the magnetic sector is scanned over a wide m/z range (magnet scanning).

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Figure 2. 2. Schematic setup of a magnetic sector field ICP-MS (in this example of the Element 2).

The ability to separate species is described by the resolving power or resolution (is the ability of an instrument or measurement procedure to distinguish between two peaks at m/z values differing by a small amount and expressed as the peak width in mass units [325]), which is defined Rp = M/dM; where Rp = resolution; M = nominal mass peak; dM = mass difference between the peaks. For magnetic sector instruments, two peaks of equal intensity are considered to be resolved when they are separated by a valley, which is 10% of the height of each peak[325]. Mass resolution is achieved by using two mechanical slits—one at the entrance to the mass spectrometer and another at the exit, prior to the detector. Varying resolution is achieved by scanning the magnetic field under different entrance and exit slit width conditions. Low resolution is achieved by using wide slits, whereas high resolution is achieved with narrow slits. [324]. Commercial instruments for SF-ICPMS can be operated at resolution settings which range from low (m/∆m = 300) to high (m/∆m = 10,000).

The detector is a discrete dynode electron multiplier. In this detector, an incoming ion strikes a conversion dynode which produces electrons, which are accelerated to the next dynode where they generate more electrons. This process is repeated with subsequent dynodes, generating a pulse of electrons that are measured at the collector anode. In order to have a broad dynamic range, current detectors count these pulses at low amounts of ions, and switch to less sensitive analogue signals measured at an intermediate dynode instead of the collector when concentrations are high.

2.2.2. Specifications of the Thermo Fisher Element II ICP-SF-MS In this study PGE analysis was performed using Inductively Coupled Plasma Sector Field Mass Spectrometry (ICP-SF-MS, Element II, Thermo Fisher Scientific Bremen GmbH, Germany). Instrumental settings are shown in Table 2.1. Sample introduction was performed with an ESI- SC-Fast introduction system. The system uses a high flow vacuum pump, 6-port valve and

75 sample loop to deliver the sample to the nebulizer and this minimizes sample uptake and rinse- out time and avoids contact of the sample with the peristaltic pump tubing. Internal standard is mixed with sample on-line before the mixture passes through the concentric PFA-ST microflow nebulizer. A Peltier cooled cyclonic spraychamber separates the fine aerosol. The plasma torch is equipped with a guard electrode to minimize capacitive coupling. The resolution settings are R=300 for low resolution (LR), R=4000 for medium resolution (MR) and R=10000 for high resolution (HR).

Table 2. 1. Specifications of the ICP-SF-MS instrument used in this thesis Instrument ELEMENT2 Thermo Finnigan Forward power 1,350 W Reflected power < 2 W Nebuliser Concentric Solution uptake rate 0.4 mL min-1 (pumped) Spray chamber Cyclonic Sampling and skimmer cones Ni (Thermo Finnigan) -1 Sample gas flow 1 min Cool argon flow rate 16 L min-1 Auxiliary argon flow rate 1.0 L min-1 Torch Capacitive decoupling Pt shield torch RF frequency 27.12 Mhz Sensitivity 1 x 106 cps per 1 ng mL-1 115In (in LR) 1 x 105 cps per 1 ng mL-1 115In (in MR) 1 x 104 cps per 1 ng mL-1 115In (in HR) Take-up time 15 s Wash time 10 s Number of acquisition 6 (3 runs and 2 pass) Mass window LR 20% Search window LR 0% for each isotope Integration window LR 20% for each isotope Mass window MR/HR 125% for each isotope Search window MR/HR 60% for each isotope Integration window MR/HR 60% for each isotope Scan type E scan for each isotope Integration type Average for each isotope

2.2.3. Stability of PGEs in solutions and instrumental memory effects A potentially significant problem associated with field work and laboratory experiments using PGEs and their analysis are the strong tendency of hydrolysis and precipitation [144, 328], the adsorption onto the surface of vessels under weakly acidic and neutral conditions [108, 328, 329], and transformations during storage and pre-treatment steps [108, 112, 113, 205]. The PGEs concentration, the sample matrix (pH, dissolved salts and organic matter, …….etc.) and

76 the vessels are important factors play simultaneously a role in the stability of the PGEs in the solution [328]. Choice of container material

For all three elements stability of PGEs in solutions follows the order: quartz> borosilicate glass (BG)> fluorinated ethylene propylene (FEP)> polytetrafluorethylene PTFE (Teflon®)> polypropylene (PP)>polyethylene (PE) [328, 329]. Stability of PGEs increase in solution in the presence of organic matter due to the formation of stable complexes [330] and in saline solutions due to the formation of stable chloride complexes [328]. Some loss has been reported for Pd(II) and Pt(IV) in FEP containers, also in solutions containing organic matter suggesting some hydrophobic interactions [328]. In acidic conditions (pH<3) solutions of Pd and Rh in MQ water are relatively stable in PE, PP, FEP and BG but loss of Pt is only stable in BG At higher pH hydrolysis products result in loss from solution which is most pronounced for Rh [331].

For the storage of natural water samples, we performed a study on the stability of spiked filtered river water acidified with 1% HNO3 in HDPE, PP and BG containers. All the materials were first washed with detergent and MQ water then soaked for 24h in %10 HCl and then with

10% HNO3 for another 24h. After that, the bottles were washed vigorously with MQ water and filled with 1 % HNO3 and this considered as the blank. Analysis of the blanks showed no difference between the containers. The containers were then filled with filtered river water -1 (Zenne River) and spiked to a concentration of 25 µg L and acidified to 1% HNO3. The solutions were sampled at the time of preparation, and daily over a period of 6 days. The results showed that the concentrations of Pt and Pd remained constant in all three containers but a 10% loss of Pd was observed after 1 day in PP and HDPE and a 4% loss in BG. The concentrations remained stable after day 1. BG bottles were thus selected for the collection and storage of field samples.

Choice of acid for storage of solutions and analysis Further investigation showed that the stability of the solutions increased using 1% HCl, which is generally used for the storage of PGE solutions. However, comparison of the analysis of samples acidified with 2% HCl and with 3% HNO3 showed that instrumental memory effects are more pronounced using 2% HCl compared to 3% HNO3 (Figures 7 and 8). Subsequently, for the field work, the samples were acidified with 1% HCl and 2% HNO3.

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Pt Pd 150 150 Rh µg/L µg/L 120 µg/L 100 100 100 80 60 50 50 40 Number of the Number of the 20 Number of the measurment measurment measurment 0 0 0 0 20 40 60 0 20 40 60 0 20 40 60 Figure 2. 3. Shows the repeated measurement (n=45) of 80 µg L-1 Pt, Pd and Rh using LR mode in 3% HCl matrix using 3% HCl carrier solution. Dilution factor is 10.

Figure 2. 4. Shows the repeated measurement (n=13) of 10 µg L-1 Pt and Rh, 6 µg L-1 using LR mode in 3% HNO3 matrix using 3% HNO3 carrier solution. Dilution factor is 10.

Choice of container material for DGT deployments

Preparation of PGE spiked solutions is best performed in glass containers which have been equilibrated with PGE spiked solutions for at least 3 days to avoid adsorption on the container walls. This solution was then discarded, and a freshly prepared test solution was placed in the containers. This procedure used for all tests to avoid loss of the PGEs due to adsorption on the container surface, especially in PE or PP containers. Glass containers showed a much higher stability. The pre-treated glass containers showed an insignificant loss of PGE’s, even at low salinities. Deployment solutions consisted of 0.01M, 0.1M, 0.5M NaCl matrix, at a concentration level of 10 µg L-1 and 1000 µg L-1 and in spiked filtered seawater (North Sea, Belgian coast), spiked filtered river water (Zenne River, Belgium) and artificial seawater. Either 20L glass aquaria or 2L PP boxes were used as deployment containers. Spiked natural waters (river water and seawater) have a higher stability than artificial low salinity waters.

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2.2.4. Interferences in PGE analysis Interferences in atomic spectrometry are generally are either matrix induced interferences (physical interferences, chemical interferences, ionization interferences) or the spectral interferences [324, 332]. 2.2.4.1. Matrix induced interferences There are basically three types of matrix induced interferences (physical, chemical, and ionization). The first is often called a sample transport effect and is a physical suppression of the analyte signal, brought on by the level of dissolved substances or acid concentration in the sample. It is caused by the sample’s impact on droplet formation in the nebulizer or droplet size selection in the spray chamber. In the case of organic matrices, it is usually caused by variations in the pumping rate of solvents with different viscosities. The second type of matrix suppression is caused when the sample affects the ionization conditions of the plasma discharge. This results in the signal being suppressed by varying amounts, depending on the concentration of the matrix components. Ionization or space-charge interferences or called mass discrimination, which is the magnitude of signal suppression in ICP-MS increased with decreasing atomic mass of the analyte ion [324]. The classic way to compensate for a physical interference[333] or correct possible instrumental drifts [334] or the loss in analyte sensitivity [335] and improve reproducibility of measurements is to use internal standardization (IS) [333]. An isotope of a different element - that is not present in the sample at any significant level- is added to all standards and samples and the mass spectrometer ion counts of analytes are normalized to those of the IS [336]. If the transport efficiency remains the same for all the samples, the signal due to the internal standard should remain constant provided that the internal standard is chosen so that the analytes and the internal standard behave in a similar manner to the changes in any other parameters[337] - no mass discrimination in transport and measurement[336]-. Then, any changes in transport efficiency should be reflected in the internal standard signal and a correction factor can be calculated accordingly [337]. Generally, useful internal standards are: 1-The isotope of the IS must provide excellent signal stability over extended runs >12 h [338] and the internal standard element response is the same as that of the analyte in the presence of concomitant elements [335]. 2-ISs have a similar ICP- indexes to that of the analyte[332]. 3- The element chosen as internal standard has to be added in relatively high concentration so its measurement is as accurate as possible[334] [335].

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Throughout this work, a mixture of 5 internal standards were evaluated: lanthanum, 139La, iridium 191Ir, indium 115In, and bismuth 209Bi. Based on the ICP-index [332] (Table 2.2) and measurement of the reproducibility of the normalized intensities of the analytes using the different internal standards in different matrices, Rh, Pd, and Pt intensities were normalized to (115In, 209Bi and 193Ir). The use of multiple internal standards permits recalculation with another internal standard in case of aberrant results. Table 2. 2. Isotopes and first ionization potentials of the analytes and internal standards (ISs) Element isotope First ionization potential (eV) [339] ICP-index [332] Rh 103 7.45 7.72 Pd 105 8.34 9.08 Pt 195 8.95 13.07 In 115 5.78 6.73 Ir 193 8.96 12.76 La 139 5.58 4.46 Bi 209 7.29 9.57 Re 185 7.88 6.83

2.2.4.2. Spectral interferences One of the main problems in determining PGEs by ICPMS is the existence of spectral interferences. In the case of PGE, spectral interferences occur due to a) isobaric overlap, b) doubly charged ions, and c) polyatomic ion interferences. Spectral interferences on the different isotopes of the PGEs and required resolution to separate the interferences from the analytes are listed in Tabel 2.3.

2.2.4.2.1. Isobaric Interferences Isobaric interferences occur when two or more elements have isotopes at the same nominal mass. The Pd isotopes at m/z 106 and 108 suffer from isobaric Cd interferences Table 2.3. (natural abundances for 106Cd and 108Cd are 1.25% and 0.89%, respectively), which require a resolution higher than 11 000, i.e. exceeding the maximum resolution power of HR-ICP–MS equipment by current commercial instruments; therefore, no separation from the analyte peak can be achieved. However, in practice, isobaric interferences are not always a problem because all elements, with the exception of In at m / z 113 and 115 (which overlap with 113Cd and 115Sn), have at least one isotope (as for 105Pd) which is free from isobaric overlap. Isobaric interferences can also be calculated based on the natural abundance ratios of the isotopes of the interfering element and the measurement of a non-inferred isotope of the interfering element.

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2.2.4.2.2. Doubly Charged Species Doubly charged ions (M+2 are common for elements with a low second ionization potential, such as barium, which has a second ionization potential of 10.0 eV. The alkaline earth elements, the rare earth elements and elements such as U and Th are most likely to form M+2 species (as 206Pb+2 on 103Rh; table 2.3). The formation of a doubly charged ion results in a loss of sensitivity for the singly charged species and generates an isotopic overlap at one half of the mass of the parent element.

2.2.4.2.3. Polyatomic Ion Interferences This form of spectroscopic interferences is far more of a problem than isobaric interferences and can be much more difficult to overcome. Polyatomic ions are molecular species which are formed in the plasma and interface region of the ICP - MS. They are caused by a variety of factors but are usually associated with either the plasma/nebulizer gas used, matrix components in the solvent/sample, other elements in the sample, or entrained oxygen/nitrogen from the surrounding air[340].These polyatomic ions typically come from precursors in the argon support gas (e.g. Ar, H, O), entrained atmospheric gases (e.g. N, O), from solvents or acids used during sample preparation (e.g. N, CI, S, P), or from the sample matrix, and are particularly problematic. These interferences are typically produced in the cooler zones of the plasma, immediately before the interface region[340]. Polyatomic ion interferences can be placed into several groupings; oxides i.e. ArO+, ClO+, MO+; hydroxides i.e. ArOH+, ClOH+, MOH+; hydrides i.e. ArH+, MH+; argides i.e. MAr+ [108, 172, 338, 341-343]. The PGEs interferences in ICPMS are studied in [88, 108, 172, 178, 338, 341-344] and summarized in table 2.3. Table 2. 3. Potential spectral interferences on the masses of Pd, Pt and Rh[178, 343] b b Analyte Interferent Minimum Analyte Interferent Minimum resolution resolution isotope abundance species a abundance m/∆m isotope abundance species a abundance m/∆m 105Pd 22.33 40Ar65Cu 0.003 7300 194Pt 32.90 178Hf16 O 27.2 8100 36Ar69Ga 0.20 92000 177Hf17 O 0.007 9600 89 16 176 18 Y O 99.76 27600 Hf O 0.01 8800 88Sr17O 0.03 1000000 176Yb18 O 0.02 9200 87Sr18 O 0.014 30900 176Lu18O 0.005 9200 87Rb18 O 0.056 28400 195Pt 33.80 179Hf16O 13.6 8200 106Pd 27.33 40Ar66Zn 27.79 7200 178Hf17O 0.01 6900

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38Ar68Zn 0.013 6800 177Hf18 O 0.04 8800 36Ar70Ge 0.002 9300 103Rh 100 38Ar65Cu 0.02 7200 90Zr16 O 51.3 26500 40Ar63Cu 68.89 8040 88Sr18 O 0.16 70600 36Ar67Zn 0.01 10100 89Y17 O 0.037 58800 87Sr16 O 6.99 102900 106Cd 1.25 27900 87Rb16 O 27.76 147000 108Pd 26.46 38Ar70Ge 0.014 8600 85Rb18 O 0.14 17200 36Ar72Ge 0.005 7300 206Pb2+ 24.14 1248 40Ar68Zn 18.7 6500

92Zr16 O 17.1 540000

91Zr17 O 0.004 216000

92Mo16O 14.8 40000

108 Cd 0.89 1080000 aMonoatomic interferents: natural abundance; polyatomic species: maximum abundance possible calculated as the product of the natural abundances of the two isotopes forming the molecular ion divided by 100 bMinimum resolution required to separate the interferent and analyte masses.

.2.4.2. Interference study We performed a detailed investigation on the interferences on the ICPMS measurements of PGEs using model solutions as described in [143, 345, 346]. Experiments were performed 10x diluted aqua regia as this will be used as eluent in the DGT measurements. Using single element solutions of the interfering compounds at appropriate concentration ranges; the intensity of the interfering species formed (e.g.179Hf16O+ on 195Pt) is plotted in function of the intensity of the interfering isotope (e.g.179Hf) and from the linear relationship interference correction equations can be established. Analysis is performed in LR, MR and HR resolution modes. For103Rh, the most important interferences of concern are from 40Ar63Cu+, 87Sr16O+, 87 16 + 36 67 + 206 2+ 194 178 16 195 179 16 + 105 Rb O , Ar Zn and Pb ; for Pt this is Hf O, for Pt this is Hf O , for Pd 40 65 + 89 16 + 88 16 + 87 18 + 106 interferences of concern are from Ar Cu , Y O , Sr OH and Rb O , for Pd these are 106Cd , 40Ar66Zn , 90Zr16 O , 88Sr18 O, 89Y17 O and for 108Pd these are 108Cd, 40Ar68Zn , 92Zr16 O, 92Mo16O (Table 2.3). In Figure 2.5 the interferences of Hf on 194Pt and 195Pt, of Cu, Zn, Pb, Sr and Rb on 103Rh , of Cu, Y, Sr and Rb on 105Pd, of Y, Zr and Cd on 106Pd and Mo, Zr, Cd and Zn on 108Pd are plotted for measurements made in LR, MR and HR modes. 10x diluted aqua regia is used as sample matrix.

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In general, the problem of spectroscopic interferences can be dealt with in two ways, either by their removal or by correcting for them mathematically [347, 348]. Instrumental approaches to remove interferences include nebulizer with cryogenic desolvation [88, 207], matrix separation of interfering species: offline/online chromatographic systems, cold plasma conditions, collision/reaction cell technologies[153], high mass resolution spectrometry [88, 143, 155, 322, 349]. The best and probably most efficient way to remove spectral overlaps is to resolve them using a high-resolution mass spectrometer [348, 350]. The possibility of separating interfering signals from analyte signals depends upon the required resolution m/∆m, [351]. A resolution of 10,000 may be sufficient to overcome some interferences of polyatomic and double-charged ions on analyte ions. For example, the resolution required to overcome interference of HfO+ on 195Pt is 8200, of 206Pb2+ on Rh is 1248 and ArCu+ on 103Rh is 8040; ( table 2.3), thus these interferences can be resolved using HR and MR modes on the Element II. Although their resolving capability is far more powerful than quadrupole-based instruments, there is a sacrifice in sensitivity if HR is used (to 1-2% of that found with LR mode) and 10% for MR compared to LR [324, 349], which can often translate into a degradation in detection capability for some elements. The analysis of Pt is interfered by Hf (Figure 2.5) and the apparent 195Pt concentrations accounts for 1% of the Hf concentrations in LR and MR and this interference can completely be removed in HR mode. 103Rh measurements by ICPMS can mainly be interfered by 206Pb2+, 40Ar63Cu+, 36Ar67Zn+, 87Sr16O+, 87Rb16O+. The apparent 103Rh concentration accounts for 0.006% of the Cu concentration in LR and MR and can be completely resolved in HR. The interference of Sr accounts for 0.02% of the Sr concentration in all resolutions. For Pb, Zn and Rb, no increase in the Rh signal was observed with increasing concentrations of interfering elements. Sr is thus the most important interferent in the Rh analysis in natural waters due to the high concentrations of Sr present and high interference factor. The required resolution (102900) makes it impossible to separate the interference in HR. 105Pd measurements by ICPMS are interfered mainly by 40Ar65Cu+, 87Sr16OH+, and 89Y16O+. Figure 2.5 shows the apparent 105Pd concentrations in function of the concentrations of interfering elements. The apparent 105Pd concentration accounts for 0.02% of the Cu concentration in LR and MR and can be completely resolved in HR. The interference of Y accounts for 1% of the Y concentration and interference of Sr for 0.1% of the Sr concentrations and both interferences cannot be resolved by increasing the resolution settings. Also, for 105Pd, Sr will be the most important interference in natural waters. The apparent 106Pd concentrations in function of the concentrations of interfering elements Y, Zr, Zn, Sr and Cd is shown in Figure 2.5. The isobaric

83 interference of 106Cd on 106Pd is 0.0003% of the 111Cd concentration. In natural waters, the 111Cd will also include possible interference by 95Mo16O on the111Cd signal. The interference of Y accounts for 0,2% of the Y concentration and interference of Zr for 0.03% of the Zr concentrations. The interference of Mo, Zr, Cd and Zn on 108Pd (Figure 2.5) shows that the interference accounts for 0.0002% of the Mo concentrations, 0.01% of the Zr concentrations 0.0004% of the Zn concentrations and 0.0025% of the Cd concentrations. Only the Zn interference can be removed in HR. Although 108Pd has an isobaric interference, the absence of the Sr interference increases the reliability of this isotope.

35 195 Pt_ 60 194Pt_ apparent 30 y = 0.0053x - 0,1295 apparent concentrat R² = 0.9992 50 ion (ng/L) concentra 25 tion (ng/L) 40 20

15 30 y = 0.0109x - 0.2960 10 20 R² = 0.9991

5 10 Hf (ng/L) 0 Hf (ng/L) 0 1000 2000 3000 4000 5000 6000 0 0 1000 2000 3000 4000 5000 6000 LR MR HR LR MR HR

9 Rh_ apparent 2,5 8 concentration Rh_ apparent ng/L concentratio 7 n ng/L 2,0 6

5 1,5 4

3 1,0 2

0,5 1 Cu (µg/L) 0 Rb (µg/L) 0 20 40 60 80 100 120 140 160 0,0 0 10 20 30 40 50 60 LR MR HR LR MR HR

4,5 Rh_ apparent 7 concentration Rh_ apparent 4,0 (ng/L) concentration 3,5 6 (ng/L)

3,0 5

2,5 4 2,0 3 1,5 2 1,0

0,5 1 Pb (µg/L) Zn (µg/L) 0,0 0 10 20 30 40 50 60 0 0 100 200 300 400 500 600 700 800 900 LR MR HR LR MR HR

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30 105Pd_ apparent 25 concentration 100 105Pd_ (ng/L) 90 apparent 20 80 concentration (ng/L) 70 15 60 50 10 40 30 5 20 Cu (ng/mL) 10 0 Y (ng/mL) 0 50 100 150 200 0 0 0,2 0,4 0,6 0,8 1 1,2 LR MR HR LR MR HR

20 105Pd_ 18 apparent concentration 1400 105 16 (ng/L) Pd_ 1200 apparent 14 concentratio 12 1000 n (ng/L) 10 800 8 600 6 400 4

2 200 Rb (µg/L) Sr (ng/mL) 0 0 0 10 20 30 40 50 60 0 200 400 600 800 1000 1200 LR MR HR LR MR HR

200 20 Rh_ apparent 106Pd_ 180 concentration 18 apparent (ng/L) concentration 160 16 (ng/L) 140 14 120 12 100 10 8 80 6 60 4 40 2 Zn (µg/L) 20 Sr (µg/L) 0 0 0 2 4 6 8 10 0 200 400 600 800 1000 1200 LR MR HR LR MR HR

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40 106Pd_ 35 apparent cocnetration 30 (ng/L) 25 20 15 10 5 Sr (µg/L) 0 0 200 400 600 800 1000 1200 LR MR HR

Figure 2. 5. Interference equivalent concentrations (IEC) in function of the concentration of interfering elements on 105Pd, 106Pd, 108Pd, 194Pt, 195Pt and 103Rh measured in low, medium and high resolutions by ICP-SF-MS.

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2.2.4.3. Mathematical correction Using single element solutions of the interfering compounds at appropriate concentration ranges in the absence of analyte; the intensity of the interfering species formed (e.g.178Hf16O+ on 194Pt) is plotted in function of the intensity of the interfering isotope (e.g.178Hf) and from the linear relationship interference correction equations can be established[345, 352-354]. For the interfering element, ideally an isotope with no interferences is chosen for the corrections. Alternatively, standard addition of interfering elements can be performed to take matrix effects into account. In this work the inteferences were evaluated in a 3% HNO3 matrix and a 10x diluted aqua regia matrix. The advantage of the standard interferent addition method over the aqueous calibration ratio method is that it takes into account the possible variation in the ratio of interferent formation in the real sample relative to the aqueous standards due to the matrix effect. It is known that the MO+/M+ ratio varies from sample to sample depending on the matrix composition, operating condition and gradual closing of the sampler orifice [155].

Interference corrections aquations Sc =Sm − (Sinter A) where Sc is the corrected signal of the analyte,

Sm the measured signal, Sinter the signal of the interfering element and A is the % of formation of the respective interfering species. Typical interference corrections obtained in this study in LR were: 194 194 178 S Ptc = S Ptm – 0.011*S Hf 195 195 178 S Ptc = S Ptm – 0.005*S Hf 105 105 88 65 89 S Pdc = S Pdm – 0.002*S Sr – 0.0000635* Cu – 0.012* Y 106 106 88 66 89 111 S Pdc = S Pdm – 0.0002*S Sr – 0.00014* Zn – 0.0005* Y – 0.0913* Cd 108 108 90 98 111 S Pdc = S Pdm -0.00724* Zr – 0.0013* Mo – 0.0657* Cd 103 103 88 65 S Rhc = S Rhm – 0.009*S Sr – 0.000128* Cu As the Hf interference on 194Pt is larger than that on 195Pt, 195Pt is the preferred isotope. Sr is the most important interference for Rh and Pd due to the high concentrations in natural waters and often the limiting factor in the reliability of the correction equations. When the contribution of the interference accounts for more than half of the observed signal, accurate measurement is not possible. Mathematical corrections have some difficulties; 1-Measurement of more isotopes. 2- Problematic if interfering ion has higher signal intensity than the isotope of interest (The relative contribution of interferences to the PGE signal is calculated as the ratio of the total concentration of interferences and the corrected PGE concentration. At an interference ratio of 1, half the signal is due to interferences, and the other half originates from the PGEs in the

87 sample. At an interference ratio lower than 1, PGE concentrations can be accurately determined with mathematical concentration, while interference ratios higher than 1 need careful estimation of interference and calculation [346, 354]). 3-More sample is required to produce standard addition. 4- One standard has to be prepared for each sample. 5- Time and material consuming. 6- Interfering matrix elements have to be mixed together. Formation probability of polyatomic ion can only be estimated and is not always reproducible[355]. 7- This mathematical correction method requires a linear dependence of the IO+ signal on the interferent I+ concentration[353]. 8- mathematical correction of polyatomic interferences generally requires more elaborate schemes and are not always accurate due to the large differences in concentration levels between the PGE and the interfering elements but also because the formation of the interfering species is also dependent on the operating conditions[342]. 9- The correction factors of interfering species needed in the correction equations were measured daily using single element standard solutions [152].10- The disadvantages are that it does not take mass-dependent sensitivity effects (mass bias) into account and that the accuracy depends on the type and magnitude of interference relative to the analyte. Also, there are some cases where theoretical corrections simply cannot be used because there is no alternative isotope of the interfering species to monitor. 11- It is well known that the contribution of interference to analyte signals is highly dependent on the instrument and the sample introduction method used. This can even change dramatically in the course of a single day working under strict repeatability conditions [141]. Therefore, chemical separation techniques are often used to isolate and pre-concentrate the target elements from the ‘parent’ elements of the interfering ions. In this context we have fully discussed this issue in our recent publication [356].

2.2.5. Practical considerations in ICPMS measurements The accuracy and the precision using MS achieved depending on several factors include tuning and peak shape, ion abundance, resolution, calibration, sample introduction, data manipulation, validation, and quality control checks. Here I will discuss only the main factors that may affect the PGEs measurements in ICPMS.

2.2.5.1. Matrix effects, calibration procedures, blanks, and detection limits The sample matrix can bias the analyte signal through spectral and non-spectral interference effects [351]. Matrix effects can be reduced by simply diluting the sample (if permitted by analyte concentration) or corrected for by using internal standardization or the use of isotope

88 dilution mass spectrometry [351]. Internal standard correction will not work efficiently when the matrix effects are too severe requiring standard addition calibration. As shown in section

2.2.3. HNO3 is a better choice of acid than HCl for the analysis regarding memory effects and wash-out time. Thus, final dilution for external calibration standards were prepared in 3%

HNO3; the internal standard mixture was also prepared in 3% HNO3 and the carrier solution also 3% HNO3. In this work aqua regia was mainly used as eluent and to obtain sufficient sensitivity, a minimal 10-fold dilution was applied. Standard addition calibration curves were made in 10-fold diluted aqua regia, as well as in all matrixes used (thiourea, diluted seawater, river water, etc.). Further considerations of making the calibration such as the calibration range, the number of the point, calculating the uncertainties, calculating the concentrations, data handling ….etc are adopted from these studies using [347, 357-368]. The Pt, Pd and Rh calibration curves ranging 0.05 to 15 µg L-1 for the metal ions yielded good correlation coefficients (R2) ranging from 0.99 to 0.999. Regarding the minimal dilution, it is important to avoid clogging of the nebulizer a maximum 0.2% (m/v) total dissolved substances is recommended in ICPMS analysis.

Changing matrix during analysis (example from 3% HNO3 to 10-fold diluted aqua regia or diluted thiourea) can result in the release of PGEs adsorbed in the sample introduction system. Therefore, several blanks of the matrix should be run before the samples until a stable background signal is obtained. An typical average reagent blank (10% aqua regia) (n= 10 replicate) of 1.10 ng Pt L-1,0.34 ng Pd L-1, and 0.355 ng Rh L-1 was obtained, resulting in a detection limit of 1.16 ng Pt L-1, 0.83 ng Pd L-1, and 0.47 ng Rh L-1 (determined as the average blank +3 times standard deviation).

2.2.5.2. Resolution The isotopes (103Rh, 105Pd, 106Pd, 108Pd, 194Pt and 195Pt) were measured in three resolutions (Low (LR), Medium (MR) and High Resolution (HR) with resolutions of 300, 4000 and 10000 respectively). The instrumental sensitivity is highest in LR mode (106 cps/ppb 115In compared to MR mode (105 cps/ppb 115In) and HR mode (104 cps/ppb 115In). Although HR can solve many of the interferences (Table 2.3), this also results in lower sensitivity and lower precision. For the laboratory experiments which were performed at µg L-1 levels, no difference was observed between the concentrations found in LR, MR or HR and the results in LR were used. For real samples and evaluation of blanks, MR and HR was also used and potential interferences evaluated.

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2.2.5.3. Contamination Common contamination sources include: the reagent water, typically with a conductivity of 18.2 MΩ cm; the cleaning solution used for the glassware; storage of the glassware pre‐use; the risk from airborne dust in the laboratory; the use of powdered gloves (especially for Zn) for sample handling and contamination risk from (coloured) pipette tips (especially Cd, Cu, Fe and Zn) [369]. Cd, Cu, Zn are the source of increasing the blanks for PGEs [356]. The use of Rh as internal standard or high concentrated standards such (Sr, Rb, Cu, Y …. etc.) or samples containing such high concentration of these elements cause the contamination effect in the instrument and those elements in return they form oxides and argid species that cause increase in the apparent concentration (interferents species) of PGEs and hence increase in the blanks. Rigorous cleaning procedures of the sample introduction system are thus required.

2.3. DGT preparation and procedures

2.3.1. Diffusive gel preparation The polyacrylamide gel (PAM) was prepared as described by Zhang and Davison [281]. Agarose diffusive gel (AGA) was prepared by dissolving 1.5g of agarose (Biorad) in a 100mL of Milli-Q water. The mixture was placed in water bath (80◦C) and gently stirred until all the agarose was dissolved and the solution became transparent. The hot gel solution was pipetted between two preheated glass plates separated by a Teflon spacer of 0.75 mm thickness and left to cool down to its gelling temperature. The set gels were cut into discs of 2.5 cm diameter and placed in deionized water for 24 hours and the water replaced several times. Then the gel discs were stored in electrolyte of 0.01M NaNO3 solution at room prior to assembly.

2.3.2. Binding gels preparation investigation Following the thorough literature study presented in Chapter 1, Ten resins were selected with a high affinity and selectivity for PGEs. The properties of these resins are shown in Table 2.4. All the selected resins are ion-exchange and/or chelating resins, which imply that binding can occur through ion exchange, electrostatic interactions and complex formation with the lone pair of the ligands of the functional groups. As such, the binding can occur regardless of the charge of the metal complex. The high stability of the polydentate PGE complexes on the resins explains the selectivity of these resins towards PGEs. Only S985 is a weak base chelating ion- exchange resin, which implies that negatively charged metal species can also be bound by ion- exchange in acidic or neutral conditions.

Table 2. 4. Physical-chemical characteristics of ion exchangers investigated

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Bead Operating size Exchanger Trade name Functional group Matrix References pH range range type (µm) Thiourea Macro porous Purolite polystyrene 300 - 0-14 Chelating [221] S914 crosslinked with 1200 divinylbenzene

Isothiouronium Macro porous + Purolite NH2 polystyrene 300 - 0-13 Chelating [186, 220] S920 crosslinked with 1200

R S NH2 divinylbenzene Polyamine Chelating Purolite Polyacrylic- 300 - Weak [112-114, 0-10 205, 240, S985 divinylbenzene 1200 base anion 241, 289] exchanger

n=1 or 2 Purolite Thiol 550- polystyrene 0-13 Chelating S924 R-SH 800 IONQUEST® chelate phosphine 300- - 0–10 Chelating [201] MPX-317 thiourea 1200 phosphine oxide IONQUEST® - 0.1–10 Chelating [201, 232] MPX-310

Amino phosphine IONQUEST® oxide Powder 0–10 0.1–10 Chelating [189] MP-102 polymer

IONQUEST® phosphine Powder - Chelating [201] MP-317 Thiourea polymer Alkylphosphines

IONQUEST® Powder - Chelating MP-101 polymer

strong Dowex sulfonic acid [154, 63-150 cation 50W-X8 R-SO -2 3 exchange 370-373]

The binding layer, or resin gel, consisted of agarose gel (Biorad) in which the resin was dispersed. Grinding and sieving of the tested resins was necessary for the DGT applications, with the exception of MP-102, which was available in powder form. In the first step, the resins were grinded in a mechanical mortar (Fritsch Pulverisette) and sieved using a 50µm

91 nylon sieve (except Purolite 985 which was sieved with 250µm because of the difficulties of grinding this resin). More homogeneous resin gel discs could be made using resins sieved on 50µm compared to 250µm, which resulted in a better repeatability. Grinding results in water loss and may affect the resin performance, thus the resins have to be freshly grinded and sieved before use. Storage of the prepared resin gels should be limited to several weeks to avoid deterioration of the performance.

Freshly grinded and sieved resins were mixed with a hot 1.5% agarose solution in optimized ratios: for Purolite S914 (1 gr/8 mL), Purolite S920 (0.5gr/7 mL), Purolite S985 (0.8 gr/8 mL), Purolite S924 (0.7 gr/10ml), IONQUEST® MPX-317 (0.5 gr/10 mL), IONQUEST® MP-317 (0.5 gr/7 mL), IONQUEST® MP-101 (0.5 gr/7 mL) and IONQUEST® MP-102 (0.5 gr/ 7mL) in polypropylene vessels. Attempts to increase the amount of resin resulted in very fragile or sticky resin gels that are impossible to remove from the plates. The resin mixture is well mixed and cast between two pre-heated acid-washed glass plates separated by a Teflon spacer (0.75 mm) and allowed to set at room temperature. The resin gel is then cut on the glass plate using a 2.5 cm diameter plastic cutter and stored in deionized water at 4 °C prior to its assembly of the DGT device.

2.3.3. DGT assembly DGT samplers were supplied by DGT Research and assembled according to the protocol from Lancaster (www.dgtresearch.com). Concisely, the resin gel disc was placed on the base plate, with the resin side facing up, then a diffusive gel was placed on it followed by a Millipore Durapore membrane filter (HVLP pore size 0.45 µm, diameter 2.5 cm, thickness 120µm). The filters were rinsed in 0.7M HNO3 for about 24 h and then rinsed 5 times in MQ water before being stored in MQ water prior to use. Assembled DGT samplers were stored at 4°C in doubled zippered plastic bags, which contained 0.01 M NaNO3, to maintain moisture.

2.3.4. Diffusion cell experiments

The procedure of Zhang and Davison for the determination of diffusion coefficients (Dcell) in agarose gels using a diffusion cell is described in detail elsewhere[270]. The diffusion cell consists of two 250 mL Plexiglas compartments, which are interconnected by a 1.5 cm diameter opening window. A 2.5 cm diameter agarose gel disc (∆g = 0.75 mm) and cellulose nitrate filter membrane (∆g = 0.17 mm) were placed between the opening windows of the two compartments. Experiments were performed in various solutions (0.01 M NaCl, filtrated river water, filtrated seawater and artificial seawater. 230 mL of carrier solution was introduced into

92 one compartment known as the receptor or receiver solution and 230 mL of carrier solution spiked with PGEs into the other (known as the source solution). Test solutions were stored in borosilicate bottles after spiking and let at least for 24h for equilibration after spiking. Both compartments were stirred continuously using an overhead stirrer. Sub-samples of 1 mL were taken from each compartment - acidified with 0.3 mL of 14M HNO3 - at various intervals and diluted 10 times before the analysis by ICP-MS. External calibration with matrix-matched standards was employed to calculate the concentration and the mass of Pt, Pd and Rh in each sample. 2M NaOH and 0.5 M HCl were added to each side to adjust the pH as needed. Diffusion cell experiments were carried out using filtered seawater, filtered river water, artificial seawater, and in 0.01M NaCl.

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Chapter 3: Development of the Diffusive Gradients in Thin-Films Technique (DGT) for platinum (Pt), palladium (Pd), and rhodium (Rh) in Natural Waters

Abdulbur-Alfakhoury, Ehab, Steve Van Zutphen, and Martine Leermakers. "Development of the diffusive gradients in thin-films technique (DGT) for platinum (Pt), palladium (Pd), and rhodium (Rh) in natural waters." Talanta 203 (2019): 34-48

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Abstract The Diffusive Gradients in Thin-Films (DGT) technique was developed for the first time for the platinum group elements (PGEs) platinum (Pt), palladium (Pd), and rhodium (Rh). Different chelating resins, specific designed for the accumulation of PGEs, namely Purolite S914, S920, S985, Italmatch Chemicals IONQUEST® MPX-317 and MP-102, were compared. The method development involved several different steps: 1) selection of an appropriate diffusive gel, 2) comparison of resins for the PGEs in terms of kinetics of uptake, 3) development of an efficient elution method for the PGEs from the resin gel, 4) Determination of diffusion coefficients for the PGEs in the diffusive gel, 5) Investigation of the influence of pH, ionic strength and dissolved organic matter on the diffusion coefficients and 6) study the selectivity of the tested resins gels in terms of potential interferences on the determination by ICPMS. Pt, Pd and Rh showed a linear accumulation over time for all resins and diffusion coefficients were independent of pH. The diffusion coefficient for Pt increased with increasing ionic strength (>0.5 M NaCl), but not for Pd and Rh. The interference study showed that Sr was the most important interferent for Rh and Pd and must be removed prior to analysis. The resins Purolite S914, Purolite S920, and Italmatch Chemicals IONQUEST® MPX-317 showed the best performance in terms of detection limits and separation of interferences. Using optimized procedures, the concentrations of Pt, Pd and Rh can be quantified at pg L-1 levels in natural waters for a 14 day deployment time.

Keywords

Platinum group elements (PGEs); Sector field inductively coupled plasma mass spectrometry (SF-ICPMS); Diffusive Gradients in Thin-Films (DGT); Surface water; Interferences

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3.1. Introduction

Several studies have reported increasing concentration of platinum group elements (PGEs, i.e. Pt, Pd, Rh, Ir, Ru, Os), especially Pt, Pd, and Rh in different environmental compartments (rain, surface and ground water, sea water, soils, sediments, sewage sludge, etc.) as a result of anthropogenic emissions [19, 20, 23, 104, 374]. Recent studies [27, 46, 375] on the emission and environmental occurrence of PGEs indicate that automobile catalysts may not be the single most important PGEs source, but also industrial processing, fossil fuel combustion and medical centres contribute significantly to the global PGEs budget at the Earth’s surface. The concentration of PGEs in river water are often below the analytical detection limits (0.1-0.5 ng L-1), and where measurable generally range from 0.4 to 10.2 ng L-1 near anthropogenically impacted areas [20, 103, 375] with a general pattern of Pd ≥ Pt > Rh [103, 376]. Platinum- based anticancer drugs in hospital result in Pt concentrations in waste waters ranging from 0.02 to 140 μg L−1 [45]. In the past, PGEs were previously considered to be present as relatively inert nano-particles, but recent studies have shown that these metals may undergo environmental transformations into more reactive species which may become bioavailable [5, 23, 109, 377]. The understanding of the mobility, pathways and bioavailability of elements requires information about their physio-chemical forms [378]. The determination of PGEs species in environmental samples is highly challenging due to problems associated with chemistry of PGEs in solution during sample handling and storage and problems related to the analysis itself [108, 144, 379]. Problems associated with the chemistry of PGEs, in particular at low concentrations, are the strong tendency of hydrolysis and precipitation [144, 328], the adsorption onto the surface of vessels under weakly acidic and neutral conditions [108, 328, 329], and transformations during storage and pre-treatment steps [108, 112, 113, 205]. PGEs can occur in various oxidation states, and in each oxidation state different complexes with inorganic and organic ligands can be formed depending on pH and ligand concentrations [108]. The most common oxidation state of Rh is (III) [8], whereas Pt and Pd can occur in either the (II) or (IV) valence state but the divalent state largely predominates over the tetravalent state at 25 ⁰C except under very oxidizing conditions [8]. Hydroxide, chloride, sulphide, ammonia and humic substances are identified as of possible importance in the complexation of Pt2+, Pd2+ and Rh3+ in the most common environmental compartments [8, 110, 328, 380, 381]. Inductively coupled mass spectrometry (ICP-MS) is the most widely used technique for the analysis of PGEs [143, 382] in environmental samples, either using a quadrupole mass spectrometer (ICP-Q-MS) [153] or sector field mass spectrometer (ICP-SF-MS) [143, 155,

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322]. Isobaric and polyatomic interferences are an important problem in ICP-MS analysis, especially when the analytes are at very low concentrations compared to the concentrations of the interfering elements. For the analysis of PGEs in environmental samples, the presence of Strontium (Sr), Rubidium (Rb), Copper (Cu), Zink (Zn), Hafnium (Hf), Lead (Pb) and Yttrium (Y) may cause interferences. The list of most important isobaric and polyatomic interferences for Pt, Rh and Pd is shown in the supplementary S1. Sector field (ICP-SF-MS) enables the separation of some of these polyatomic interferences by increasing the resolution settings of the mass spectrometer [88, 383]. Alternatively, selective extraction/pre-concentration of PGEs from the interfering matrix elements can be performed. In this context, the in situ passive sampling technique diffusive gradients in Thin-Films (DGT), invented in 1994 by Hao Zhang and Bill Davison [267, 268] pre-concentrates selected metals over the deployment time, improving detection limits and reducing matrix effects in environmental matrices such as water. The method is based on the accumulation of solute species on a resin or sorbent gel (a resin/sorbent immobilized in a thin hydrogel layer) after passing through a diffusive hydrogel layer (agarose or polyacrylamide) [267, 268]. The concentration gradient built between the bulk solution and the resin gel leads to a diffusive flux from the bulk solution and a pre- concentration of solutes in aquatic systems when the deployment time is sufficiently long. CDGT can be calculated with the help of equation 1.7. The role of the diffusive hydrogel is to allow diffusion of the solutes and the role of the sorbent/resin is the fast and irreversible binding of the solutes of interest. In the ideal case, interfering compounds are not bound on the resin/sorbent gel. The most commonly used resin gel incorporates Chelex®-100 resin with functional groups of iminodiacetate acid (IDA), which act as chelators for polyvalent metal ions [267, 268], but many other resins and sorbents have been used such as Diphonix for U[284, 285], MnO2 for Ra[286], 3-mercaptopropyl functionalized silica gel for Hg[287], Metsorb, ferrihydrite or ZrO for elements such as As, P, Mo, V[288]. For the PGEs the DGT technique has not yet been developed. Several resins are available which have a high selectivity towards PGEs. These include ion-exchange resins with polyamine [112-114, 205, 240, 241, 289], thiol [186], thiourea [186, 201, 215], isothioureum [186, 220], phosphine oxide [189, 201, 232] functional groups. Based on the hard and soft acids and bases theory (HSAB), the ion exchangers of the functional groups containing one or more soft-bases donor atoms interact strongly with soft acids. Pt (II), Pd(II) and Rh(III) ions are soft acids and show affinity for soft bases with donor atoms: O < N < S while hard metal ions such as Hf (IV), Zr (IV) show affinity for hard bases with the donor atoms: O > N > S [185, 186]. These resins are generally used for the recovery of PGEs from hydrometallurgical process

97 waters (refining processes, recycle catalytic converters etc.) consisting of acidic solutions with high chloride content or oxidizing solutions in the case of phosphine oxide resins [112-114, 186, 205, 220, 240, 241, 289] and their performance under neutral conditions and for a water composition corresponding to the composition of river water or seawater has seldom been investigated. The aim of the study was therefore to investigate the performance of selected resins with different characteristics (Purolite S985, S914, S920 with polyamine, thiourea and isothioureum functional groups, IONQUEST® MPX-317 and MP-102 with phosphine oxide thiourea and phosphine oxide functional groups) for their applicability as binding resin for DGT applications for Pt, Pd and Rh in natural waters. The influence of pH, ionic strength and organic carbon content was evaluated. A detailed interference study was performed as well as the study of the accumulation of interfering compounds on the resin gels and how these interferences can be removed.

3.2. Materials and methods

3.2.1. General Procedures All general procedures are described in section 2.1. 3.2.2. DGT preparation and assembly 3.2.1.1.Diffusive gel preparation Details were previously discussed in section 2.3.1. 3.2.1.2.Binding gels preparation investigation Following the literature study described in chapter 2, five resins (S914, S920, S985, MPX-317 and MP-102) with a high affinity and selectivity for PGEs were selected for detailed investigation and published in the manuscript. A number of other resins were also tested (MPX- 310, S924, MP-317, MP-101) and the results are presented in the Annex. The properties of these resins are shown in Table 2.4. Details were previously discussed in section 2.3.2. 3.2.1.3.DGT assembly Details were previously discussed in section 2.3.3. 3.2.3. Preparation of the deployment solutions A major issue in the development of DGT for PGEs is the stability of the PGE solutions, especially at low salinities and neutral or alkaline pH. Preparation of PGE spiked solutions is best performed in glass containers which have been equilibrated with PGE spiked solutions for at least 3 days to avoid adsorption on the container walls. This solution was then discarded, and a freshly prepared test solution was placed in the containers. This procedure used for all

98 tests to avoid loss of the PGEs due to adsorption on the container surface, especially in PE containers. Glass containers showed a much higher stability. The pre-treated glass containers showed an insignificant loss of PGE’s, even at low salinities. Deployment solutions consisted of 0.01M, 0.1M, 0.5M NaCl matrix, at a concentration level of 10 µg L-1 and 1000 µg L-1 and in spiked filtered seawater (North Sea, Belgian coast), spiked filtered river water (Zenne River, Belgium) and artificial seawater. Artificial seawater was composed of 27.5g NaCl, 6.78g

MgSO4 H2O, 5.38g MgCl2.6H2O, 1.4g CaCl2.2H2O, 0.72g KCl and 0.2g NaHCO3 per litre. Commercially available sea salt (Instant Ocean, Aquarium Systems, France) was used in a later stage and was also used to prepare to low salinity solutions (0.03 M sea salt). 3.2.4. Sample analysis Sample analysis was performed using Inductively Coupled Plasma Sector Field Mass Spectrometry (SF-ICP-MS, Element II, Thermo Fisher Scientific Bremen GmbH, Germany), as discussed in detail in section 2.2

3.2.5. Characterization of DGT performance 3.2.5.1.Chemical interactions between diffusive gels and PGEs An essential pre-requisite of the DGT procedure is that there is no reaction between the diffusive gel and the measured solute. In order to test the absence of a chemical reaction between the gel membranes and solute, 10 PAM diffusive gels (0.8 mm) and 10 AGA gels (0.75 mm) were exposed to continuously shaken seawater spiked with PGEs (15 µg L-1) for 6, 12, 24, 48, and 72 hours. Spiked seawater was used due to the high stability of the PGE solutions in this matrix. At each interval gel discs were taken out (n=2), washed with MQ water and eluted with 2 mL of 1.6 M HNO3 for 24 h. A 1 mL subsample of the deployment solution was taken at each sampling interval and acidified with 0.3 mL of 14 M HNO3 for analysis and at the day of analysis is diluted up to 10 mL. 3.2.5.2. Uptake kinetics on resin gels In order to determine the uptake kinetics of Pt, Pd and Rh simultaneously by the various binding gels (S914, S920, S985, MPX-317 and MP-102), three resin gels were placed in 50 mL polypropylene tubes containing 20 mL of 1 mg L-1 Pt, Pd and Rh of each in artificial seawater matrix, and 0.1 mL samples were taken over a period of 48 h. The pH of each solution was adjusted with 2 M NaOH to 8.1± 0.2, and a temperature of ∼23 °C was maintained throughout the experiment. A control vial without gel showed that there was no significant loss due to adsorption on the container walls.

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3.2.5.3. Elution factors Two different elution methods were evaluated: aqua regia at 70°C for 24h and thiourea (0.66 M in 0.3 M HCl). In order to determine the elution factor, resin gels were spiked with 50 µl of 1 mg L-1 of a mixture of Pt and Pd and Rh (in 0.24M HCl) on the surface of the resin gel disc and left for 24h. For the aqua regia procedure 1 mL of freshly prepared aqua regia solution was added to the resin gel in a Teflon bottle and placed in an oven at 70°C for 24 h. After cooling, 8.95 mL of MQ water was added. For resin gels that did not dissolve completely in aqua regia the solution was filtered using 0.45µm disposal mixed cellulose ester syringe filters

(Chromafil) pre-washed with 10 mL of 1.4 M HNO3. For the thiourea procedure 1 mL of thiourea (Merck) 0.66 M in 0.3 M HCl was added to the spiked discs and shaken for 24h, then 8.95 mL of MQ water was added then the resin gel separated. For the ICPMS measurements, calibration curves were made in either a 0.066 M thiourea/ 0.03 M HCl matrix or a 10% aqua regia extracted blank resin gel matrix to obtain matrix-matched standards to calculate the recovery. The tests of PGEs elution efficiencies were conducted in 3 replicates. The elution factor (fe) was calculated as the ratio of mass recovered through elution from the binding gel compared to mass spiked. 3.2.5.4. Diffusion cell experiments

The procedure of Zhang and Davison for the determination of diffusion coefficients (Dcell) in agarose gels using a diffusion cell is described in detail in section 2.3.4. Diffusion cell experiments were carried out using filtered seawater, filtered river water, artificial seawater, and in 0.01 M NaCl. 3.2.5.5. Time series DGT deployment: general procedures The performance of DGT for measuring Pt and Pd and Rh in natural waters was tested in filtered seawater and filtered river water spiked with PGEs at a concentration range of 10-20 µg L-1. In several experiments, artificial seawater, and 0.03 M seasalt solutions were used. The solutions were left for at least 24h for equilibration. The uptake of PGEs over time were tested using DGT units with a 0.75-mm-thick agarose diffusive gel layer and filter (0.17 mm) overlaid on a 0.5-mm-thick binding resin gel. The DGT probes (n=2 at each sampling-time) were exposed for 6, 12, 24, 48, 72 hours in the 2 L stirred-solutions spiked with PGEs. At each sampling interval, a subsample of the deployment solution was taken and acidified with 0.3 mL of 14M HNO3 for analysis. The DGT assemblies were rinsed with MQ water immediately after retrieval. The pistons were disassembled, and the resin-gels transferred into a Teflon bottle and eluted as described in section 2.5.3. The slopes of the linear regressions of accumulated

100 mass in function of time were used to determine the effective diffusion coefficients, D (cm2s- 1), of PGEs using equation 1.7. after rearrangement. 3.2.5.5.1. Effect of pH on uptake of PGE’s by the resin gels Using a 0.01M NaCl deployment solution, the pH of tested solutions was adjusted to cover a range of 5−8, with 0.5 M HCl (Merck, Germany) or 2 M NaOH (Merck, Germany). Temperature and pH were measured before, during, and after the 24h duration of the experiment. 3.2.5.5.2. Effect of ionic strength on uptake of PGE’s by the resin gels The ionic strength of solutions was adjusted to cover a range of 0.01−0.5 M with a solution of 2 M NaCl (Merck, Germany). The pH of each solution was adjusted using either HCl or NaOH, to obtain a pH between (6.5 -7) for the 24h duration of each experiment. 3.2.5.5.3. Influence of DOC on uptake of PGE’s by the resin gels Experiments were conducted using a 0.01M NaCl solution at pH 6.5 -7. The dissolved organic matter (DOM) concentration covered a range of 2.7−11.9 mg C L−1 by dilution of humic acid stock solution (540 mg C L−1, determined using a TOC analyzer). The humic acid stock solution was prepared by dissolving humic acid sodium salt (Sigma Aldrich) in 2 M NaOH, dilution with Milli-Q water and adjusting the pH between (6.5-7) with 0.5M HCl [384]. A 24h deployment time was used. 3.2.5.6. Binding gel selectivity and interferences For the interference tests, concentrations of the interferences were chosen to be comparable to concentrations found in natural samples. Ten DGT devices with binding gels made from S914, S920, S985, MPX-317 and MP-102 of each were deployed in 10 litter filtrated river collected from the study site and spiked with 10 µgL-1 of PGEs and spiked with Cu (12 µg L-1), Pb (10 µg L-1), Hf (0.7 µg L-1), and Y (0.5 µg L-1) due to their lower concentration while Sr and Rb were not added due to their high concentration in riverine water (Sr 280 µg L-1, Rb 76 µg L-1). The test solution was left for 48h for equilibration before the deployment. Master variables (T = 21 ⁰C, pH = 8.6, salinity = 0.4, DOC = 4.3 mg L -1) were measured before the deployment and at the end of the test. The DGT probes (n=2 at each sampling-time) were exposed for 24, 48, 72, 96, and 120 hours in the stirred-solutions. 3.3. Results and Discussion 3.3.1. Chemical interactions between diffusive gels and PGEs Figure 3.1 shows the relative accumulation of PGEs in the diffusive gels (mass of PGEs in diffusive gel/mass of PGEs in solution *100). Accumulation of Pt is negligible in both AGA and PAM gels, whereas both Rh and Pd show a significant increase in PAM gels in function

101 of the time (up to 2.5% for Pd after 72h) and negligible accumulation in AGA gels. The same has previously also been reported for Hg [385] and Au [386]. The results of this experiment led us to continue the further steps of the development using AGA as diffusive layer.

Figure 3. 1. The relative accumulated mass of Rh, Pd and Pt in diffusive gels in function of the time (n= 2).

3.3.2. Uptake kinetics of PGEs by binding gels Figure 3.2 shows the uptake kinetics of Pt, Pd and Rh from a spiked artificial seawater solution at pH 8.1± 0.2 on the different resin gels. MPX-317 shows a fast uptake of Pt and Rh and a somewhat slower uptake for Pd, whereas MP-102 shows a fast uptake for Pd compared to Pt and Rh. S985 and S920 show very fast uptake for all three elements, whereas S914 shows a very fast uptake for Pt and Pd and a slower uptake for Rh. Under the conditions tested in this experiment (filtrated seawater) quantitative uptake of PGEs (20µg of each metal) is observed after 48h for all resins with the exception of Pd on S985 and Rh on S914. The fastest simultaneous uptake of Rh, Pd and Pt is observed for S920. Approximately half of the spike is taken up in the first 30 min accounting for a flux to the resin gel of 300 ng min-1 of each metal.

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MPX-317 MP-102

Uptake (%) Uptake (%) 100 100

80 80

60 60

40 40

20 20

Time (min) Time (min) 0 0 0 1000 2000 3000 4000 0 1000 2000 3000 4000 Rh Pd Pt Rh Pd Pt

S914 S920 S985 uptake uptake uptake (%) (%) (%) 100 100 100

80 80 80

60 60 60

40 40 40

20 20 20

Time (min) Time(min) Time (min) 0 0 0 0 1000 2000 3000 4000 0 1000 2000 3000 4000 0 1000 2000 3000 4000 Rh Pd Pt Rh Pd Pt Rh Pd Pt

Figure 3. 2. Uptake kinetics of Pt, Pd and Rh on different resin gels (n= 3).

3.3.3. Elution factor Elution of PGE with 0.66 M thiourea (TU) in 0.3 M HCl gave excellent recoveries for the S920 resin, but for the other resins, only Rh could be extracted with thiourea. The elution factors of Pt and Pd with thiourea were around 0.1 or lower. A much more aggressive hot aqua regia (AR) extraction resulted in high recovery of all PGEs as shown in Table 3.2. An efficient eluant to extract PGEs, and other soft metals, from chelating resins containing thiourea or phosphine oxide groups is difficult to find due to the high stability of the metal–resin complexes [201, 215]. A 50-60% recovery of Pd from a resin containing thiourea has been reported using 1 M thiourea in 2M HCl [195] and complete recovery could be achieved only after ignition of the resin [215]. Better recoveries of Pd from MP-102 resin using thiourea were obtained using

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several consecutive extractions at 60°C and acidifying with HNO3 [189]. The binding of Rh to these resins is generally weaker than the binding of Pt and Pd [205] and this can explain the better recoveries found for Rh. However, due to the low recoveries found for our TU extraction and high recoveries by using AR we used the AR procedure further in this work.

Table 3. 1. Elution factor fe

Rh Pd Pt Resin gel AR TU AR TU AR TU S914 0.990 ± 0.017 0.732 ± 0.030 0.935 ± 0.033 0.018 ± 0.003 0.910 ± 0.026 0.081 ± 0.012 S920 0.954 ± 0.012 0.911 ± 0.021 0.916 ± 0.029 0.975 ± 0.015 0.811 ± 0.007 1.03 ± 0.02 S985 0.831 ± 0.005 n 0.919 ± 0.001 n 0.817 ± 0.009 n S924 0.92±0.02 n 0.90±0.02 n 0.89± 0.05 n MPX-310 0.874±0.035 n 0.914±0.025 n 0.9399± 0.052 n MPX-317 0.899 ± 0.011 0.759 ± 0.027 0.866 ± 0.012 0.096 ±0.012 0.942 ± 0.042 0.071 ±0.035 MP-317 0.98±0.035 n 0.95±0.025 n 0.98± 0.052 n MP-101 0.91±0.005 n 0.92±0.03 n 0.95± 0.035 n MP-102 0.880 ± 0.020 0.868 ± 0.046 0.972 ± 0.017 0.191 ± 0.019 0.976 ± 0.010 0.126 ± 0.019 n; not tested. AR: aqua regia. TU: Thiourea. Results are presented as (Mean ± standard deviation of 3 replicates)

3.3.4. Diffusion coefficients using the diffusive cell Typical results of diffusion cell experiment are shown in Figure 3.3. Using equation2.7 after rearrangement, the diffusion coefficient of Pt, Pd and Rh obtained from Figure 3.3 are (4.70×10-06, 3.38×10-06, 2.90×10-06) cm2 s-1, respectively at filtrated seawater matrix of pH 8.3±0.2, salinity = 33.2, DOC ˂1 mg L-1 and a temperature of 19.9°C. Applying the Stokes−Einstein equation for temperature correction described in [387], the diffusion coefficient of Pt, Pd and Rh were calculated to be 5.48× 10-06 , 3.94× 10-06 , 3.39× 10-06, cm2 s-1 at pH 8.3±0.2, salinity = 33.2 and temperature 25 °C. No value could be found in literature for comparison, for either aqueous solutions or hydrogels. The average diffusion coefficients based on 3 replicates of the diffusion cell are shown in Table (3.3). For all three elements, the diffusion coefficients in artificial seawater or filtered seawater are about 30% higher than in 0.01M NaCl or in filtered river water. This may be due to speciation changes of the dissolved 0 species from neutral hydrated hydroxy species in freshwater ([Pt(OH)2(H2O)2] , 0 0 - 2- 2- [Pd(OH)2(H2O)2] , [Rh(OH)3(H2O)3] to chloride complexes in seawater (PtCl4 , PtCl6 PdCl4 -3 , RhCl6 ). The difference between the artificial solutions and spiked natural waters samples may be due to the presence of organic ligands forming complexes with the metals.

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Receptor 1000 Mass (ng) 800

600

400

200 Time (sec) 0 0 2000 4000 6000 8000 10000 12000 Rh Pd Pt y = 0,043x + 98,235 y = 0,047x - 10,224 y = 0,0707x + 109,35 R² = 0,9925 R² = 0,998 R² = 0,9981 Figure 3. 3. The accumulation of mass of PGEs in the receptor compartment of a diffusion cell over time. The experimental conditions for the data were as follows: pH of source and receiving solution = 8.3 ±0.2, salinity = 33.2, DOC ˂ 1 mg.L-1, source solution concentraion.

Table 3. 2.Diffusion coefficients (mean ± standard deviation of 3 replicates, ×10-6 cm2 s-1, at 25⁰C) obtained in the diffusion cell experiments in seawater: artificial seawater (ASW), filtrated seawater (FSW) and in fresh water: 0.01M NaCl and filtrated river water (FRW).

Seawater

DCell_ASW DCell_FSW Pt 5.95 ± 0.80 5.38 ± 0.11 Pd 5.14 ± 0.36 3.59 ± 0.32 Rh 4.79 ± 0.45 3.47 ± 0.06 Fresh water

DCell_0.01M NaCl DCell_FRW Pt 4.27 ± 0.16 4.07 ± 0.21 Pd 3.95 ± 0.18 1.94 ± 0.11 Rh 3.65 ± 0.64 2.98 ± 0.23

3.3.5 Effective Diffusion coefficient measurements using time-series DGT deployments For a DGT binding phase to be valid for a particular analyte it must demonstrate a linearity of mass accumulated over the deployment time, implying that the binding to the resin is strong, irreversible, almost instantaneous and that the accumulated metal amount is well below the capacity of the resin [269]. The mass of PGEs accumulated on the resin gel after 72h (600-800 ng) is much less than the amount accumulated during the uptake kinetics experiment (20µg) indicating that the capacity is not an issue and uptake rate (300 ng min-1) is fast enough to maintain a constant flux to the resin gel. The accumulation of PGEs using the five resin gels samplers was assessed for a 72 h deployment in different solutions: spiked filtrated river water

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(FRW) at salinity = 0.5, pH =8.6, T=19 ⁰C, DOC= 4.2 mg L-1, spiked filtrated seawater (FSW) at salinity=33.6, pH=8.4, T=19 ⁰C, DOC ˂ 1 mg L-1, artificial seawater at salinity=33, pH=8.1, T=21⁰C, DOC ˂ 1 mg L-1 and 0.03M sea salt at salinity=1.7, pH=6.5, T=19⁰C, DOC ˂ 1 mg L-1. Figure 3.4 shows an example of the DGT time series experiment for Pt, Pd and Rh measurements. Using eq 2.7, and applying the Stokes−Einstein equation for temperature -6 2 -1 correction [270], the effective diffusion coefficients DDGT (×10 cm s , at 25 ⁰C) were determined and the results are summarized in Table 3.4. Pd 700 Mass (ng) 600

500

400

300

200

100 Time (h) 0 0 10 20 30 40 50 60 70 80 S914 S920 S985 MPX 317 MP 102 y = 7,8028x - 1,7584 y = 7,9032x + 2,4807 y = 5,3071x + 4,3325 y = 7,0182x + 2,6469 y = 3,0932x + 0,7569 R² = 0,9992 R² = 0,9935 R² = 0,9963 R² = 0,9979 R² = 0,9925

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700 Rh Mass (ng) 600

500

400

300

200

100

0 Time (h) 0 10 20 30 40 50 60 70 80 S914 S920 S985 MPX 317 MP 102 y = 5,216x + 10,058 y = 8,2756x + 25,199y = 8,3747x + 29,43y = 8,3031x + 24,689 y = 4,4679x + 13,891 R² = 0,9959 R² = 0,99 R² = 0,9809 R² = 0,9924 R² = 0,9861

1000 Pt Mass (ng) 800

600

400

200

0 Time (h) 0 10 20 30 40 50 60 70 80 S914 S920 S985 MPX 317 MP 102 y = 10,049x + 29,281y = 9,4538x + 22,128y = 9,8821x + 50,806y = 11,913x + 29,742 y = 8,1077x + 8,6318 R² = 0,9956 R² = 0,9959 R² = 0,9819 R² = 0,9923 R² = 0,9914 Figure 3. 4. Time-series experiment (The accumulated mass in function of time). The diffusion coefficients D (cm2s-1), of Pt, Pd and Rh in the diffusive layers Δg = 0.092 cm (the agarose diffusive gel (0.75 mm) and the filter (0.17 mm)) using DGT units (n=2 at each time) with AGA diffusive gel layers overlaid on binding gel of AGA with S914, S920, S985, MPX-317 and MP-102 and deployed in filtrated river water (FRW) spiked with Pt, Pd, and Rh 24h before the deployment. The experimental conditions for the data were as follows: pH~8.64, salinity = 0.5. Area of window (A) = 3.14 cm2, deployment solution concentration (C= ng mL-1) for each resin gel test (Pt, Pd, Rh) (average concentration ± standard deviation over the 72h): S914 (19.9±1.7,14.0±1.1, 14.7±0.8), S920 (19.87±1.7, 16.8±1.22, 23.4±1.1), S985 (19.9±1.7, 13.9±0.5, 23.4±1.0), MPX-317 (19.9±1.7, 14.2±0.9, 23.4±1.1), MP-102 (12.4±1.0, 9.1±0.6, 14.9±0.7); temperature 19°C; diffusive gel thickness and filter Δg = 0.092 cm; graph gradient of mass accumulation over time (α) = ng/h. Mean values and standard deviation (error bars) of duplicate measurements are given.

107

Table 3. 3. Shows the Diffusion coefficients (×10-6 cm2 s-1, at 25⁰C) with different binding gel obtained in time series experiments in artificial seawater, filtrated seawater, 0.03 M sea salt and filtrated river water. The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution Artificial seawater S914 S920 S985 MPX-317 MP-102 Pt 5.13 ± 0.45 5.56 ± 0.55 Pd 4.13 ± 0.43 4.87 ± 0.83 Rh 2.85 ± 0.10 2.29 ± 0.55

SEAWATER S914 S920 S985 MPX-317 MP-102 Pt 5.80 ± 0.82 5.10 ± 0.59 4.18 ± 0.42 6.00 ± 0.68 5.871 ± 0.392 Pd 4.39 ± 0.66 4.27 ± 0.42 2.05 ± 0.17 2.63 ± 0.23 2.812 ± 0.726 Rh 2.81 ± 0.11 3.11 ± 0.16 2.74 ± 0.19 2.78 ± 0.31 3.211 ± 0.429

0.03M sea salt S914 S920 S985 MPX-317 MP-102 Pt 4.02 ± 0.33 5.13 ± 0.29 4.68 ± 0.55 4.98 ± 0.34 6.00 ± 0.67 Pd 5.19 ± 0.25 4.87 ± 0.33 4.48 ± 0.42 3.95 ± 0.22 4.62± 0.53 Rh 2.45 ± 0.22 2.52 ± 0.22 3.34 ± 0.26 (1.09 ± 0.07) (1.22 ± 0.12)

Filtrated river water S914 S920 S985 MPX-317 MP-102 Pt 5.19 ± 0.55 4.99 ± 0.43 5.19 ± 0.49 5.66 ± 0.55 5.94 ± 0.54 Pd 4.22 ± 0.13 4.20 ± 0.32 3.32 ± 0.17 2.80 ± 0.14 3.15 ± 0.44 Rh 2.69 ± 0.17 3.30 ± 0.24 3.33 ± 0.32 3.31 ± 0.21 2.63 ± 0.21

Mass accumulation over time was linear for each experiment with R2 values > 0.95 obtained Figure 3.4. This confirms that the new binding gels functioned in accordance to the assumptions of the DGT equation 1.7. and that PGEs species were being taken up irreversibly.

The Pt diffusive coefficient measured with the diffusion cell, (Dcell,) and the effective diffusion coefficients (DDGT) using DGT in seawater show similar results for all resins with exception of

S985 where a slightly lower DDGT is obtained and this could be explained by the fact the adsorption ability of this resin for Pt and Pd is affected by the competing effect of the high chloride concentration as the resin is an anion exchange resin [205]. In river water, the Pt DDGT is comparable to Dcell, but lower than in seawater; whereas DDGT in 0.01M NaCl are very similar for all resins and comparable to DDGT in seawater. For Pt, at low chloride concentrations

(<0.1mM) and typical freshwater pH, the dominant forms of Pt are Pt(OH)2 for Pt (II) and − - Pt(OH)5 for Pt (IV) [188], and Pt(OH)2 predominates over Pt(OH)5 [104]. Chloride and - 2- hydroxychloride complexes (PtCl3OH , Pt(OH)nCl6-n ) appear at chloride concentrations > - 2- 0.4mM and the complexes (PtCl4 , PtCl6 , dominate in seawater (chloride > 0.7 M).

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For Pd, the effective diffusion coefficients DDGT using resin gels of S914 and S920 are comparable between river water and seawater and within the range found for Dcell. DDGT using resin gels MPX-317, MP-102 and S985 somewhat lower, both in seawater as in river water. The relatively slower binding kinetics of Pd on MPX-317, binding of Sr on MP-102 (discussed in section 3.8) and competition of chlorides on the anion exchange resin S985 may explain the difference observed. The inorganic aqueous speciation of Pd is comparable to that of Pt: in 0 2- river water it is dominated by the neutral form: [Pd(OH)2(H2O)2] [388]; whereas PdCl4 2- predominates over PdCl3OH throughout the normal pH range of seawater [8, 389]. Although the equilibrium chemistries of Pt(II) and Pd(II) are comparable, their kinetic behaviour is very different with complexes of Pt(II) and Pt(IV) being much more kinetically inert. This may explain the difference between seawater and river water found for Pt, but not for Pd.

The Rh effective diffusive coefficient DDGT using DGT units with different binding gels did not show a significant difference between FSW and FRW and are comparable for all resin gels.

DDGT of S985 shows the same performance as the other resins’ gels in FSW and isn’t influenced by salinity as for Pt and Pd, which indicates that there is a preferential binding of Rh on the resin. MP-102 and MPX-317 show good results for Rh in spiked FRW or FSW but problems are encountered when artificial solutions with low salinity are spiked with Rh. The expected +3 -3 Rh species of spiked solutions range from the completely hydrated Rh(H2O)6 , to RhCl6 , n-3 with intermediate mixed aquo-chloro complexes,RhCl6-n(H2O)n , also present. The extent to which each complex exists depends primarily on the chloride concentration. It has been found -3 that, at high chloride concentrations ([Cl] ca. 0.1M), the main Rh species are RhCl6 and -2 2- RhCl5(H2O) , although RhCl4(H2O) can also occur. As the chloride concentration decreases the hydration reactions occur more readily. A rrecent study by [110] has shown that under equilibrium conditions, that is when the solution has been left to “age” Rh under typical 0 − freshwater pH (6–8), is composed by [Rh(OH)3(H2O)3] and [Rh(OH)4(H2O)2] and the “tentative” Rh(III) speciation in seawater under typical pH conditions, is dominated by − 2− hydroxylated complexes ([Rh(OH)4(H2O)2] and [Rh(OH)5(H2O)] ), while chlorides are only relevant at pH<7. These aqua hydroxy complexes are chemically inert and can influence the sorption behaviour on some resins [112, 113]. Overall, the resin S920 showed the least variation in the diffusion coefficients for all three metals in function of the sample matrix indicating an efficient uptake independent of the speciation of the metals or competing ions.

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3.3.6. The effects of pH, ionic strength, and DOM on uptake on DGT measurements The surface charge on the adsorbent particles, the protonation of ligands on the resin as well as the speciation of the PGEs are influenced by the pH [390], which may lead to a change in the performance of DGT resins. Figures 3.5 shows the ratio of DGT measured concentrations

(using the diffusion coefficients for the different resins, Table 3.4), CDGT, to the independently measured concentrations of the bulk solutions, CSol, over a pH range of 5-8. Between pH 5 and 8, Pt, Pd and Rh showed quantitative uptake onto tested binding gels, accounting for 100 ± 10% of the directly measured concentration. These results indicate that the Pt, Pd and Rh uptake efficiency by binding gels is independent of pH, demonstrating that the charge of the analyte or binding gels does not significantly affect uptake efficiency across the pH range studied.

Ionic strength

1,4 C /C 1,3 DGT sol 1,2 1,1 1 0,9 0,8 0,7 0,6 0,5 0,4 0,3 0,2 0,1 0 S914 S920 S985 MPX MP S914 S920 S985 MPX MP S914 S920 S985 317 102 317 102 Pt Pd Rh 0.01M (NaCl) 0.1M (NaCl) 0.5M (NaCl)

pH_effect

1,3 C /C 1,2 DGT sol 1,1 1 0,9 0,8 0,7 0,6 0,5 0,4 0,3 0,2 0,1 0 S914 S920 S985 MPX MP S914 S920 S985 MPX MP S914 S920 S985 317 102 317 102 Pt Pd Rh 5.1 6.02 7.03 8.05

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DOC_effect

1,3 C /C 1,2 DGT sol 1,1 1 0,9 0,8 0,7 0,6 0,5 0,4 0,3 0,2 0,1 0 S914 S920 S985 MPX MP S914 S920 S985 MPX MP S914 S920 S985 317 102 317 102 Pt Pd Rh 2.66 mg L-1 4.65 mg L-1 6.9 mg L-1 11.92 mg L-1 Figure 3. 5. Effect of pH, ionic strength and DOM on the ratio of DGT Pt, Pd and Rh concentrations, CDGT, to their concentrations in the bulk solution, C sol. Mean values and standard deviation (error bars) of triplicate measurements are given (n = 3).

Varying the ionic strength (using NaCl) of the solution over the studied range did not have a substantial effect on the concentration of Pt, Pd and Rh as measured by DGT (Figure 3.5). This is critical for environmental deployments of DGT, as the ionic strength of natural waters can vary from river water to seawater. Ionic strength needs to be measured when DGT with resin gels of S985 is used for Pt or Pd in natural waters, as high ionic strengths (≥ 0.1 M) can change the diffusion coefficient, D, used in calculating CDGT from eq 2.7 after rearrangement. The DOC may influence the DGT uptake by blocking the open pores of the diffusive gel or the forming of strong metal-DOC complexes hindering the metal binding on the resin gel [296]. As shown in Figure

−1 3.5, the ratios of CDGT/Csol shows that DOM in the range of 2.7−11.9 mg L has no effect on the Pt and Pd uptake by a DGT sampler indicating that the metal-DOC complexes are labile. For Rh, the ratios of

CDGT/Csol were lower than 1 (0.83 to 0.90) which may be due to the slow reaction kinetics of Rh (III) [23]. 3.3.7. The effect of interferences on ICPMS measurement and selectivity of binding gels Interferences on the ICPMS measurements were checked by using model solutions described in [143, 345, 346]. Using single element or mixed element solutions of the interfering compounds at appropriate concentration ranges; the intensity of the interfering species formed at another mass (e.g.179Hf16O+) is plotted in function of the intensity of the interfering isotope (e.g.179Hf) and from the linear relationship interference correction equations can be established. Analysis is performed in LR, MR and HR resolution modes. For103Rh, the most important interferences of concern are from 40Ar63Cu+, 87Sr16O+, 87Rb16O+, 36Ar67Zn+ and 206Pb2+; for 195 Pt this is 179Hf16O+ and for 105Pd interferences of concern are from

111

40 65 + 89 16 + 88 16 + 87 18 + 195 Ar Cu , Y O , Sr OH and Rb O (Table 2.3). In Figure 2.5. the interferences of Hf on Pt, of Cu, Zn, Pb, Sr and Rb on 103Rh and of Cu, Y, Sr and Rb on 105Pd are plotted. For DGT measurements, the importance of the interference will depend on the uptake of the interfering element on the resin gel; thus the selectivity of the resin towards the PGEs. In Figure 3.6, the accumulation of the interfering elements on the different resin gels is plotted. 3.3.7.1. Pt The analysis of Pt is interfered by Hf (Figure 2.5) and the apparent 195Pt concentrations accounts for 1% of the Hf concentrations in LR and MR and this interference can completely be removed in HR mode. Deployment of the resin gels in filtrated river water spiked with Hf at a concentration level of 0.778 µg L-1 shows a linear increase in the concentrations of Hf on the resins gels S914, S920, S985 and MP- 102, whereas the resin MPX-317 shows very little accumulation of Hf (Figure 3.6). The affinity for Hf followed the order S985 > S914 > S920 > MP-102 >> MPX-317. Concentrations of Hf in river waters generally range from 0.004 to 0.11 µg. L-1 [391], and in this study, the average measured spiked Hf in river water samples was 0.778 µg.L-1 and resulted in accumulated mass at time 120 h for S985 binding gel of 3.21 ng which causes an apparent concentration increase of Pt of 0.6ng L-1. Both the application of interference corrections or measurements in the HR mode can resolve this interference, but as MPX- 317 shows very little affinity towards Hf; this is the preferred resin in terms of selectivity for Pt. 3.3.7.2. Rh 103Rh measurements by ICPMS can mainly be interfered by 206Pb2+, 40Ar63Cu+, 36Ar67Zn+, 87Sr16O+, 87Rb16O+. Figure 2.5. shows the apparent 103Rh concentrations in function of the concentrations of interfering elements. The apparent 103Rh concentration accounts for 0.006% of the Cu concentration in LR and MR and can be completely resolved in HR. The interference of Sr accounts for 0.02% of the Sr concentration in all resolutions. For Pb, Zn and Rb, no increase in the Rh signal was observed with increasing concentrations of interfering elements. All resin gels show a linear accumulation of Pb, Zn and Cu in function of time (Figure 3.6), but the interference of Pb and Zn on the Rh signal is negligible. For Cu, this results in an apparent 103Rh concentration of 0.6 ng L-1 after 120h in LR and MR and this can be resolved in HR mode. Rubidium showed a constant value on the tested binding gels of MPX-317, S914, S920 and S985 (around 25 ng Rb) whereas an increase on MP-102 with increasing the deployment time was observed (up to 60 ng after 120 h). However, the interference of Rb on the Rh signal is negligible Figure 2.5. Comparable to Rb, Sr showed a constant value in the resin gels MPX-317, S914, S920 and S985 (around 50 ng Sr) whereas an increase on MP-102 with increasing the deployment time was observed (up to 200 ng after 120h). This results in an apparent Rh concentration of 0.15 ng L-1 (0.6 µg L-1 for MP-102) for a Sr concentration in water of 28 µg L-1. Strontium concentrations in natural waters can reach 8000 µg L-1, making it necessary to either apply mathematical corrections or remove Sr from the eluent solution before measurement.

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3.3.7.3. Pd 105Pd measurements by ICPMS are interfered mainly by 40Ar65Cu+, 87Sr16OH+, and 89Y16O+. Figure 2.5. shows the apparent 105Pd concentrations in function of the concentrations of interfering elements. The apparent 105Pd concentration accounts for 0.02% of the Cu concentration in LR and MR and can be completely resolved in HR. The interference of Y accounts for 1% of the Y concentration and interference of Sr for 0.1% of the Sr concentrations and both interferences cannot be resolved by increasing the resolution settings. All binding gels show a linear accumulation of Cu in function of time (Figure 3.6). The accumulation is highest for S985 and S920, followed by MP-X317, S914 and MP- 102. A Cu concentration in solution of 12 µg L-1 results in an accumulation of 500 ng Cu on the resin gel after 120h, accounting for an apparent Pd concentration of 2 ng L-1. This interference can be resolved in HR mode. Y shows a linear accumulation over time for MP-102, whereas the other resins show a strong deviation from linearity after 48h with the lowest uptake found for S914. Y in river waters range from 0.05–1.40 µg L–1 [391], and in our experiments a spike of was 0.515 µg.L-1 used which caused an apparent Pd concentration of 2 ng L-1 for MP-102 and 0.3 ng L-1 for S914. For the Y interference, S914 is the most suitable resin and the application of mathematical corrections is necessary. As already mentioned for Rh; the accumulation of Sr shows a constant value in function of deployment time with exception for MP-102 which shows a linear increase of Sr in function of time. This accumulation accounts for an apparent Pd concentration of 7 ng L-1 for S914, S920, S985 and MPX- 317 resin gels and 28 ng L-1 for the MP-102 resin gel for a Sr concentration of 280µg L-1 in the deployment solution.

Rb Cu 80 Mass (ng) 700 Mass( ng) 70 600 60 500 50 400 40 30 300 20 200 10 100 Time (h) Time (h) 0 0 0 50 100 150 0 50 100 150 S914 S920 S985 MPX 317 MP 102 S914 S920 S985 MPX 317 MP 102

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3000 Zn Sr Mass (ng) 300 Mass(ng) 2500 250 2000 200 1500 150 1000 100 500 50 Time (h) Time (h) 0 0 0 50 100 150 0 50 100 150 S914 S920 S985 MPX 317 MP 102 S914 S920 S985 MPX 317 MP 102

Pb 200 Mass(ng) 14 Y Mass (ng) 12 150 10 8 100 6

50 4 2 Time (h) 0 Time (h) 0 0 50 100 150 0 50 100 150 S914 S920 S985 MPX 317 MP 102 S914 S920 S985 MPX 317 MP 102

4 600 Mass (ng) Hf Mass (ng) Rh

3 400

2 200 1 Time (h) 0 Time (h) 0 0 50 100 150 0 50 100 150 S914 S920 S985 MPX 317 MP 102

S914 S920 S985 MPX 317 MP 102

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600 Mass (ng) Pd 800 Pt 500 Mass (ng) 600 400 300 400 200 200 100 Time(h) Time (h) 0 0 0 50 100 150 0 50 100 150 S914 S920 S985 MPX 317 MP 102

S914 S920 S985 MPX 317 MP 102

Figure 3. 6. The uptake (ng) of Pt, Pd, Rh, Zn, Cu, Sr, Rb, Y, Hf and Pb over time using DGT units with diffusive layers (agarose diffusive gel (0.75 mm) and filter (0.17 mm)) overlaid on binding resin gel of AGA with S914, S920, S985, MPX-317 and MP-102 and deployed in spiked filtrated river water. The experimental conditions for the deployment solution were as follows T = 21 ⁰C, pH = 8.6, salinity = 0.4, DOC = 4.3 mg L -1. Mean values and standard deviation (error bars) of duplicate measurements are given (n =2).

Removal of Sr and Rb from the resin gels Due to the high concentrations of Sr in surface waters (10-1300 µgL-1), the high Sr accumulation on the binding gels and the important interference of Sr on Rh and Pd SF- ICPMS- measurements (Figure 2.5), the application of mathematical corrections for the Sr interference will results in large uncertainties on the measurement[108, 352] making it the necessary to remove this interference from the binding gels prior to the ICPMS measurement. The accumulation of Rb and Sr on the binding gels S914, S920, S985 and MPX-317 shows a constant accumulation over the deployment time (Figure 3.6), which could be explained either by saturation of the binding sites or an equilibrium effect between resin gel and diffusive gel [392]. As the linearity of the accumulation of Pt, Pd and Rh is not lost whereas the accumulation of Rb and Sr is constant during the same deployment time, this cannot be due to saturation of the binding gels. As Sr and Rb are present in high concentrations in the deployment solution, they can diffuse into the binding gels without binding on the resins as the resins have little affinity for those metals. If this is the case, washing the resin gel with Milli-Q water prior to extraction could remove those metals from the binding gels. To verify this, 10 DGT units consisting of agarose diffusive gel (0.75 mm) and filter (0.17 mm) overlaid on binding resin gel of agarose (0.5 mm) with resin S914 were deployed in 2 litre filtrated seawater with a Sr concentration of 8 mg L-1 and a Rb concentration of 150 µg L-1 and spiked with Pt, Pd, Rh at a concentration around 10 µg L-1 and continuously stirred. The DGT probes were exposed for 16 hours in the stirred-solutions. After 16h, the DGT assemblies divided into two groups. The first five gels (No-wash) were eluted immediately as described in section 2.5.3. The second group

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(Washed) were washed for 12h with 10 mL Milli-Q water on a shaker then resin gels were separated from the MQ water and eluted as described in section 2.5.3. After elution the eluents were measured and the results are shown in Figure 3.7. After the 12h Milli-Q water wash, the accumulated mass of Pt, Pd and Rh did not show any difference between the washed and unwashed resin gels and this indicates that the wash does not cause a leach of the PGE metals from the resins gels (Figure 3.7). At the concentration levels of the PGE spike (µg L-1) the effects of the interferences are insignificant. More than 95% of Sr and Rb are removed by a single 10 mL Milli-Q water wash indicating that the resin gels S914 has very little affinity for Sr and Rb and they enter the resin gels by diffusion (Figure 3.7).

200 1800 Mass Mass 180 1600 (ng) (ng) 160 1400 140 1200 120 1000 100 800 80 60 600 40 400 20 200 0 0 Rh Pd Pt Sr Rb

Washed No-wash Washed No-wash Figure 3. 7. The accumulated mass (ng) (mean ± the standard deviation of 5 replicates), of Rh, Pd, Pt, Rb and Sr using DGT units (n=10) of diffusive layers (agarose diffusive gel (0.75 mm) and filter (0.17 mm)) overlaid on binding resin gel of agarose (0.5 mm) with Purolite S914 and deployed in filtrated seawater spiked with Pt, Pd and Rh, 24h before the deployment. 5- Binding gels were washed for 12 h with 10 mL MQ-water before the elution (washed). 5- Binding gels were eluted immediately after the deployment without any wash (No-wash). Recovery factor is considered 1 for Sr and Rb.

The experiments in spiked river water were repeated but now a Milli-Q water rinse was added prior to the elution step. Figure 3.8 shows the accumulated mass over the time after 12h Milli-Q water wash for all resin gels before the elution step. The results show a very slight Sr accumulation on S920 and MPX-317, reaching an equilibrium value of 2.5 ng Sr (compared to 1500 ng without wash), whereas the blanc value of S914 is slightly higher (5 ng) and this value is also found at the end of the deployment. The resin MP-102 on the other hand binds 100 times more Sr and Rb which cannot be removed by the Milli-Q wash. Further

116 investigation showed that using 3 consecutive washes with Milli-Q completely removed the Sr and Rb interference from S914, S920, MPX-317 resin gels, whereas it was harder to remove from S985 and impossible to remove from MP-102. Thus, resins S914, S920, MPX- 317 are the most suitable resins regarding the removal of interferences.

6 Sr MP-102 Mass 200 Mass (ng) (ng) 4

100 2

Time (h) Time (h) 0 0 0 20 40 60 80 0 20 40 60 80 Sr Rb S914 S920 MPX-317

Mass 2 Rb (ng)

1

0 Time (h) 0 20 40 60 80 S914 S920 MPX-317 Figure 3. 8. The uptake (ng) of Sr and Rb over time using DGT units with AGA diffusive gel layers (0.75mm) overlaid on binding resin gel of AGA with S914, S920, MPX-317 and MP- 102 and deployed in spiked filtrated river water. Resin gels were washed with MQ water for 12h before the elution step. Data of Sr and Rb are normalized to 113In. Recovery for Sr and Rb considered 1 and the dilution factor is 10. The experimental conditions for the deployment solution were as follows T = 21 ⁰C, pH = 8.6, salinity = 0.4, DOC = 4.3 mg L-1. STD is the for duplicate pistons measurement at each time.

3.3.8. Binding gel blanks and DGT detection limits. Table 3.5 shows the blank values obtained for an aqua regia extraction of the resin gels. The methodological DGT detection limit (MDL) for each element was calculated based on three times the standard deviation of 4 replicate measurements of blank resin gels and using eq 1 with a standard DGT configuration (3.14 cm2 exposure window area, 0.75 mm diffusive gel, 0.17 mm filter membrane), diffusion coefficient in riverine water at a temperature of 25 °C, diffusion coefficients in river water shown in Table 3.2. The data show that low pg L-1 detection limits can be achieved in using a 14-day deployment. Resin MP102 has higher blank values for Rh, Pd and Pt compared to the other resins and is thus less suitable for field studies.

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Table 3.4. DGT blank values (n=4) and method detection limits (MDL) for the different resin gels. MDLs are calculated based on a deployment time of 14 days, ∆g =0.092 cm, A=3.14 cm2, D river water at 25°C, pH=8±0.2. S914 S920 S985 MPX-317 MP-102 blank 0.013±0.001 0.0087±0.0016 0.008±0.0014 0.006±0.002 0.134±0.046 Pt (ng/disc) MDL (ngL-1) 0,007 0,007 0,007 0,014 0,335 blank 0.036±0.002 0.041±0.006 0.046±0,018 0.043±0.019 0.605±0.035 Pd (ng/disc) MDL (ngL-1) 0,017 0,051 0,153 0,161 0,297 blank 0.032±0.012 0.004±0.0002 0.013±0.002 0.009±0.006 0.033±0.014 Rh (ng/disc) MDL (ngL-1) 0,102 0,017 0,018 0,051 0,093

3.4. Conclusions

In this study the DGT technique was successfully developed for the analysis of Pt, Pd and Rh in natural waters. Agarose showed no interaction with PGEs and was thus used as diffusive layer and for the preparation of DGT binding layers. The five different resins gels, incorporating either Purolite S914, S920, S985, Italmatch Chemicals IONQUEST® MPX-317 and MP-102, showed a linear accumulation of PGEs in function of time, in both natural, spiked fresh water and seawater as in artificial solutions with ionic strength ranging from 0.01M to 0.5M NaCl and a pH range of 5 to 8. Elution factors ranging from 0.81 to 0.99 were obtained using an aqua regia extraction of the resin gels. The choice of most suitable resin for the different elements is thus based on the blank values of the resin gel, selectivity towards interfering elements and uptake kinetics. Hence, Pt- resin gel selection follows the order: MPX- 317 > S920 ≥ S914> MP-102 > S985. Pd-resin gel performance follows the order S914 ≥ S920> MPX-317 > S985> MP-102 and the Rh- resin gel selection follows the order S920 > S914 > MPX-317 > S985 > MP-102. The high concentrations of Sr in aquatic samples forms the most important interference in the analysis of Rh and Pd. The resins S920, S914 and MPX-317 show very little affinity towards Sr and Sr accumulation in the resin gel can easily be removed by three consecutive Milli-Q washes. The resins Purolite S920, S914 and Italmatch Chemicals IONQUEST® MPX-317 are thus the most suitable to apply in field studies. For the simultaneous determination of Pt, Pd and Rh with one resin, Purolite S920 is the most promising due to the low blank values, high uptake kinetics of all three elements and limited influence of matrix composition on the diffusion coefficients.

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Acknowledgements:

We would like to thank David Verstraeten for the measurement of DOC. We thank Bertrand Gallet (Purolite France), Valerie Treffkorn-Maura (Purolite France), for providing the resin samples. We thank Edgar Berreby (Purolite France) and Mikhail Mikhaylenko (Purolite Russia) for revision of the manuscript. We thank the Hercules Foundation for financing the ICP-SF-MS instrument (UABR/11/010).

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Chapter 4: Evaluation of the effect of solution ageing on the DGT speciation of Rh and Pt

Abdulbur-Alfakhoury, Ehab and M. Leermakers, “Evaluation of the effect of solution ageing on the DGT speciation of rhodium (Rh) and platinum (Pt)”. Journal of Analytical Atomic Spectrometry, 2021. DOI: 10.1039/D0JA00442A

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Abstract:

Synthetic solutions of Rh and Pt at near neutral pH, as well as spiked natural waters are subjected to an ageing effect, which means that over time the speciation of the PGEs will change from the original composition due to aquation and hydrolysis, forming various hydroxy and aqua complexes. Due to the slow ligand exchange reactions equilibrium speciation may not rapidly be attained. The ageing may affect the reactivity of the PGE solutions towards chelating resin used in the analysis of PGEs by Diffusive Gradients in Thin-Films (DGT). The aim of the study was therefore to investigate the performance of the chelating resins (Purolite

S914, S920 with thiourea and isothioureum functional groups, and IONQUEST® MPX-317 with phosphine oxide thiourea functional group) for their applicability as binding resin for DGT applications for Pt and Rh in natural waters using 17 days aged spiked solutions. Uptake kinetics experiments revealed that direct uptake of Pt on the resins from aged solutions was fast and quantitative, whereas only 30% of Rh was directly bound on the resin. Likewise, much lower apparent diffusion coefficients were obtained for Rh in aged spiked river water (70% decrease) which can be explained by the formation of Rh(OH)3(s) which is not taken up by the resin gel. Addition of dissolved organic matter results in an increased solubilisation of

Rh(OH)3(s) and an increased DGT labile fraction. Formation of Rh(OH)3(s) is less pronounced in seawater. For Pt, no species are formed which are not DGT labile. Thus, DGT accurately measures labile metal species in solution, discriminating labile from inert metal species.

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4.1. Introduction:

Platinum and Rhodium are among the least abundant trace elements in the environment, with a crustal concentration of around 0.54, and 0.06 ng g−1 respectively ]393[ and concentrations below 0.9 ng L−1 and 0.5 ng L−1 [103], respectively in water bodies. Near anthropogenically impacted areas, especially close to roads, industrial areas and medical centres enrichment over 103 − fold [20, 103, 375] is found leading to the need to develop reliable analytical methodologies for the determination of their environmental concentrations, mobility, reactivity, and biological impact. Although there are similarities in the applications, physicochemical properties and thermodynamic equilibrium chemistries between Pd, Pt and Rh, their kinetic behaviour towards ligand exchange processes are very different: ligands exchange processes for Pd (II) are very fast (within minutes); whereas those for Pt(II) and Pt(IV) are much slower (hours)[394, 395] and even days to weeks for Rh (III)[110, 116]. The differences in their kinetic behavior are even used to separate these elements in hydrometallurgical processes [396] and for Pt, the kinetically inert Pt complexes form the basis of their anticancer activity [397]. The slow complexation kinetics of Pt and Rh result in a so-called “ageing” at near neutral solutions spiked with Pt and Rh, which means that over time the speciation of the PGEs in a 2– 3– solution will change from its original composition ([PtCl6] and [RhCl6] in the stock solutions) due to aquation and hydrolysis in which different aqua chloro- and aqua hydroxo- complexes are formed. This may also influence sorption and elution of PGEs on anion exchange and chelating resins used to pre-concentrate the PGEs from solution [112, 396]. The in situ passive sampling technique Diffusive Gradients in Thin-Films (DGT), based on the preconcentration of solute species on a resin immobilized in a thin hydrogel layer after passing through a diffusive hydrogel layer [267, 268] was recently developed for PGE’s in water [221]. The method development was based on deployment in spiked synthetic solutions and spiked natural waters after 24h equilibration of spiked solutions. However, due to the slow exchange kinetics of Pt and Rh, full equilibration of the spiked Rh and Pt solution with inorganic and organic ligands present and the hydrolysis and aquation reactions may not have been fully achieved. In the present study, we evaluated the DGT characteristics using 17 days aged spiked solutions as deployment solutions and the resins (Purolite S914, S920 with thiourea and isothioureum functional groups, and IONQUEST® MPX-317 with phosphine oxide thiourea functional group. Effective diffusion coefficients were calculated in aged solutions and

122 compared with freshly spiked solutions. The influence of ionic strength and organic carbon content was evaluated using aged solutions. 4.2. Materials and methods

4.2.1. DGT probe preparation Experimental and analytical procedures are described in Chapter 2. Briefly, agarose was used as diffusive gel and for the preparation of resin gels. Three resins (Purolite S914, Purolite S920 and IONQUEST MPX-317) were used as binding resins and are presented in the manuscript. Tests were also performed on MPX-310 and are included in the Annex. The resins were grinded and sieved on 50µm before DGT applications. Millipore Durapore membrane filter (HVLP pore size 0.45 µm, diameter 2.5 cm, thickness 120µm) were used for all experiments. 4.2.2. Preparation of the deployment solutions Deployment solutions consisted of either of a NaCl matrix (0.01M - 0.5M), filtered seawater (North Sea, Belgian coast) or filtered river water (Zenne River, Belgium). Pt and Rh were spiked at a concentration level of 10 µg L-1 for DGT deployment tests or 1000 µg L-1 for uptake kinetics tests. For the DGT deployment tests, the pH of each solution was adjusted using either HCl or NaOH, in order to obtain a pH between (6.00 - 6.2) for the 24h duration of each experiment. Solutions were aged for 17 days after spiking Pt and Rh and kept under continuous stirring at a lab temperature around 20⁰C. The influence of dissolved organic matter was evaluated by addition of humic acid as described previously[109]. The dissolved organic matter (DOM) concentration covered a range of 4−12 mg C L−1. 4.2.3. Uptake kinetics on resin gels The uptake kinetics of Pt and Rh by the various binding gels (S920, S914 and MPX-317) were determined by deploying three resin gels in 50 mL borosilicate bottles containing 20 mL of 1 mg L-1 Pt and Rh of each in a 0.01 M NaCl matrix aged for 17 days, and 0.1 mL samples were taken over a period of 24 h. The pH of each solution was adjusted with 2 M NaOH to 6.1±0.2, and a temperature of ∼24 °C was maintained throughout the experiment. A control vial without gel showed that there was no significant loss due to adsorption on the container walls. 4.2.4. Effective diffusion coefficients using time series DGT deployment Effective diffusion coefficients for Pt and Rh were determined in spiked seawater and freshwater solutions at a concentration range of 10-15 µg L-1 which were allowed to age for 17 days before DGT deployment. The accumulation of Pt and Rh using three resin gels samplers was assessed for a 72 h deployment in different solutions: spiked filtrated river water (FRW)

123 at salinity = 0.4, pH =8.57, T=20 ⁰C, DOC= 4.3 mg L-1, spiked filtrated seawater (FSW) at salinity=35.9, pH=8.06, T=18 ⁰C, DOC ˂ 1 mg L-1. The uptake of PGEs over time was tested using DGT units with a 0.75-mm-thick agarose diffusive gel layer and filter (0.17 mm) overlaid on a 0.5-mm-thick binding resin gel. The DGT probes (n=2 at each sampling-time) were exposed for 6, 12, 24, 48, 72 hours in the 10 L stirred-solutions spiked with Pt and Rh. At each sampling interval, a subsample of 1 ml of the deployment solution was taken and acidified with

0.3 mL of 14 M HNO3 for analysis. The DGT assemblies were rinsed with MQ water immediately after retrieval. The pistons were disassembled, and the resin-gels transferred into a Teflon bottle and eluted in aqua regia at 70°C overnight as described previously [221]. The slopes of the linear regressions of accumulated mass in function of time were used to determine the effective diffusion coefficients, D (cm2s-1), of PGEs using equation (1.7) after rearrangement. 4.2.5. Sample analysis Sample analysis was performed using Inductively Coupled Plasma Sector Field Mass Spectrometry (SF-ICP-MS, Element II, Thermo Fisher Scientific Bremen GmbH, Germany), equipped with ESI fast auto-sampler. The isotopes 195Pt and 103Rh were measured in low resolution mode (106 cps/ppb 115In). 193Ir and 115In are used as internal standards for 195Pt and 103Rh respectively. Matrix-matched calibration curves were used for all types of matrixes during the study (seawater, fresh water, resin gel extracts, etc.). 4.3. Results and Discussion

4.3.1. Uptake kinetics of PGEs by binding gels Figure 4.1 shows the uptake kinetics of Pt and Rh from a spiked 0.01 M NaCl solution at pH 6.1± 0.2 on the different resin gels for solutions with different ageing times. For Pt, all resin gels showed the same performance characteristics at both ageing times with an average 10% lower uptake of Pt in 17 days aged solution compared to 1 day aged solutions[221]. Thus, despite the ageing process of the solutions, the resin gels still show fast uptake of all Pt species. This implies that all the species that are formed during the ageing process are DGT labile. In contrast, Rh shows significant different uptake characteristics between a 1-day aged solution and a 17 days aged solution for a 0.01 M NaCl solution, with only 40% of the dissolved species taken up by the resins. In low salinity solutions without organic matter, hydration reaction leads to the formation of Rh(OH)3(s) [110, 116], which shows poor binding on the resin gels. As will be shown in section 3.3. the presence of organic matter in natural waters leads to the partial solubilisation of Rh(OH)3(s) and the formation of more DGT labile complexes.

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Pt_ kinetics Rh_ kinetics 100 100 Uptake % Uptake %

80 80

60 60

40 40

20 20

time (min) time(min) 0 0 0 500 1000 1500 2000 0 200 400 600 800 1000 1200 1400 1600 S920 17 day aged S920 1 day aged S914 17 day aged S920 17 day aged S920 1 day aged S914 17 day aged S914 1 day aged MPX-317 17 day aged MPX-317 1 day aged S914 1 day aged MPX-317 17 day aged MPX-317 1 day aged

Figure 4. 1. Uptake kinetics of Pt and Rh on different resin gels and different solution ageing (1and 17) days (n= 3) at the same conditions (T = 24⁰C, pH = 6.1±0.2).

4.3.2. Effective Diffusion coefficient measurements using time-series DGT deployments Figure 4.2 shows the DGT time series experiment for Pt and Rh measurements in solutions aged for 17 days. Using equation 1.7., and applying the Stokes−Einstein equation for -6 2 -1 temperature correction [270], the effective diffusion coefficients DDGT (×10 cm s , at 25 ⁰C) in 17 days aged solutions were determined and compared to 24h aged solutions. The results are summarized in Table 4.1.

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Rh_ FRW_ 17 days Aged solution Rh_FSW_17 days Aged solution

100 350 Mass (ng) Mass (ng) 90 300 80 70 250 60 50 200 40 150 30 20 100 10 Time (h) 50 0 Time (h) 0 10 20 30 40 50 60 70 80 0 S920 S914 MPX 317 0 20 40 60 80 S920 S914 MPX 317 y = 1,2503x + 3,6836 y = 1,158x + 1,7968 y = 1,0334x + 0,8036 y = 4,4507x + 5,326 y = 2,9947x + 5,5557 y = 3,3672x + 13,99 R² = 0,9918 R² = 0,9888 R² = 0,9991 R² = 0,9945 R² = 0,9984 R² = 0,9904

Pt_ FSW_17 days aged solution Pt_ FRW _ 17 days aged solution 700 Mass (ng) 400 Mass (ng) 350 600 300 500 250 400 200 150 300 100 200 50 Time (h) 100 0 Time (h) 0 20 40 60 80 0 S920 S914 MPX 317 0 10 20 30 40 50 60 70 80 y = 4,583x - 5,2581 S920 S914 MPX 317 y = 2,9047x + 3,7534 y = 4,5533x - 5,4964 R² = 0,9993 y = 7,3115x + 1,6118 y = 7,5217x + 22,921 y = 7,2173x + 26,928 R² = 0,9976 R² = 0,9984 R² = 0,9989 R² = 0,9924 R² = 0,9928

Figure 4. 2. Accumulated mass of metals in function of time for deployments in 17 days aged solutions. Filtrated river water (FRW). Filtrated sea water (FSW).

Mass accumulation over time in the aged solutions was linear for each experiment with R2 values > 0.99 obtained Figure 4.2. This confirms that the binding gels functioned in accordance to the assumptions of the DGT equation 1.7., and that labile Pt and Rh species were being taken up irreversibly. For DGT to work at least one species must be DGT labile. If the dissociation of the non DGT labile species can occur in the diffusive gel during the diffusion process the species are DGT labile. Table 4. 1. Diffusion coefficients (×10-6 cm2 s-1, at 25⁰C) with different binding gel obtained in time series experiments in filtrated seawater and filtrated river water fresh obtained from our previous work [221] and aged for 17days after spiking with Pt and Rh. The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution Resin gel Filtrated seawater Filtrated river water Fresh Aged Fresh Aged S920 5.10 ± 0.59 4.65 ± 0.18 4.99 ± 0.43 4.33 ± 0.19 Pt S914 5.80 ± 0.82 4.78 ± 0.29 5.19 ± 0.55 2.76 ± 0.11 MPX-317 6.00 ± 0.68 4.59 ± 0.22 5.66 ± 0.55 4.36 ± 0.16 S920 2.81 ± 0.11 2.79±0.16 3.30 ± 0.24 0.98±0.05 Rh S914 3.11 ± 0.16 1.87±0.08 2.69 ± 0.17 0.90±0.06 MPX-317 2.78 ± 0.31 2.11±0.14 3.31 ± 0.21 0.80±0.018

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After 17 days ageing of spiked Pt in filtrated seawater, the effective diffusion coefficients

(DDGT) were similar for all resin gels and 20% lower than those obtained in freshly spiked solutions. A similar decrease in DDGT is also observed for spiked river water for the resins S920 and MPX-317, whereas an almost 50% decrease was observed for S914. The latter can however be due to the fact that an older batch of S914 was used in this experiment. As the differences are quite consistent between the resin gels, the decrease can probably be attributed to differences in the water composition of the deployment solutions (natural water samples) between the current and previous experiments (differences in organic matter content). 2− In the Pt stock solution, Pt is present as the hexachloroplatinate (IV) complex [PtCl6] . With the increase in pH, complexes of Pt(II) are formed and coexist with chloride complexes of Pt(IV). Additionally, aquation and hydrolysis occur with the decrease in acidity of the solution. As a result, different aquachloro-complexes and aquahydroxo-complexes of Pt(II) and Pt(IV) are formed during solution ageing [111, 114]. In freshwater, the thermodynamic data available for typical freshwater pH indicates that at equilibrium the dominant form of Pt is Pt(OH)2° for − Pt(II) over Pt(OH)5 for Pt(IV). At typical seawater pH (7.5–8.4) the dominant inorganic forms 2− 2- 2− 2- are Pt (IV) PtCl5(OH) and PtCl6 over Pt(II) PtCl4 and PtCl3(OH) [109, 110]. Our results show that all forms of Pt are adsorbed on the chelating resins, regardless of their charge and the speciation does not affect the diffusion in the diffusive gel.

The Rh effective diffusion coefficients (DDGT) in filtrated seawater after 17 days ageing also show a 20% decrease compared to 24h ageing for the resins S920 and MPX-317, whereas a 40% decrease is observed for S914. In filtrated river water, a much larger decrease is observed (around 70% for S920 and S914 and 75% for MPX-317). This indicates that non labile Rh species are more readily formed in low ionic strength solutions. The calculation of the diffusion coefficient requires the knowledge of the concentration of labile species in solution. However, as we can only measure the total dissolved concentrations in the test solution, this will lead to an underestimation of the diffusion coefficient. The decrease in effective diffusion coefficient corresponds to the fraction, which can be taken up by the resin gels directly (around 30%). If we use this estimation to calculate the labile fraction in solution, the calculated diffusion coefficients are comparable to those of the 1 day aged solutions. This means that the inert complexes either do not enter the diffusive gel or are not transformed into mobile species during the deployment time.

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In aged solutions at neutral pH, the Rh(III) speciation is governed by the slow formation of

Rh(OH)3(s), comparable to what has been observed for Cr(III)[16]. Rh(OH)3(s) has very little affinity for the resin gels, and dissociation (solubilisation) within the diffusive gel does not take place within the time of deployment. Comparable to what we observe for Rh(III), for Cr(III) a drop in CDGT/Csol from 0.7 to 0.3 was observed in 5 days at pH 7 and from 0.3 to 0.11 after 17 days at pH 10 [16]. These authors showed that a fraction of Cr(OH)3(s) was taken up by the resin, requiring stronger elution procedures. In our case, an aqua regia digestion was applied to the resin, so that the uptake op Rh(OH)3(s) on the resin is accounted for. Comparable to

Cr(III), where in simple solutions Cr(III) will essentially be present as Cr(OH)3° at a pH 2+ - between 6.6 and 10.5, and as CrOH at pH < 6.5 and Cr(OH)4 at pH >10 [398], for Rh, in - freshwater at typical pH values of 6-8, both Rh(OH)3° and Rh(OH)4 are thermodynamically 2+ + - 2- stable. Positively [Rh(OH) , Rh(OH)2 ] and negatively, [Rh(OH)4 , Rh(OH)5 and Rh(OH)6 3-] charged species predominated under pH 4 and above pH 7, respectively [110]. Trivalent metal ions are also more prone to aquation reactions compared to divalent metals [399]. The 3- spiked solutions, in which Rh is present essentially as RhCl6 will result in the formation of aqua-hydroxy, aqua-chloro-, mixed aqua-chloro complexes in the form of 3 – x – y [RhClx(OH)y(H2O)z] , where x + y ≤ 6 and z = 6 – x – y and at equilibrium the hydroxide complexes will prevail[110]. The solubility of Rh(OH)3(s) is minimal at pH 6-7 [116], corresponding to a maximum concentration of Rh of 26 µg/L. For Cr(III), the maximum concentrations is 3 µg/L and it has been shown that the solubility can decrease by two to five orders of magnitude due to the presence of Fe, forming mixed Fe-Cr-hydroxides [398]. A similar process is expected for Rh, as it has a very high affinity for FeO(OH) phases [116]. As we do not observe a drop in the total dissolved concentrations in water (Csol), the Rh(OH)3(s) is in a colloidal form in solution. In seawater, speciation calculations predict the dominance of hydroxylated species - 2- Rh(OH)4(H2O)2 and Rh(OH)5(H2O) [110], thus as Rh(OH)3 is not a predominant species, the formation of Rh(OH)3 is less pronounced and apparent diffusion coefficients show a much lower decrease.

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4.3.3. The effects of ionic strength and DOM on uptake on DGT measurements

Pt_ Ageing effect_ D 17 days aged solutions Rh_ Ageing effect_ D 17 days aged solutions 0,50 1,40 CDGT/Csol CDGT/Csol 1,30 1,20 0,40 1,10 1,00 0,90 0,30 0,80 0,70 0,60 0,20 0,50 0,40

0,30 0,10 0,20 0,10

0,00 0,00 0.03 0.1 0.5 4 8 12 0,03 0,1 0,5 4 8 12 M (NaCl) mg/L (DOC) M (NaCl) mg/L (DOC) S920 S914 MPX 317 S920 S914 MPX 317

Figure 4. 3. Effect of ionic strength and DOC on the ratio of DGT Pt, Pd and Rh concentrations, CDGT, to their concentrations in the bulk solution, Csol Mean values and standard deviation (error bars) of triplicate measurements are given (n = 3). DDGT for each binding gel is adopted from our previous work [221] taking into account temperature correction.

Varying the ionic strength (using NaCl) of the solution over the studied range after 17 days ageing did not have a substantial effect on the concentration of Pt, or the diffusive coefficient. Using the effective diffusive coefficient of FSW of freshly spiked solution from our previous work [221], a good correspondence is observed between the CDGT calculated concentration and total dissolved Pt in solution, which means that the possible speciation changes of Pt after 17 days did not affect the DGT uptake. The effect of DOC is somewhat more pronounced than in our previous work[221], with a 25% decrease in the CDGT/Csol at a DOC concentration of 12 mg/L. This is however, comparable to what is observed for other metal complexes [400] . As soft metals, the PGEs have a high affinity for dissolved organic matter, forming mixed hydroxy-fulvate metal complexes [115, 116, 328]. This has an important influence on the mobility of the PGEs in the environment. The lower diffusion coefficient is either due to the formation of very stable complexes which will do not dissociate within the time of diffusion, or due to slower diffusion of high molecular weight complexes. The difference with our previous work, where the influence of DOC was not so pronounced, may be the result of slow formation of the stable complexes.

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For Rh, CDGT/Csol, calculated with the diffusion coefficients in 1 day aged solutions (Table 4.1) vary between 0.1 and 0.3 and show no significant influence of ionic strength. The difference observed between the resins may indicate that a small fraction of Rh(OH)3(s) is taken up by the resin gel and this is lower for MPX-317 compared to S920 and S914 as was also demonstrated in the uptake kinetics experiment (Figure 4.1). An important increase in CDGT/Csol is observed in the presence of dissolved organic matter.

The solubility of Rh(OH)3 increases drastically in the presence of fulvic acids. An increase in the fulvic acid concentrations by an order of magnitude results in the solubilisation of Rh(OH)3 by two orders of magnitude[115, 116] with the formation of Rh(III) hydroxyfulvate complexes. Organic matter can effectively bind Rh [115, 116, 328], but the kinetics of the formation are slow, requiring 15 days to attain equilibrium[110]. Our results show that Rh(OH)3(s) is partially solubilized by the humic matter and that the organo Rh complexes can be taken up by the resin gel, in contrast to Rh(OH)3, for which the uptake is very limited. Thus, DGT could be applied to determine the kinetics of both the formation of the inert Rh(OH)3(s) as its solubilization by organic matter. The results point out that long-term lab deployment experiments in synthetic or spiked solutions, especially at µg/L levels, will be problematic for Rh due to the formation of inert complexes. Re-using previously spiked solutions for subsequent lab experiments should also be avoided. The formation of Rh(OH)3 in natural waters reduced its DGT lability, but most likely also the bioavailability and mobility of Rh, thus DGT will be measuring the labile and bioavailable fraction of Rh in solution. It is well know that Rh has a much higher particle reactivity compared to Pt and Pd, resulting in its removal from solution[116]. However, also in natural waters, the mobility of Rh will increase due to complexation with dissolved organic matter [110, 116, 328]. 4.4. Conclusions

Ageing of solutions of Pt results in the formation of various mixed aqua and hydroxy complexes. For Pt, all the complexes formed are DGT labile, whereas for Rh, this will lead to the formation of Rh(OH)3(s), which is not DGT labile. Addition of dissolved organic matter results in an increase of the labile Rh fraction due to the solubilisation of Rh(OH)3(s). Thus, this ageing effect will influence long term lab experiments, especially at higher concentrations. The effective diffusion coefficients determined with aged solutions are lower than freshly spiked solutions and will thus lead to an underestimation of the diffusion coefficients in natural waters, these hydrolysis reactions will reduce DGT lability of Rh.

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Chapter 5: Elimination of Interferences in the Determination of Platinum, Palladium and Rhodium by Diffusive Gradients in Thin-Films (DGT) and Inductively Coupled Plasma Mass Spectrometry (ICP MS) using selective elution

Optimized DGT typical work for PGEs

ABDULBUR-ALFAKHOURY, E. AND M. LEERMAKERS (2021). "ELIMINATION OF INTERFERENCES IN THE DETERMINATION OF PLATINUM, PALLADIUM AND RHODIUM BY DIFFUSIVE GRADIENTS IN THIN FILMS (DGT) AND INDUCTIVELY COUPLED PLASMA MASS SPECTROMETRY (ICP MS) USING SELECTIVE ELUTION." TALANTA 223(PT 2): 121771.

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Abstract

The analysis of platinum (Pt), palladium (Pd) and rhodium (Rh) in aquatic samples by the diffusive gradients in Thin-Films (DGT) technique using chelating resins, specific designed for the accumulation of PGEs, namely Purolite S914, S920 and Italmatch Chemicals IONQUEST® MPX-317 may however, still be influenced by the accumulation of other elements such (Cu, Zn, Pb, etc.) which will be extracted simultaneously by the hot aqua regia extraction and interfere with the Inductively Coupled Plasma Mass Spectrometry (ICPMS) analysis of the Platinum Group Elements (PGEs). Selective extractions were investigated to release the interfering elements without loss of the Platinum Group Elements (PGEs) from the resin gels. A rinse with deionized water removes over 95% of Sr and Rb and a second rinse -1 with 0.05 mol L H2SO4 can be used to as a common eluent to remove an important fraction of the interfering elements from S920 and S914 without loss of PGEs but this results in loss of around 15 % of the PGEs from MPX-317. It was shown that selective extractions can be used to remove specific interferences from each resin gel.

Keywords

Platinum group elements (PGEs); Sector field inductively coupled plasma mass spectrometry (SF-ICPMS); Diffusive Gradients in Thin-Films (DGT); Interferences; Selective elution

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5.1.Introduction

Analysis of Platinum Group Elements (PGEs) in environmental samples is challenging because of the very low concentration levels of these analytes, the lack of proper reference materials and limitations of the analytical techniques used in the final determination step [142]. The inadequate detection power of the majority of instrumental analytical techniques puts severe constraints on the determination of these elements in environmental samples [143]. Separation and pre-concentration steps often precede the final detection of these metals [144]. Numerous interfering effects from the sample matrix, in particular when examining complex environmental samples of extremely low PGE concentrations, may limit direct application of particular techniques [144]. The approaches quite often used are based on Adsorptive voltammetry (AV) [145, 146], electro-thermal atomization atomic absorption spectrometry (ETA–AAS) [147], neutron activation analysis (NAA) [148], inductively coupled plasma optical emission spectrometry (ICP–OES) [149, 150] and inductively coupled plasma mass spectrometry (ICP–MS)[88, 143, 151-157], often in combination with demanding pre- concentration procedures [158-161]. Due to the very low detection limit, wide dynamic range, analysis of a large number of samples, and multi-element capability of ICP-MS, this method is considered as the most appropriate for direct determination of PGEs in various environmental samples [157]. Commercially available instrumentation fall into three groups, namely quadrupole (Q-ICPMS) based instruments, high- resolution sector field instruments (SF-ICPMS) and time-of-flight (TOF-ICPMS) systems [166] but the (Q-ICPMS and SF-ICPMS) are the most commonly used for environmental applications. The Q-ICP-MS spectrometers possess a mean mass resolution of approximately 300 m/∆m, whereas the SF-ICP-MS, working under high resolution (HR) settings, offers an enhanced resolution (10 000 m/∆m) whereas in low resolution (LR) mode, SF-ICPMS the sensitivity is about ten times higher than Q-ICP-MS[401]. For Rh, the most important interferences of concern are from ArCu+, SrO+, RbO+, ArZn+ and Pb2+; for Pt this is HfO+ and + + + + + + for Pd interferences of concern are from ArCu , YO , SrO , ZrO , MoO , RbO and Cd isotopes (Table 2.3). The extent of interfering effects directly depends on the concentration of matrix elements [346] as well as the ability of the formation of interfering species in the plasma [370] [144]. The problem of spectral interferences can be dealt with in two ways, either by their removal or by correcting for them [347, 348]; Either theoretical approaches (mathematical corrections), or instrumental approaches (nebulizer with cryogenic desolvation, matrix separation of interfering species: offline/online chromatographic systems, cold plasma

133 conditions, collision/reaction cell technologies[153] or high mass resolution spectrometers [88, 143, 155, 322, 349] can be used. The use of high-resolution sector field ICP–MS (HR-ICP–MS) allows the elimination of many polyatomic interferences by resorting to mass resolutions (m/∆m) of LR= 300, MR= 4000 or HR= 10000 modes. For example, the resolution, m/∆m, required to overcome interference of 179Hf16O+ on 195Pt is 8200, of 40Ar65Cu+ on 105Pd is 7300 and 38Ar65Cu+on 103Rh is 7200. Separations of ions with isobaric overlapping signals (106Cd+ on 106Pd+ signal) require a resolution which is not obtainable by current commercial instruments. Hence, not all interferences are resolved by the HR mode. Furthermore, In HR mode, a significant decrease in sensitivity (to 1% of that found with LR mode) is observed which, accordingly, impairs the LOD [401, 402]. Unresolved spectral interferences by the HR mode could be corrected alternatively by using mathematical correction, such as elemental equations [345, 352-354] and/or multivariate technique [347, 348]. Mathematical correction procedures, based on signal ratio measurement, have been extensively applied for the evaluation of the PGE content of environmental samples [152, 346, 353, 354]. These methods involve quantifying the interferents (atomic or molecular) signal that overlaps the analyte Pt, Pd and Rh signal and subtracting it by mathematical equations to give a corrected value; Scorr =Smeas − (Sinter A) where Scorr is the corrected signal of the analyte, Smeas the measured signal, Sinter the signal of the interfering element and A is the % of formation of the respective interfering species (e.g. oxide). The successful application of the mathematical correction is depended on the relative contribution of the interferent signal to the PGE signal [346, 354]. Mathematical correction, based on the evaluation of the contribution of interferent signal in that one for analyte, was successfully applied for elimination of spectral interferences occurring during determination of Pt by ICP-MS technique [129, 151, 155, 353], but was inefficient for the determination of Pd [129, 155, 345] due to the high contribution of the interferent signal relative to the Pd signals. Diffusive Gradients in Thin-Films (DGT) was introduced by Davison and Zhang in 1994 [267, 268] pre-concentrates selected metals over the deployment time, improving detection limits and reducing matrix effects in environmental matrices such as water. The principle of the DGT technique is based on the diffusion of the dissolved species through a diffusive layer and their accumulation on a binding resin. A hydrogel and a membrane filter are commonly used as the diffusive layer and the binding layer commonly consists of a resin incorporated into a hydrogel [267, 268]. The binding phase gel disc is placed on the base of the DGT sampler, then a diffusive gel disc is placed on it followed by a membrane filter. Finally, a plastic outer sleeve

134 with the exposure window is clamped to the base of the DGT sampler. The time-average concentration of metal in the solution, CDGT, can be calculated with the help of Fick's first law of diffusion equation 1.7. A membrane filter protects both gels from particulate matter and holds all underlying layers packed. Its main role is to protect the DGT device from particulate matter and other potential physical harm. The diffusive gel role is to lessen the impact of variations in water turbulence and to regulate and slow down the diffusion or supply of ions to the resin gel layer, by creating a diffusive gradient [259]. The diffusion limiting layer can also have other functions, such as to exclude analyte species that are too large to pass through the pores, and to reduce the sampler sensitivity to variations in turbulence [259], and the role of the sorbent/resin is the fast and irreversible binding of the solutes of interest. In the ideal case, interfering compounds are not bound on the resin/sorbent gel. The most commonly used resin gel incorporates Chelex®-100 resin with functional groups of iminodiacetate acid (IDA), which act as chelators for polyvalent metal ions [267, 268], but many other resins and sorbents have been used such as Diphonix for

U [284, 285], MnO2 for Ra [286], 3-mercaptopropyl functionalized silica gel for Hg [287], Metsorb, ferrihydrite or ZrO for elements such as As, P, Mo, V [288]. For the PGEs, the DGT technique has been recently developed[221], where several resins with high selectivity towards PGEs were evaluated. These include ion-exchange resins with polyamine [112-114, 205, 240, 241, 289], thiol [186, 217], Thiourea [186, 201, 215], Isothioureum [186, 220], phosphine oxide [189, 201, 232] functional groups. These functional groups have a high selectivity for Pt, Pd and Rh but they show accumulation for undesired elements such Cu, Zn, Y, ……etc. [221], whose presence in the elution will cause interferences during the ICPMS measurements. The potential interferences during the previous study [221] were resolved using HR mode together with mathematical corrections which deteriorated the detection limit of the method by the HR mode and increasing the uncertainty by mathematical corrections. The binding of metal ions on the functional groups of ion-exchange and chelating resins will depend on the stability constant of the metal chelates, the acid/base properties of the ligands and metal ions, the solution pH, relative position of the functional groups, spatial configuration and steric effects [403]. The ligands of the functional groups in resins behave as bases (i.e., electron donors) and the metal ions as acids (i.e., electron acceptors) [170]. . Precious metal ions of low oxidation state (e.g. Au(I), Pd(II) and Pt(II)) belong to “soft acids”[183, 404] while Cu, Zn and Pb are borderline elements, whose behaviour depends on their oxidation state and considered as soft acids when they have low or zero oxidation states[404]. Phosphorus, sulphur and nitrogen donor ligands belongs to the soft bases, while oxygen is a hard base [404, 405].

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The ion exchangers of the functional groups containing one or more soft-bases donor atoms interact strongly with soft acids with the order P > S > N > O. Hard metal ions such as Sr(II), Rb(I), Zr(IV), Mo(IV), Y(III) are ions with low electronegativity, non-polarizable, have high charge to ionic radius ratio [228] and they show affinity for hard bases with the donor atoms O > N > S >P [185, 186]. Hence, resins containing functional groups with soft base characteristics are used for the adsorption of metal ions with soft acid characteristics. For example, ion-exchange resins containing resins based on thiourea and its conjugate thiouronium are classified as soft Lewis bases and strong cationic resins with chelating properties. In the present study, we investigated, based on the hard-soft acid properties of PGEs and interfering ions, which selective elution procedures could be used to separate the PGEs and interfering ions from the resin gels, thus improving the detection limits of the method.

5.2.Materials and methods

5.2.1. General Procedures

All chemicals were of analytical reagent grade or greater. High purity nitric acid (HNO3) and hydrochloric acid (HCl) were used in all experiments (Fisher Scientific, Trace metal grade). Standard solutions used were Rhodium ICP Standard 1000 mg L-1 in 6% HCl (SIGMA -1 -1 ALDRICH), Palladium 1000 mg L in 2% HNO3 and Platinum 1000 mg L in 2% HCl ICP- MS ULTRA grade(TM) Standard (Ultra Scientific, North Kingstown, England). All further dilutions of the stock solutions were prepared in 2% HCl. Deionized water (Milli-Q Advantage with Element Pod, Merck Millipore, USA), named MQ-water hereafter, was used for the preparation of the solutions, gels and cleaning glassware and containers. Solutions of eluents were prepared by dilution/dissolving appropriate amount with MQ-water or H2SO4 (Merck, pa), Na2SO4 (Merck, pa), NH4SCN (Merck, pa) , NaCl (Merck, pa), CH3OH (Merck, pa),

NaNO2 (Merck, pa), KOH (Merck, pa), CH3COONa (Merck, pa), CH4N2S (Merck, pa ), -1 CH3COOH (Merck, pa), C2H5OC2H5 (Merck, pa), Cadmium standard solution 1 g·L (AVS Titrinorm, VWR) and lead standard solution 1 g·L-1 (SpectrosoL, BDH, VWR) were used to prepare the model solutions. 5.2.2. DGT preparation and assembly

5.2.2.1.Diffusive gel preparation Details were previously discussed in section 2.3.1.

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5.2.2.2.Binding gels preparation investigation Details were previously discussed in section 2.3.2. 5.2.2.3.DGT assembly Details were previously discussed in section 2.3.3. 5.2.2.4.Preparation of the deployment solutions Preparation of PGE spiked solutions is prepared as previously described in [221], where glass containers employed which have been equilibrated with PGE spiked solutions for at least 3 days to avoid adsorption on the container walls. This solution was then discarded, and a freshly prepared test solution was placed in the containers. Deployment solutions consisted of spiked filtered river water (Zenne River, Belgium). 5.2.3. Metal uptake on resin gels:

For the performance of DGT test for simultaneous accumulation of Pt, Pd, Rh, Zn, Cu, Sr, Rb, Hf, Y, Pb, Cd and Mo was evaluated using concentrations of the interferences were chosen to be comparable to concentrations found in natural river water samples. DGT devices with 10 binding gels of each resin S914, S920 and MPX-317 were deployed in 10 liter filtrated river water and spiked with Pt, Pd, Rh, Zn, Cu, Hf, Y, Pb, Cd and Mo at 10 µg L-1 level while Sr and Rb were not added due to their high concentration in riverine water (Sr 280 µg L-1, Rb 76 µg L-1[221]). The ratio of the concentrations of interfering elements to the concentrations of PGEs in solution thus range from 2 to 30 depending on the concentrations of the elements in the unspiked natural water. The test solution was left for 24h for equilibration before the deployment. Master variables (T = 21 ⁰C, pH = 8.6, salinity = 0.4, total dissolved solids (TDS)= 700 mg L-1, DOC = 4.3 mg L -1) were measured before the deployment and at the end of the test. The DGT probes (n=2 at each sampling-time) were exposed for 6, 12, 24, 48, and 72 hours in the stirred-solutions. At each sampling time, the DGT assemblies were rinsed with MQ water immediately after retrieval. The pistons were disassembled, and the resin-gels transferred into a Teflon bottle and eluted in 1 ml aqua regia in an oven at 70°C for 24 h as described in [221]. After cooling, 8.95 mL of MQ-water was added. For resin gels that did not dissolve completely in aqua regia the solution was filtered using 0.45µm disposal mixed cellulose ester syringe -1 filters (Chromafil) pre-washed with 10 mL of 1.4 mol L HNO3. 5.2.4. Removal the interferences from the resins gels

In order to test the selective separation of the accumulated interferences from the resin gels prior the elution with hot aqua regia, 26 DGT units of each resin S914, S920 and MPX-317 were deployed in 7 litre filtrated river water spiked with Pt, Pd, Rh, Zn, Cu, Hf, Y, Pb, Cd and

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Mo at 10µgL-1 level (Sr and Rb were not added due to their high concentration in riverine water) and continuously stirred. The DGT probes were exposed for 24 hours to the stirred solutions. After retrieval, the pistons were disassembled, and the resin gels transferred into Teflon bottles. The first step, all the resin gels were subjected to a wash with MQ water to remove the Sr and Rb as described in our previously published work [221]. In the next step, the resin gels were divided into 13 groups (two pistons of each group). The first group called the reference and this group is eluted with hot aqua regia. The reference will have no loss of bounded Pt, Pd and Rh and the other elements and it will be used as the comparison for the other groups. The other groups were subjected to different eluent solutions 1) 0.5 mol L-1 -1 -1 -1 - Na2SO4, 2) 0.5 mol L NaNO2, 3) 1 mol L NaCl, 4) 0.5 mol L CH3COONa, 5) 0.1 mol L 1 -1 -1 -1 CH3COOH, 6) 0.01 mol L HNO3, 7) 0.05 mol L H2SO4, 8) 0.1 mol L KOH, 9) 50% -1 -1 CH3OH, 10) 0.5 mol L NH4SCN, 11) 0.5 mol L CH4N2S, and 12) 50% C2H5OC2H5. After the second elution step, a third wash step of MQ-water of 12h was conducted to remove the last remaining eluent. Finally, all resin gels were removed from the MQ-water and eluted using the hot aqua regia procedure. 5.2.5. Sample analysis

Sample analysis was performed using Inductively Coupled Plasma Sector Field Mass Spectrometry (ICP-SF-MS, Element II, Thermo Fisher Scientific Bremen GmbH, Germany), equipped with ESI fast auto-sampler as described previously [45] and instrumental parameters are shown in Table 2.1. 195Pt, 105Pd and 103Rh intensities are normalized to 193Ir, 209Bi and 115In, -1 115 respectively as internal standards and 0.42 mol L HNO3 as carrier solution. In was used as internal standard for the other metals. Calibration standards were prepared in a 10% aqua regia extracted blank resin gel matrix to obtain matrix-matched standards to construct the calibration functions. The PGEs elution factors (fe) of (Rh, Pd and Pt) for S914 (0.99, 0.935 and 0.910); S920 (0.954, 0.916, and 0.811) and MPX-317 (0.899, 0.866 and 0.942) were used from previously work [221], whereas the elution factor for the other elements was considered to be 1 in this study. 5.3. Results and Discussion

5.3.1. Uptake of metals on resin gels The accumulation of Sr, Rb, Mo, Y, Zn, Hf, Cu, Pb, Cd, Zr as well as Pt, Pd and Rh on the resin gels in function of time is shown in Figure 5.1. Three types of accumulation patterns can be observed: 1) no increase in function of time as observed for Sr and Rb; 2) an initial linear accumulation followed by a rapid flattening of the curve and no further increase in the

138 concentration in function of the time as observed for Mo and Y and 3) a linear accumulation as observed for Zn, Hf, Cu, Pb as well as Pt, Pd and Rh. The pattern of Sr and Rb is typical of an earth alkaline and alkaline metal with little affinity towards the resin. Hence, the uptake in the resin gel is determined by diffusion: a rapid equilibrium between the concentrations in the diffusive and resin gel and bulk solution is obtained and no further increase in function of time is observed. If these metals are bound to the functional groups of the resin, they will rapidly be displaced by the metals with a higher affinity towards the resin. The accumulation pattern of Mo and Y is typical for unselective resins towards a metal [269]. Competition effects [406] arise when the amount of one metal bound to the resin is dependent, for a fixed bulk metal concentration of the analyte, on the concentration of other cations present in the system [406], and starts when less resin sites are available with the result that the flux starts to decrease; the slope of the accumulation curve bends and the accumulation reaches a plateau, indicating negligible net metal binding. Accordingly, there is no metal concentration gradient i.e., the free metal concentration in the resin domain and bulk solution is in equilibrium. Hence, the Mo and Y have very low binding affinity to the selected resins in this study and they bind much less strong comparing to Rh, Pd and Pt. The accumulation pattern is not identical for all resins: a higher accumulation of Mo (about 4 times higher) is observed on S920 compared S914 and MPX-317 whereas accumulation of Y on S920 and S914 is about two times higher than on MPX-317. A general linear accumulation is observed for Cu, Zn, Hf, Pb, Cd besides the analytes Rh, Pd and Pt. Thus, these elements will not rapidly be displaced from the resin within the deployment time. Cu and Zn show a behaviour similar for all resins with a somewhat highest accumulation of Zn after 72h for S920 (factor 1.2) compared to S914 and MPX-317. Hf shows similar behaviour with the highest accumulation on S920 and lower accumulation on S914 (2x less) and MPX-317 (3x less). On the other hand, for Cd and Pb the highest accumulation is observed on resin S914 compared to S920 and MPX-317 (factor 1.5-2 for Pb and factor 6 for Cd). The accumulation of Zr is comparable for 290 and S914 and much higher (factor 70) than on MPX- 317.

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Rh Pd 500 Mass 800 Mass 400 (ng) (ng) 600 300 400 200 200 100 time (h) time (h) 0 0 0 20 40 60 80 0 20 40 60 80 S914 S920 MPX-317 S914 S920 MPX-317

Pt Cu 400 Mass 300 Mass (ng) (ng) 300 200 200 100 100 time (h) time (h) 0 0 0 20 40 60 80 0 20 40 60 80 S914 S920 MPX-317 S914 S920 MPX 317

Zn Pb 1500 60 Mass Mass (ng) (ng) 1000 40

500 20

time (h) time (h) 0 0 0 20 40 60 80 0 20 40 60 80 S914 S920 MPX-317 S914 S920 MPX-317

Hf Zr 15 2,5 Mass Mass (ng) 2 (ng) 10 1,5

1 5 0,5 time (h) time (h) 0 0 0 20 40 60 80 0 20 40 60 80 S914 S920 MPX-317 S914 S920 MPX-317

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Y Mo 250 Mass 80 Mass 200 (ng) (ng) 60 150 40 100 20 50 time (h) time (h) 0 0 0 20 40 60 80 0 20 40 60 80 S914 S920 MPX-317 S914 S920 MPX-317

Sr Rb 150 Mass 40 Mass (ng) (ng) 30 100 20 50 10 time (h) time (h) 0 0 0 20 40 60 80 0 20 40 60 80 S914 S920 MPX 317 S914 S920 MPX-317 Figure 5. 1. The uptake (ng) of Pt, Pd, Rh, Zn, Zr, Cu, Mo, Cd, Sr, Rb, Y, Hf and Pb over time using DGT units with diffusive layers (agarose diffusive gel (0.75 mm) and filter (0.17 mm)) overlaid on binding resin gel of AGA with S914, S920 and MPX-317 and deployed in spiked filtrated river water.

5.3.2. Removal the interferences from the resin gels Platinum group elements including Pt, Pd and Rh can be retained on chelating resins with P, N or S-containing chelating groups such as phosphine oxide thiourea, isothiouronium and thiourea [170, 221]. Due to their high affinity for these resins, it is often difficult to elute these elements from the resin [111, 182, 221], Hence, high concentrations or large volumes of eluents are required to recover the adsorbed metals [114, 170], or digestion of the resins using hot aqua regia [221] or fuming (ignition) [186, 205, 215] in necessary to achieve total recovery. ) Ammonium thiocyanate (NH4SCN), Sulfuric acid (H2SO4 , Thiourea (CH4N2S), Sodium hydroxide (NaOH), Ammonium hydroxide (NH4OH) and many others have been applied for elution of sorbed metals on resins[112-114, 205, 220]. Complete leaching of PGE metal ions sorbed on the resins was not a success using any of the above-mentioned elution agents separately. However, their combination might allow carrying out the separation [186, 205, 236]. These eluents were now evaluated for possible elution of the interfering elements which are less strongly bonded to the resins. The concentration of these eluents was kept low to

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prevent the elution of the PGE from the resin gel but high enough to remove the undesired elements. 5.3.2.1. Purolite S920 Figure 5.2 shows the accumulated mass of the elements after extractions with the different eluents for S920. As shown in our previous work [45] the reference wash with MQ-water leaches over 98% of the Sr and Rb on the resins without loss of the PGEs. In comparison with the reference, Rh leaches using a wash solution of KOH, while Pd and Pt leach occur using

(C2H5)2O. Hence, both eluents (8 and 12) cannot be used for the elution of interferences. Cd,

Pb and Zn can be eluted with the acids H2SO4, HNO3 and CH3COOH and for Zn also by

CH3OH. Y and Hf can be removed by the salts Na2SO4, NH4SCN, NaCl, CH3COONa as well

as by the acids HNO3, H2SO4 and CH3COOH. Mo can be removed by the salts Na2SO4,

NH4SCN, NaCl, CH3COONa but only by H2SO4 as acid and Cu can be removed by H2SO4 and

CH4N2S. The residual Sr and Rb can partially be removed by the acids H2SO4, HNO3 and

CH3COOH as well as NH4SCN and CH3OH. Zr can partially be removed by H2SO4 as well as -1 with KOH. Of all the washes tested on the resin gel S920, the wash with 0.05 mol L H2SO4 results in an elution of almost undesired elements, without elution of the PGEs. Over 50% of Cu, Mo and Zr can be removed, 70% of the accumulated Hf and over 95% of Zn, Cd, Pb and Y.

50 Mass Pt Rh 150 Mass Pd 30 Mass (ng) (ng) 40 (ng) 20 30 100

20 50 10 10

0 0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12

4 Sr Rb Mass 1,0 Mass 2000 Mass Cu (ng) (ng) (ng) 0,8 3 1500 0,6 2 1000 0,4 1 500 0,2

0 0,0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12

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Mass Zn 15 Y Hf 3000 (ng) Mass 1,2 Mass (ng) 1,0 (ng) 2000 10 0,8 0,6 1000 5 0,4 0,2 0 0 0,0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12

Mass Cd 40 1000 Mass Mo 4 Zr (ng) Mass (ng) 800 (ng) 30 3 600 20 2 400 10 1 200

0 0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12

Mass 300 Pb (ng) 250

200

150

100

50

0 ref. 1 2 3 4 5 6 7 8 9 10 11 12

Figure 5. 2. Accumulated mass of metals (ng) on S920 after different elution procedures: ref. is the -1 -1 reference method with only a deionized water rinse; 1) 0.5 mol L Na2SO4, 2) 0.5 mol L NaNO2, 3) -1 -1 -1 -1 1 mol L NaCl, 4) 0.5 mol L CH3COONa, 5) 0.1 mol L CH3COOH, 6) 0.01 mol L HNO3, 7) -1 -1 -1 -1 0.05 mol L H2SO4, 8) 0.1 mol L KOH, 9) 50% CH3OH, 10) 0.5 mol L NH4SCN, 11) 0.5 mol L CH4N2S, and 12) 50% C2H5OC2H5. A single 12h elution was performed in each treatment.

Purolite S920 is a weak base resin with isothiouronium functional groups with a very high selectivity towards Hg, Au and PGEs. Depending on the pH of the medium, isothioureum groups can occur as free base and acid conjugated form which tend to the acid–base equilibrium [186]. R NH R + NH 2 H /X - S S C + +X (2) NH 2 NH2 acid conjugated form Free base form

143

Adsorption of metals on the resin can occur both through ion-exchange mechanisms as − chelation and chelation (coordination) mechanisms as shown for [푅ℎ(푂퐻)4(퐻2푂)2]푎푞 and 2− [Pd퐶푙4] below. Anion-exchange:

R R NH2 NH2 - - -x S C + [Rh(OH) (H O) ]- S C + X + [Rh(OH)4(H2O)2] 4 2 2 (3)

NH2 NH2

(4)

Chelation: + − − − + (푅1푅2)푁퐻2 퐶푙푠표푟푏+[푅ℎ(푂퐻)4(퐻2푂)2]푎푞 ---> (R1R2)NH →[푅ℎ(푂퐻)4(퐻2푂)2]푠표푟푏+ 퐻푎푞+ (5) − 퐶푙푎푞 + − 2− − + − (푅1푅2)N퐻2 퐶푙푠표푟푏 + [Pd퐶푙4]푎푞 → (푅1푅2)푁퐻 → [푃푑퐶푙3 ]푠표푟푏 + 퐻푎푞 + 2퐶푙푎푞 [170, 186, 205] (6)

Complexes can be formed using S and both N atoms as ligands.

(7)

Ion exchange mechanism is predominant at high acidities while those chelating or coordinating at low acidities [215]. The resin is unstable in alkaline pH and loses the thioureum functional group forming methylthiol and urea although it is not clear whether this also occurs when metals are complexed to the resin. The loss of Rh in alkaline medium can be explained by the reaction:

R NH R NH2 S S C + - [Rh(OH)4(H2O)2] + 3KOH K3[Rh(OH)6] + (8)

NH2 NH2

The hard acids Y, Hf show little affinity for the resin and can easily be removed by increasing

ionic strength with salts such as Na2SO4, NaCl …etc, or by decreasing the pH, resulting in a protonation of the amines and electrostatic repulsion between the protonated amines and the positively charged metals. Sr and Rb are already efficiently removed from the resin by the deionized water rinse, reducing the amount in the resin gel from around 80 to 2 ng as has

144 already been reported [45] and can further be removed by the elution with salts or acids as these are hard acids with low affinity for the resin. Intermediate metals such as Pb and Zn as well as the soft acid Cd, which form metal chelates can be removed by a decrease in pH with acids such as H2SO4, HNO3 and CH3COOH but for Cu, which forms more stable complexes with S ligands only H2SO4 as well as thiourea removes Cu from the resin.

5.2.2.1. Purolite S914 Figure 5.3 shows the accumulated mass of elements on the S914 resin gel after use of the different eluents. It can be seen that several eluents leach the PGEs from solution. Loss of Rh,

Pd and Pt is observed using the salts NaNO2, NaCl and CH3COONa as well as with KOH and in addition loss of Rh is observed using Na2SO4 and loss of Pd and Pt using NH4SCN. Cd can efficiently be removed by the acids H2SO4, CH3COOH and HNO3. The same acids efficiently remove Zn; which can also be eluted with KOH. The same eluents can be used for Pb, but leaching is less efficient. Y can be leached by the acids H2SO4 and HNO3, whereas Zr and Hf can partially be removed only by H2SO4. Mo can be leached by the salts NaNO2, NaCl, Na2SO4,

NH4SCN and CH3COONa as well as in an alkaline leach KOH. Only H2SO4 as acid can leach

Mo from the resin gel. Cu is the most difficult to elute. Only H2SO4 results in a 20% leach of

Cu. The residual Sr and Rb can efficiently be removed by the acids H2SO4, CH3COOH, HNO3 and to a lesser extent by the salts NaNO2, NaCl, Na2SO4, NH4SCN and CH3COONa as well as -1 the solvents CH3OCH3 and CH3OH, Hence, for the resin’ gel S914, the wash of 0.05 mol L

H2SO4 is the common eluent for the partial removal of of the interfering elements. The 0.05 -1 mol L H2SO4 wash step is capable removing over 95% of Sr, Rb, Zn, Y and Mo and a 20- 30% of Cu, Cd, Pb and Zr and Hf.

145

Mass Mass Mass 100 (ng) Pt 200 (ng) Pd 30 (ng) Rh

80 25 150 20 60 100 15 40 10 50 20 5 0 0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 Mass Mass Mass 10 (ng) Sr 3 (ng) Rb 20 (ng) Cu

8 3 15 2 6 2 10 4 1 5 2 1 0 0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 Mass Mass Mass (ng) Zn 4 Y Hf (ng) 30 (ng) 400 3 300 20 2 200 10 100 1

0 0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 Mass Mass Mass 8 (ng) Cd 30 (ng) Mo 2,0 (ng) Zr 25 6 1,5 20 4 15 1,0 10 2 0,5 5 0 0 0,0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 Mass 600 (ng) Pb 500 400 300 200 100 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12

Figure 5. 3. Accumulated mass of metals on S914 after different elution procedures: ref. is the -1 -1 reference method with only a deionized water rinse; 1) 0.5 mol L Na2SO4, 2) 0.5 mol L -1 -1 -1 NaNO2, 3) 1 mol L NaCl, 4) 0.5 mol L CH3COONa, 5) 0.1 mol L CH3COOH, 6) 0.01 -1 -1 -1 -1 mol L HNO3, 7) 0.05 mol L H2SO4, 8) 0.1 mol L KOH, 9) 50% CH3OH, 10) 0.5 mol L

146

-1 NH4SCN, 11) 0.5 mol L CH4N2S, and 12) 50% C2H5OC2H5. A single 12h elution was performed in each treatment.

Purolite S914 contains thiourea functional groups. Thiourea groups exhibit selectivity order Hg2+ > Ag+ > Au+ = Au3+ > Pt2+ = Pt4+ > Cu2+ > Pb2+ = Pb4+ > Bi2+ > Sn2+ = Zn2+ > Cd2+ > Ni2+[228].

SH S

R N NH R N NH2 (9)

H H Thione form Thiol form

This resin can also be protonated in acidic conditions resulting in both anion exchange and chelation binding mechanisms. The two N atoms a long with S of the thione form coordination complexes with metals ions (Figure 5.5).

Figure 5. 4. Coordination complex of the thione form with metals ion

The chelation mechanisms of PGE complexes was proposed by Hubicki [186] − − (푅1푅2)C=S + [PdCl4] → (푅1푅2)퐶 = 푆 → [푃푑퐶푙3 ]푠표푟푏 + 퐶푙푎푞 (10) Possible binding modes for soft and intermediate acids have been demonstrated, confirming the involvement of C=S in coordination with metals in the resin’s backbone as shown for Cd below (Figure 5.6) [201, 234].

Figure 5. 5. Possible binding modes for Cd with thiourea functional groups.

For Cu, a similar binding mechanism can be expected and the binding strength of the Cu complex is much stronger than that of Cd and Zn, which is most likely due to the specific sulfur-containing functionality of this material [182, 201], making it more difficult to

147

selectively remove Cu from the resin. Higher concentrations of thiourea and H2SO4 would be required which would also possibly leach a fraction of the PGEs [186].

5.2.2.1. Italmatch MPX-317 Figure 5.4 shows the accumulated mass in function of the eluent using binding gels of MPX-

317. In comparison with the reference, the HNO3 elution influences the recovery of Pt, Rh and

Pd (10-30% loss) whereas thiourea and H2SO4 also decreases the recovery the PGEs by 15%.

Only the washes with CH3OH, NH4SCN, Na2SO4, NaNO2, (CH3)2O and CH3COONa do not affect the recovery of PGEs. These eluents also remove the residual Sr and Rb from the resin.

Cd, Pb, Y and Zn can efficiently be removed by acid eluents (H2SO4, HNO3, CH3COOH), but

also by CH3OH. Mo can be removed by the salts NH4SCN, Na2SO4, NaNO2, NaCl,

CH3COONa as well as with an alkaline leach KOH. Hf can partially be removed by the salts

Na2SO4, NaNO2, NaCl and acids (H2SO4, HNO3, CH3COOH) as well as KOH. Cu cannot be -1 removed by any of the different eluents. Thus, a 0.5mol L NaNO2 rinse efficiently removes the residual Sr and Rb (85%), Zr and Mo (>99%) and partially Hf, Y, Cd, Pb and Zn (20-50%) without eluting the PGEs.

Mass Mass Mass 50 (ng) Pt 200 (ng) Pd 15 (ng) Rh

40 150 10 30 100 20 5 50 10

0 0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 Mass Mass Mass 5 (ng) Sr 1,2 (ng) Rb 500 (ng) Cu

4 1,0 400 0,8 3 300 0,6 2 200 0,4 1 0,2 100 0 0,0 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 Mass Mass Mass Zn 20 Y 0,2 Hf 200 (ng) (ng) (ng)

150 15 0,1

100 10 0,1 50 5

0 0 0,0 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12

148

Mass Mass Mass 25 (ng) Cd 400 (ng) Mo 0,04 (ng) Zr

20 300 0,03 15 200 0,02 10 100 0,01 5

0 0 0,00 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 ref. 1 2 3 4 5 6 7 8 9 10 11 12 Mass 1200 (ng) Pb 1000 800 600 400 200 0 ref. 1 2 3 4 5 6 7 8 9 10 11 12

Figure 5. 6. Accumulated mass of metals (ng) on MPX-317 after different elution procedures: -1 ref. is the reference method with only a deionized water rinse. 1) 0.5 mol L Na2SO4, 2) 0.5 -1 -1 -1 -1 mol L NaNO2, 3) 1 mol L NaCl, 4) 0.5 mol L CH3COONa, 5) 0.1 mol L CH3COOH, -1 -1 -1 6) 0.01 mol L HNO3, 7) 0.05 mol L H2SO4, 8) 0.1 mol L KOH, 9) 50% CH3OH, 10) 0.5 -1 -1 mol L NH4SCN, 11) 0.5 mol L CH4N2S, and 12) 50% C2H5OC2H5 .A single 12h extraction was performed in each treatment.

MPX-317 contains phosphine oxide and thiourea functional groups and has a high selectivity towards PGEs. Phosphines and its derivatives are excellent ligands in coordination chemistry and are used for solvent extraction of PGEs [230, 407]. Due to the structure of the resin, a bidentate complex can be formed between the P=O groups in the resin and metal ions or metal ion complexes and this binding is a high selectivity for PGEs [189, 201]. Metal binding can occur by chelation with the P=O groups, the C=S groups as well as amine groups [189, 201, 234]. Adsorption of the surface of the resin is the dominating mechanism with possible binding modes illustrated for Cd below (Figure 5.7) [80].

Figure 5. 7. MPX-317 resin and Cd2+ Possible bonding modes of cadmium to the chelating resins [234].

149

Although coordination has been shown to be the main binding mechanism, electrostatic interactions have also been shown to play a role, also resulting in a poorer uptake of Rh compared to Pt and Pd. Although the binding of PGEs has been shown to be very efficient especially for Pt and Pd at low pH and high chloride concentrations, we observed a 5-20% loss of PGEs using a HNO3 or H2SO4 rinse, possibly due to changes in electrostatic interactions and changes in speciation of the PGE species. Thiourea has been shown to partially release PGEs from the resin [63] and this release increases with acidification. No common rinse removes all interferences without loss of PGEs and Cu cannot be removed by the rinses due to the strong interaction with the S groups. Compared to S920 and S914, the accumulation of Hf is much lower on MPX-317 (factor 3-5), which make this resin particularly interesting for Pt, as HfO is the most important interference for Pt. Washing with a salt such as NaNO2 partially removes the remaining Hf from the resin, without loss of PGEs.

5.4. Conclusion

Sr and Rb are major interferences in the analysis of PGEs by ICPMS. These elements can largely be removed from the resin gels where they are predominantly present in the gel solution by rinsing the gels with deionized water. Removal of these elements from the resins can further be performed by addition of salts or acids. The elution of PGEs from the resin gels S920 and -1 - S914 using 0.05 mol L H2SO4 is negligible, but this is not the case for MPX-317. 0.05 mol L 1 H2SO4 can be used as a common eluent for the removal of interferences on resin gel S920. Over 50% of Cu, Mo and Zr can be removed, 70% of the accumulated Hf and over 95% of Zn, -1 Cd, Pb and Y. For the resin gel S914, the 0.05 mol L H2SO4 wash step is capable removing over 95% of Sr, Rb, Zn, Y and Mo and a 20-30% of Cu, Cd, Pb and Zr and Hf. Although the -1 0.05 mol L H2SO4 can efficiently remove several interferences from the MPX-317 resin, it also results in a 15% loss of PGEs. Cu also cannot efficiently be removed by any of the rinses -1 tested on the MPX-317 resin gel. A 0.5mol L NaNO2 rinse efficiently removes the residual Sr and Rb (85%), Zr and Mo (>99%) and partially Hf, Y, Cd, Pb and Zn (20-50%) without eluting the PGEs. The MPX 317 resin accumulates less Hf, Y and Zr compared to S920 and S914. The MPX-317 resin is particularly interesting for the determination of Pt when high Hf concentrations are expected as this resin accumulates 3-5 times less Hf compared to S920 and

S914 and the residual Hf can be removed by a rinse with a salt solution such as NaNO2. The use of optimized rinsing procedures prior to the extraction of PGEs results in an improved reliability of DGT measurements of PGEs, especially in polluted aquatic systems with high concentrations of interfering elements.

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Chapter 6: Developing SPE to measure PGEs from surface water

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6.1. Introduction Low concentrations of Pt, Pd and Rh in environmental samples (ng or pg g-1 level) with the complexity of the matrix causing many interferences during the determination, which may require application pre-concentration/matrix separation procedures before ICPMS detection of the analytes. Among different separation techniques, solid phase extraction (SPE) based on adsorption [149], ion-exchange or chelating properties of solid sorbents have been demonstrated to be most effective [370]. Ion exchange technique is very suitable for eliminating spectral interference in the determinations of PGE by ICP-MS. Ion exchange resins have been classified based on the charge on the exchangeable counter ion (cation or anion exchanger) and the ionic strength of the bound ion (strong or weak exchanger)[174]. Ion-exchange methods are based on platinum group elements form stable anionic chlorine complexes (preferably at 0.5M HCl [373] or pH = 0.3 for natural water samples), while the majority of transitional group or rare-earth elements form weaker anionic or cationic complexes. The high affinity of PGE chlorine complexes for strongly basic anion-exchange resins as well as their weak affinity for cation-exchange resins can be used for separating these metals from sample matrix. Cation-exchange methods have been developed in which the matrix material of the complex sample is adsorbed on the resin and the PGEs are eluted requiring much lower acid concentrations, thus, rendering the procedure safer and less expensive [154, 373] but it does not provide any pre-concentration. Non-boiling evaporation method to increase the PGEs sample concentration have been used in [373, 408, 409]. Despite the simplicity of the method but it requires ultra-clean environment. Chelating resin is broadly classified as one type of ion exchange resins[182]. The partition coefficients of many metal ions are much higher on complexing resins than on the ion exchange ones so that a much lower free metal ion concentration can be determined using complexing resins. This allows to evaluate free metal concentration at extremely low levels in natural waters[200]. The principal advantage of chelating ligands is their high selectivity for certain precious metal ions or a group of ions, whereas, their main drawback lies in the difficulty of recovering the retained metals[183]. Various chelating groups contain N, O, and/or S donor atoms, which can coordinate with elements based on the Hard and Soft Acid and Base (HSAB) Theory [185, 186, 194], soft metal ions, for instance Au, Pd, Pt ions, show affinity for soft bases with donor atoms: S>N>O [182, 184]. These elements can also be recovered using chelating resins immobilizing S-containing chelating groups such as Thiourea, whereas it is often difficult to elute because of its strong affinity [182]. Generally, for elution, stronger

152 stripping reagents, e.g. acidic solutions of Thiourea or the application of the hot aqua regia [221, 410, 411]. In this context, we evaluate the use of cation exchange resin (Dowex 50W-X8) combined with non-boiling evaporation pre-concentration method and the use of chelating resins (S914, S920, S985, MPX-317 and MPX-310) for PGEs determination in filtrated river water.

6.2.Experimental 6.2.1. Reagents

All chemicals were of analytical reagent grade or greater. Concentrated nitric acid (HNO3) (69%) and hydrochloric acid (HCl) (37%) were used in all experiments (Fisher Scientific, “Suprapure” grade). Standard solutions used were Rhodium ICP Standard 1000 mg L-1 in -1 6% HCl (SIGMA ALDRICH), Palladium 1000 mg L in 2% HNO3 and Platinum 1000 mg L-1 in 2% HCl ICP-MS ULTRA grade(TM) Standard (Ultra Scientific, North Kingstown, England). All further dilutions of the stock solutions were prepared in 2% HCl. Deionized water (Milli-Q Advantage with Element Pod, Merck Millipore, USA), named Milli-Q hereafter, was used for the preparation of the solutions, gels and cleaning glassware and containers. Cadmium standard solution 1 g L-1 (AVS Titrinorm, VWR) and lead standard solution 1 g·L-1 (SpectrosoL, BDH, VWR) were used to prepare the model solutions. A single element standard (Alfa Aesar Spectrapure) 1000 mg L-1 was used for calibration standards of Mo. All experiments were carried out under laminar flow hood (class-100) in a clean room. All

equipment used was acid washed in 5% (v/v) distilled HNO3 and rinsed with Milli-Q water and then dried in a laminar flow hood (class-100) in a clean room before storage in cleaned polythene bags. All handling of equipment and samples was with polythene gloves. Temperature and pH measurements were performed using pH probe (WTW GmbH, Germany) and monitored during the experiment. Cation-exchange resin Dowex 50W-X8 200-400 mesh (Bio-Rad Laboratories, Richmond, CA) in H+ form was used for separation of matrix elements from the samples of filtrated river water before determination of Rh, Pd and Pt by ICPMS. Selected resins with different characteristics (Purolite S985, S914, S920 with polyamine, thiourea and isothioureum functional groups IONQUEST® MPX-317 and MPX-310 with phosphine oxide thiourea and phosphine oxide functional groups) for their applicability as binding resin for SPE applications for Pt, Pd and Rh in natural waters table 2.4.

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6.2.2. Quantification of analytes Sample analysis was performed using Inductively Coupled Plasma Sector Field Mass Spectrometry (SF-ICP-MS, Element II, Thermo Fisher Scientific Bremen GmbH, Germany), as described in Chapter 2. All standard solutions were prepared using matrix match calibration procedure either in 0.25 mol L−1 HCl (for samples after cation-exchange separation and samples of batch process) or in 0.1 M Thiourea in 0.2 M HCl (for samples after anion-exchange elution), or in ten times diluted aqua regia matrix. Interferents were measured using isotopes 89 90,92 95,98 111,114 206,207,208 and the corresponding resolution: YLR, ZrLR, MoLR, CdLR, PbLR, 63,65 66,67,68 85,87 87,88 178,179 CuMR, ZnMR, RbMR, SrMR, and HFLR. To compensate for instrumental drift and non-spectral matrix effects internal standardization was applied using the ICP index for ICP-MS determinations [332], where Rh, Pd, and Pt intensities are normalized to (115In, 209 193 -1 Bi and Ir), respectively, at 2 µg L of each in 0.42M HNO3 and 0.42M HNO3 as carrier solution for tests with aqua regia or only HCl matrix involved but using 0.42M HCl and 0.42M HCl as carrier solution for Thiourea and HCl elunets. During the research, I was using five internal standards (115In, 139La, 191Ir, 193Ir, 185Re and 209Bi) to check possible weird out coming results. If the standardising is changed, it is indicated the change and the reason. Some elements precipitated when left to stand for several hours, not surprisingly, since Thiourea (TU) has been used to quantitatively precipitate some of the PGE and gold facilitate their separation. Furthermore, fresh Thiourea is continually added to the column, so the equilibrium for formation of these complexes is always shifted towards dissociation; analysis with ICPMS is so difficult because the high TDS from Thiourea[412]. The recovery (R)% of analytes (ng Rh, ng Pd, ng Pt) from interfering elements (of Hf, Rb, Sr, Y, Cu, Zn, Cd, Mo, Zr, Pb) in model solution in 0.5 M HCl. After an overnight equilibration, the mass of Rh, Pd, Pt, Hf, Rb, Sr, Y, Cu, Zn, Cd, Mo, Zr and Pb of analytes in spiked samples was directly determined by ICPMS and subjected to separation via cation in column performed in 4 replicates figure (6.1).

The recovery (R)% of analytes (ng Rh, ng Pd, ng Pt) from interfering elements (of Hf, Rb, Sr, Y, Cu, Zn, Cd, Mo, Zr, Pb) in river water solution filtered through PVDF filters (Whatman, 0.45 μm), adjusted to pH 0.3 ± 0.1 with HCl. After an overnight equilibration, the mass of Rh, Pd, Pt, Hf, Rb, Sr, Y, Cu, Zn, Cd, Mo, Zr and Pb of analytes in spiked samples was directly determined by ICPMS and subjected to separation via cation in column performed in 4 replicates figure (6.2).

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The uptake % of analytes (ng Rh, ng Pd, ng Pt) in filtrated river water solution filtered through PVDF filters (Whatman, 0.45 μm), adjusted to pH 0.3 ± 0.1 with HCl. After an overnight equilibration, the mass of Rh, Pd, and Pt of analytes in spiked samples was directly determined by ICPMS and subjected to separation via S920, S914, S985, MPX-317 and MPX-310 in batch procedure performed in 4 replicates figure (6.4).

The uptake% of analytes (ng Rh, ng Pd, ng Pt) from interfering elements (of Hf, Rb, Sr, Y, Cu, Zn, Cd, Mo, Zr, Pb) in river water solution filtered through PVDF filters (Whatman, 0.45 μm), adjusted to pH 0.3 ± 0.1 with HCl. After an overnight equilibration, the mass of Rh, Pd, Pt, Hf, Rb, Sr, Y, Cu, Zn, Cd, Mo, Zr and Pb of analytes in spiked samples was directly determined by ICPMS and subjected to separation via S920 in column performed in 4 replicates figure (6.5). 6.2.3. Separation solutions preparations Borosilicate glass were used to prepare all model solutions and samples by spiking Rh, Pd and Pt and potential interferents in 0.5 mol L−1 HCl. Model solutions were prepared by dissolving the appropriate amount of HCl and spiking micro volume of the standard solutions. Filtrated river water prepared by adjusting the pH = 0.3 ±0.1 with HCl. All samples were allowed to stand for 24h before the tests. 6.2.4. Cation exchange pre-treatment, column preparation and pre- concentration with heating procedure New or used resin, Dowex 50W-X8(H) (Bio-Rad Laboratories, Richmond, CA) (20 g) to be carefully cleaned by stirring with 20 mL of 4mol L−1 HCl solution (six times or until no further discoloration of the acid was observed) to remove most cationic species from the resin, replacing them with H+ ions to maintain overall neutrality of the resin and this not only removes matrix elements, but also effectively desorbs any residual PGE’s and Au[373]. Next with Milli- Q water (twice). Then, with 0.5 mol L−1 HCl (twice) for about 20 min. The glass columns tube of dimension 0.8 cms × 10 cms to be packed with 3.5 g of the clean resin. Care was taken to avoid air bubbles in the column. Close the tap, and add a small amount of 0.5 M HCl to the column. Pour 3.5 cm of slurried resin into the column. Open the tap, and allow the acid to drain from the column as the resin settles. Add additional 0.5 M HCl to the column, if necessary, to prevent the solution level dropping below the top of the resin bed. The settled height should be 3.5 cm; resin may have to be removed or added to achieve this. If additional resin has to be added, it is essential that the resin bed is re-homogenised to prevent dis-continuities and

155 preferential channelling of solutions through the resin bed. Both ends of column tube were blocked with glass wool. Conditioning the resin: The resin to be equilibrated with 10 mL of 0.5 mol L−1 HCl with maximum flow rate (1 ml min-1) through the column. The blank, obtained by passing 0.5 mol L−1 HCl through the column, is to be always determined prior to loading the sample in order to control possible contamination. Loading the sample: 200 ml sample of PGEs in 0.5M HCl (or pH = 0.3 for filtrated river water samples) (50 ml sample for each batch). The batches are collected with 50 ml enlimars and measured separately then the results are summed. Batches are filtered using 0.45 μm disposal mixed cellulose ester syringe filters (Chromafil) pre-washed with 10 mL of 1.4 M HNO3. The PGE eluate should remain colourless, indicating a successful separation. Wash the column with 3 ml of 0.5 M HCl and combine it with the sample. Fractions were diluted two times just before the ICPMS analysis. Cleaning the resin by adding 10 ml 4 M HCl and elute the column. This fraction contains the matrix elements and is discarded. Remove the resin from the columns after each run. Combine the resin from all columns, homogenise and batch clean the resin with 4 M HCl. Homogenisation of the resin minimises memory effects, ensures that separations are performed under reproducible conditions, and allows the determination of representative procedural blanks.

The recovery % is calculated as mass difference between the initial and fraction passed the column Figures (6.1 and 6.3). Repeatability of the results expressed as a relative standard deviation of 4 replicates Figure (6.1). The reproducibility of results expressed as RSD of 4 replicate determinations of analytes in river water Figure (6.3). In the next step is to test the pre-concentration by evaporation procedure, the fractions are collected in one borosilicate glass stock and homogenised. The fractions are transferred into 4 borosilicate beakers pre-cleaned with hot 10% HNO3 for 12h followed by hot 10% HCl for another 12h (200 ml each) and covered with 1 µm filter paper to prevent anything to fall in the beakers. The beakers are placed on electric heating plate at 85°C in the fume hood and the volume brought to about 20ml. Transfer the solution from the 200-ml beaker to a 50-ml borosilicate glass beaker pre-cleaned as the other glass beakers, rinsing the walls of the 200-ml beaker with 0.5 M HCl. Evaporate to around 2 ml on the heating plate and transfer the contents to a l0 ml volumetric flask, rinsing the walls of the 50-ml beaker with 0.5 M HCl until made to volume. Total evaporation or boiling the solutions should be avoided during this test. The last step, transfer the sample solution to a PP vial for storage in the fridge at 4°C for measurement

156 with the ICPMS. The recovery % is calculated as mass difference between the initial and the measured mass in the 10 ml of the last remaining Figure (6.2).

6.2.5. Chelating resins pre-treatment for static mode (Batch process) If the resin (S920, S914, S985, MPX-317 or MPX-310) supplied contained a much wider range of grain-sizes than the nominal fraction, these to be removed by slurring the resin with deionised water, allowing the resin to settle for a short time, and then decanting off the suspended fines with the supernatant. This was repeated at least three times to remove most of the fine material [412]. New resin were prepared according to standard methods and loaded with 1 M NaCl aiming to convert the resins into the chloride form[112] by soaking in 10 times the volume of the wet settled resin. This is followed by rinsing the resin with MQ-water for 3 times for 30 min to wash off any remaining chloride-containing solution[205]. Then 1 gr of each resin was brought in contact (replicates =3) with test solutions (200 ml filtrated river water pH = 0.3 ±0.1 adjust with HCl) using borosilicate glass for doing and placed on the shaker for 48h at room temperature. The uptake % of Rh, Pd and Pt on the resin was determined by the difference between its concentrations in the initial and equilibrium solutions Figure (6.4). 6.2.6. Chelating resin S920 pre-treatment and column preparation The resin supplied contained a much wider range of grain-sizes than the nominal fraction, these to be removed by slurring the resin with deionised water, allowing the resin to settle for a short time, and then decanting off the suspended fines with the supernatant. This was repeated at least three times to remove most of the fine material [412]. New resin was prepared according to standard methods and loaded with 1 M NaCl aiming to convert the resins into the chloride form[112] by passing 10 times the volume of the wet settled resin. This is followed by rinsing the resin with MQ water at a flow rate of 10 BV/h for 30 min to wash off any remaining chloride-containing solution[205]. Loading on the column, columns packed with an ion exchanger (3.5 gr) of commercial granulation[220]. The column is conditioned by elution with 0.5 M HCl, three times the bed column volume. The blank, obtained by passing 0.5 M HCl through the column, is to be always determined prior to loading the sample in order to control possible contamination. Loading the sample: 200 ml sample of PGEs in filtrated river water samples of pH = 0.3 ± 0.1 adjusted with HCl, where the resin shows very good sorption capacity at these condition [186]. The batches are collected with 200 ml enlimars of borosilicate and measured using ICPMS. The PGE eluate

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should remain colourless, indicating a successful separation. Wash the column with 3 ml of 0.5 M HCl and combine it with the sample Figure (6.5). To test possible elution of PGEs from the column further tests were conducted. In the first step, the columns washed with 0.3M HCl [220] to remove the remaining of the sample and to wash out the undesired elements. Elution with acidified Thiourea (1M) in 2M HCl or with hot aqua regia at 60°C were investigated. The elution with acidified Thiourea was tested by passing (2 ml×3 extractions) of the acidified Thiourea at room temperature then passing the other (2 ml×3 extractions) of the acidified Thiourea heated to around 80°C. The fractions are collected with Borosilicate glasses and diluted ten times with MQ water and filtered using 0.45 μm disposal mixed cellulose ester syringe filters (Chromafil) pre-washed with 10 mL of 1.4 M HNO3 then measured separately with ICPMS. Results are calculated by summing the measured fractions Figure (6.6). Elution with hot aqua regia at 80°C was tested by removing the resin from the column into Teflon bottles then adding 3ml of freshly prepared aqua regia and then close the Teflon bottles and place them in an oven at 80°C for 24 h, then samples are left to cool down and the bottles are opened, then 27ml MQ water was added and homogenised and then the resin separated using filtered using 0.45 μm disposal mixed cellulose ester syringe filters

(Chromafil) pre-washed with 10 mL of 1.4 M HNO3 and solution measured with ICPMS Figure (6.7). Resin regeneration; The resin is developed for industrial applications and quantitative recovery of PGEs from this resins can be only achieved by ignition or very aggressive eluent which destroy the resin hence difficult to reuse this type of resin (This type of resin is quite cheap = few Euros/ litre). 6.2.7. Limits of detection Detection and determination limits to be calculated using the method outlined [154, 413]. The calculated detection and determination limits for a given mass or isotope are based on the slope of a line generated by a blank and standards of known concentrations. The minimum distinguishable signal, Sm, is calculated using the following relationship:

Sm = Sb + k×σ 1

Where Sb = mean blank signal, k=constant, σ = standard deviation of average blank signal (n=4). Detection limits are calculated using k=3, while determination limits use k=10, a value proposed by the American Chemical Society and Committee on Environmental Improvement

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1980. Calculated detection limits provide a concentration at which the ICP-MS technique can differentiate between random background noise and a particular isotope in solution. Such detection limits will vary from day to day due to machine sensitivity and background noise levels, making it difficult to achieve such low detection limits on a reproducible basis. Therefore, determination limits yield a more reasonable working concentration range that can be accurately reproduced on a regular basis for a particular isotope and provide a good approximation of a lower limit of quantitative analysis.

The recovery “fR” of PGEs in the samples after chelating separation were calculated based on the following quantitative formula the ratio of eluent mass to sample mass. 6.3.Results and discussion 6.3.1. Blanks and Limits of detection Table (7.2) shows the blank values are obtained for different eluents and are reported as concentrations in solution (µg. L-1). Blanks were measured to establish the background level of the Rh, Pd and Pt throughout the run, and blank corrections were applied where necessary. The instrumental blanks obtained by preparing the corresponding eluent in PP and measuring with ICPMS. The resin blank, obtained by passing eluent through the column, was always determined prior to loading the sample in order to control possible contamination. The methodological detection limit (MDL) for each element was calculated based on three times the standard deviation of 4 replicate measurements of blank columns and using eq 1. The measured method detection limits for all elements were better than 10 ng L-1 (except for TU). However, these do not translate directly to improved lower limits of determination for samples, because limitations on the types of solution analysed are imposed by the level of total dissolved solids (TDS). Upper limits of approximately 0.2% TDS may be aspirated into an ICP-MS instrument fitted with a standard interface, without causing significant instrumental drift and/or matrix effects, necessitating higher dilution factors for solutions. Nevertheless, with such low detection limits, ICP-MS is ideally suited to the determination of low concentrations of Rh, Pd and Pt in simple solutions, and is also far less prone to interferences than most other instrumental techniques.

Table 6. 1. Blank and MDL (ng L-1) 103Rh 105Pd 106Pd 108Pd 195Pt Instrumental blanks of (TU_HCl) _ blank 5.91 38.23 9.914 7.19 1.55 0.2MTU in 0.4M HCl MDL 8.16 52.84 11.56 8.72 2.12 S920 blanks of (TU_HCl) _ blank 8.9 53.86 4.82 6.98 2.16

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0.2MTU in 0.4M HCl MDL 11.68 69.11 8.12 24.6 8.82 Instrumental blanks of blank 3.03 6.34 6.81 3.4 0.56 0.5M HCl MDL 3.66 7.98 7.54 5.3 0.94 S920 column blank 4.1 3.92 4.30 7.67 0.20 (0.5M HCl) MDL 4.4 5.79 13.5 10.64 1.1 Cation column blank 3.54 6.58 6.87 3.83 0.90 Dowex 50W-X8 (0.5M HCl) MDL 4.17 7.38 7.16 6.1 1.24 Instrumental blanks of blank 1.18 12.98 5.12 4.7 2.53 0.1TU in 0.2M HNO3 MDL 3.91 41.88 6.92 7.14 4.63 S920 blanks of blank 1.45 12.07 7.51 7.09 4.19 0.1TU in 0.2M HNO3 MDL 2.35 16.46 12.31 8.7 11.25 Cation column Dowex 50W-X8 blank 16.14 174.6 17.06 10.64 3.7 Followed by non-boiling evaporation (0.5M HCl) MDL 19.35 205.314 24.59 14.67 4.92

6.3.2. Separation of matrix ions on cation-exchange resin:

Ion-exchange methods are based on the fact that PGEs forms highly stable anionic chloro complexes, while other metals form weaker anionic or stable cationic species[414]. The interfering cations are exchanged and adsorbed on the cation- exchange resin, while the analyte is passed through the clean-up column. The efficiency of such separation depends on the kind of resin, acidity of the sample, presence of sufficient amount of chloride ions, size of the column and mesh of the resin, as well as occurrence of other complexing reagents in the sample solution. In order to evaluate the performance of Dowex 50W-X8 for determination of Rh, Pd and Pt by ICP-MS in a model solution, the isotopic of most abundant isotopes of Rh (103Rh), Pd (105Pd, 106Pd, or 108Pd) and Pt (195Pt) were measured in effluate and it was found high recoveries obtained. The obtained recovery of analytes (470 ng Rh, 480 ng Pd and 580 ng Pt) in the model solution from interfering elements (3.3 – 51543 ng of Y, Cu, Zn, Hf, Mo, Zr, Pb, Rb and Sr) in solution prepared in model solution of 0.5 M HCl was 97.01 ± 1.25% (n = 4) for Rh, 93.75 ± 4.86 % (n = 4) for Pd and 97.81 ± 2.31 % (n = 4) for Pt (Figure 7.1). These recoveries are similar of the previous findings as for Pt and Pd similar in [370], and Rh in [372] for similar modelling solution. Repeatability of the results expressed as a relative standard deviation of 4 replicates and it was 1.29% for Rh, 5.18% for Pd and 2.36% for Pt. Almost complete separation (retention) of Rb (98.2%) and Sr (99.4%) in the column using the model solution (Figure 6.1). Low separation efficiency in cation exchange or simultaneous retention on anion exchange was previously reported for Pb, Zn, Cu, Y [150, 415], Hf, Zr, Y

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[154, 204, 371] and Mo[108], introducing spectral interference during detection step. These metals tend to form stable anionic complexes; this propensity depends on sample type and dissolution technique. Zr and Hf have high standard deviation due to their leach profile, when we applied the sample in 4 steps and we collected the four fractions separately and we measured them separately then the Hf and Zr showed different leach (concentrations) in different fractions but the total leach is equal to the original amount added (Figure 6.1). Complete separation is reported for Hf, Y and Cu in[370] while in this study Hf (54%), Y(22.8%) and Cu(86.8%) are retained in the column when using similar model solution but several order magnitude mass is used in comparison with this study. In this study, Mo and Cd leach completely with Rh, Pd and Pt from the column as Cd[416] and Mo[417] form anionic species at these conditions and they are not retained on cation column at similar pH. Copper, Zink and lead show high retention in the column of 86.8%, 91.5% and 95.9 %, respectively (Figure 6.1).

150 Recovery % Recovery %

100

50

column 0 C1 C2 C3 C4 number Rh Pd Pt Sr Rb Cu Zn Y Hf Mo Cd Zr Pb

Dowex 50W-X8 150 Mean Recovery %

100

50

0 Element Rh Pd Pt Sr Rb Cu Zn Y Hf Mo Cd Zr Pb

Figure 6. 1. Recovery % (not retained by the column) (±S.D.%, n = 4) in model solutions (200 ml/ 4 fractions) of Rh (470 ng), Pd (480 ng), Pt (537 ng), Sr (51543 ng), Rb (15184 ng), Cu (2319 ng), Zn (9139 ng), Y (3.3 ng), Hf (5.9 ng), Mo (1685 ng), Cd (69 ng), Zr (4.2 ng), and Pb (680 ng) in 0.5M HCl on Dowex 50W X8 200-400 mesh cation-exchanger determined by ICPMS. The first graph: C1, C2, C3 and C4 are four replicates of the column separation (uncertainty is the standard deviation of the C1, C2, C3 and C4 in %). The high uncertainty represents the unrepeatable behaviour of the elution. The second figure represents the mean recovery of the columns C (1-4). Dilution factor is 2 for ICPMS measurements.

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Because of the ultralow concentration levels of PGEs in surface water [20, 103, 375] and the cation exchange procedure does not serve as pre-concentration step, it is necessary to pre- concentrate the samples before the instrumental analyses to improve the detection limit. Pre- concentration was performed by non-boiling evaporation for the model solution to increase the concentrations by a factor of ∼20 after the Dowex 50W-X8 cation exchange to remove most of the cations including partial removal of the Na [418], which high concentration hinder the ICPMS measurement. The recovery and precision of the pre-concentration step was confirmed with n = 3 (T2, T3 and T4) replicates of the model solution collected from the effluent in the previous step after the cation exchange step and presented as recovery mean of the three replicates (R_ Mean) Figure (6.2) for Rh, Pd and Pt (100.13± 3.05, 98.83±9.75, 95.92± 8.00) %, respectively Figure (6.2). During the experimental test, total evaporation was avoided but * accidently T1 was totally evaporated. After total evaporation, the walls of the beaker were washed with 0.5 M HCl and followed the same steps of the others replicates but the measured recoveries for Rh, Pd and Pt were lower than the others and hence the replicate T1 is excluded from the calculations. Experimental blanks were 16.4169 ± 1.359, 170.3178 ± 2.1523, and 4.09027 ± 0.79778 ng L-1 for Rh, Pd and Pt, respectively. Despite the ultra-care taking during this test but increased the blank was observed (see table 6.1).

R%_ preconcentration by non-boiling evapration 150 100 50 0 T1* T2 T3 T4 R_Mean

Rh Pd Pt

Figure 6. 2. Pre-concentration by heating procedure. Mean recovery R_ Mean%: Rh (100.13±3.05), Pd (98.83±9.75), and Pt (95.92±8.0). (±S.D. %, n =3; T2, T3 and T4). The * individual replicates considered are (T2, T3 and T4). Replicate (T1 ) is excluded from the calculation due to the high loss. Error bars represent the standard deviation in% of 3 replicates.

Besides the use of model solutions (containing standards and studied interferents), it was of interest to evaluate the performance of the proposed approach for the real environmental river samples after pre-treatment procedure. In this case, samples were collected from the study site (Zenne River, Belgium). Then the water is filtrated and the pH is adjusted to 0.3 using HCl then spiked with micro-volumes with highly concentrated standards of Rh, Pd, Pt, Zn, Cu, Hf,

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Zn, Mo, Zr, Cd and Pb due to their low concentration while Rb and Sr are not spiked due to their high concentration in river water. The pH was not changed after the spiking. The solution was left over night before perform cation column separation (Figure 6.3). The average recovery of rhodium is about 81.4% ± 0.83% and this is similar of the previous published work for river water [372] while for Pd is 85.8 % ± 2.4% and this is similar of the previous published work at 0.5M HCl [414] and for Pt is 103.9%± 2.2%. The reproducibility of Rh, Pd and Pt results expressed as RSD of 4 replicate determinations of analytes in river water and they are 1.0, 2.8 and 2.1, respectively (Figure 7.3). Similar findings to the model solution, about average (99.9±0.0028) % of Sr, (96.4±0.1) % of Rb, (73.9±5.7) % of Cu, (71±5.5) % of Zn and (97.4±1.2) % of Pb are retained on the column after separation of filtrated river water matrix at pH 0.3 ±0.1. Unlike the model solution, Y shows better separation when filtrated river water matrix is used where about (96.6±0.2) % of Y retained on the column (Figure 7.3). Cd was not separated from the PGEs when pH =0.3 ±0.1, as Cd form anionic species at highly acidic conditions pH < 2 [416, 419]. Mo is not separated at acid conditions due to the formation of anionic species at pH 0.3 ±0.1 [419, 420]. The sorption of kinetically inert complexes of rhodium and iridium requires prolonged heating [421]. Different techniques of the labialization of inert complexes of platinum-group metals in solutions are used for more efficient pre-concentration of noble metals. For example, the recovery of rhodium substantially increases in the presence of SnCl2 [422].

FRW 140 R %

120

100

80

60

40 column 20 number

0 C1 C2 C3 C4 Rh Pd Pt Sr Rb Cu Zn Y Hf Mo Cd Zr Pb

163

FRW 120 R %

100

80

60

40

20 Element 0

Rh Pd Pt Sr Rb Cu Zn Y Hf Mo Cd Zr Pb

Figure 6. 3. Recovery (Not sorbet fraction by the column) (±S.D.%, n = 4) in filtrated river water (200 ml/ 4 fractions) of Rh (470 ng), Pd (480 ng), Pt (537 ng), Sr (51543 ng), Rb (15184 ng), Cu (2319 ng), Zn (9139 ng), Y (3.3 ng), Hf (5.9 ng), Mo (1685 ng), Cd (69 ng), Zr (4.2 ng), and Pb (680 ng) on Dowex 50W X8 200-400 mesh cation-exchanger determined by ICPMS. The first graph: C1, C2, C3 and C4 are four replicates of the column separation (uncertainty is the standard deviation of the C1, C2, C3 and C4 in %). Dilution factor is 2 for ICPMS measurements.

6.3.3. Rh, Pd and Pt pre-concentration using different chelating resins using batch process

The static sorption mode is used to evaluate the best resin for pre-concentration of Rh, Pd and Pt at low acidity pH = 0.3±0.1 (eq of 0.5M HCl), where PGEs show high stability in Borosilicate glass. The resins particles (1 gr of each) are shaked with a large volume (200 ml) of filtrated river water solution of 1 µg L-1 of Rh, Pd and Pt at pH= 0.3 ±0.1 adjusted with HCl (n= 3 replicates for each resin). Subsequently the uptake % was calculated as the difference between the initial and the equilibrated solution in % Figure (7.4). The adsorption of Rh on all resins is presented in Figure (7.4). Assessment of these recoveries indicates that all resins exhibit poor adsorption selectivity for Rh in the given system (0.5M HCl = pH=0.3). The widely acceptable explanation behind the relatively small extraction of Rh is the higher kinetic inertness of Rh chloride complexes as compared to Pt and Pd chloride complexes. This follows the order in which metal-chloro complexes are more likely to form 2− 2− 3− ion pairs with anion-exchangers: [MCl6] > [MCl4] ≫ [MCl6] > aqua species. This order is determined by the charge-to-size ratio or charge density of the species. Species with low

164 charge density are more easily paired than species with higher charge density[423]. This can be explained by the size of the hydration shells surrounding the ions: densely charged species have larger hydration shells which in turn have lower coulombic interaction with their counter ions than those with smaller hydration shells [423]. The dominant PGE species typically encountered in aqueous chloride media by[424] are outlined in Table S.III.2. According to 3− 2− [425], the predominant Rh(III) species in strong HCl media are [RhCl6] and [Rh(H2O)Cl5] complexes. With the dilution of rhodium solutions, the formation of aqua chloro-complexes as + well as of the cis- and trans-isomers [Rh(H2O)2Cl4] and [Rh(H2O)4Cl2] takes place[114]. 3– Moreover, Rh metal forms binuclear complexes [Rh2Cl9] . Therefore, the existing Rh chloride complexes in the solution are less favourite adsorbed on the resins comparing to Pt and Pd. In a study [426] suggests the use of large amount of tin to the rhodium solution to activate all trivalent rhodium complexes into reduced form of monovalent rhodium complexes by the oxidation of Sn(II) to Sn(IV). Better recovery for Rh over 94% is obtained using Purolite S985 reported in [113] and [205], but relative two amount of the resin is used in model solutions in[113] and 5 times amount of the resin is used in [205] comparing to my study, which may cause the improve the recovery of Rh. Better Rh uptake is up to 94% obtained by MPX-317, S914 and S920 in [411], and this might be explained by the lower concertation used in their study of 50 ng L-1 and different matrix and the pH. Furthermore, we used in our study filtrated river as matrix and the competition effect might be the reason for lowering the Rh sorption uptake as explained for S985 and XUS 43600.00 when leach solution contains interfering elements is used instead of standard solution containing only PGEs. The ratio of the amount of the resin to the volume of the contacting solution has an influence on the recovery in batch process[205]. The adsorption isotherms of Rh on resin called XUS 43600.00 holding the similar functional group of S920 indicate a poor adsorption selectivity for Rh in high HCl acidity. Rhodium uptake is influenced by the extraction temperature, acid concentration, aging of the solution, contact time and the presence of Sn(II) [426, 427]. Improvement in the sorption of the Rh by the tested resins could be achieved by the use of higher temperature during the adsorption process [425]. The improvement could be explained by reducing the formation of cationic and neutral Rh chloride complexes at higher temperature, which in turn reduces the so-called “ageing” of the PGE solution[424]. These tests were not evaluated due the time of the thesis.

165

Table 6. 2. PGE species found in aqueous chloride media[424] Platinum Palladium Rhodium 3- [RhCl6] 2- 2 2 Pt(II) [PtCl4] Pd(II) [PdCl4] - Rh(III) [RhCl5(H2O)] - - [RhCl4(H2O)2] 2 2 2 Pt(IV) [PtCl6] - Pd(IV) [PdCl6] - Rh(IV) [RhCl6] -

The capacity of weakly, intermediate, and strongly basic anion exchange resins is decreased at increased HCl concentration in sample solution [236], which is attributed to a different mechanism of Pd(II) complexes binding[186]. The mechanism of palladium(II) complexes sorption on the ion exchangers under discussion (chelating resins) can be anion-exchange, chelating (coordinating) or mixed[186]. The anion-exchange mechanism is predominant at high acidities whereas the chelating or coordinating one at low acidities[186]. In the chloride solution palladium(II) can form stable chloride, hydroxychloride and hydroxide complexes + − 2− 2− 2− 2− which are PdCl , PdCl2, PdCl3 , PdCl4 , PdCl3(OH) , PdCl2(OH)2 , PdCl(OH)3 and 2− Pd(OH)2, Pd(OH)4 . The major species in the acidic solution containing 0.1 M and higher 2− 2− chloride concentration is PdCl4 [178, 428]. Decrease in the sorption yield of PdCl4 observed with the higher hydrochloric acid concentration in the systems under discussion is caused by − − − competitive sorption of Cl and HCl2 type ions. HCl2 ion as the ion of acid stronger than HCl exhibits a greater affinity for chelating ion exchangers [186, 236] Thiol[215], Thiourea[186, 215] functional group . This competitive sorption is not observed in the solutions without hydrochloric acid or in small hydrochloric acid concentration[186]. The acidity of the contacting solution may have some detrimental effects towards the sorption recovery of Pt and Pd on S985 [205]. Sorption recovery of S985 for all PGEs increases with decreasing the acidity of the contacting solution due to deprotonation of nitrogen atoms of functional groups[113], and clearly when increasing the amount of the resin. Purolite S985 has high capacity but poor adsorption selectivity for Pd because the isotherm has a near-linear-shaped isotherm in when high (4.11M HCl concentration [205]) and the same trend is observed in this study using 0.5 M HCl. A resin called XUS 43600.00 holding the same functional group of Purolite S920, where XUS 43600.00 shows convex isotherm profile for Pd and Pt in 4.11M HCl, which indicate high selectivity for Pt and Pd in high HCl acidity condition. High sorption selectivity for Pd in 0.3M HCl media was also reported using Purolite S920 and S924 in [220]. The chelating weak-base resin S920 with a Thiouronium functional group was introduced in this work based on the consideration of the hard and soft acids and base theory (HSAB). According to this theory, ion exchangers, which have a functional group that contains S donor atoms will interact strongly

166 with soft acids such as precious metal ions. This is in line with the experimental results obtained in this present study wherein S920 exhibited high selectivity for Pt and Pd regardless of the high acidity in the media and the competing effect from the high chloride concentration and this is in accordance with study conducted using XUS 43600.00 [205], and the same finding for resins Lewatit TP-214 holding Thiourea functional group similar to Purolite S914[215]. It is reported that increasing the HCl concentration decreasing the working capacity for Pd accumulation on S920 [186]and resin called Lewatit TP 214 [215], which has the same functional group of S914. Capacity drop with the increasing HCl concentration is attributed to a different mechanism of ion binding[215]. Ion exchange mechanism (one ion exchange) is predominant at high acidities while those chelating or coordinating (number of possible coordination is three for S914) at low acidities[215]. In the in-between media both mechanisms are assumed to compete with each other. Purolite S985, despite its high loading capacity, has relatively low selectivity for the anionic form of Pt(IV) in strong acid media when compared to S920. It has been suggested that this may be related to the effect of high acidity in the contacting medium, which reduces the deprotonation of the nitrogen atom in the functional amine group of the sorbent [113, 114]. In other words, the strong acidic media reduces or deactivates the complexation ability of the weak base sorbent Purolite S985, which results in the sorption of chloride complexes of Pt(IV) exclusively through anion exchange mechanism. Hence, the chemical structure of Purolite S985, being a weak base anion exchange resin with functional amine group, reveals itself as a poor option for the adsorption recovery of Pt and Pd in highly acidic media[205]. The predominant species of palladium (II), platinum (IV) and rhodium (III) in HCl (0.1–5 M) 2− 2− − 3− are [PdCl4] , [PtCl6] , [AuCl4] and [RhCl6] , respectively[429]. Figure (7.4), Pt sorption uptake order under pH = 0.3±0.1 follows the order S920 ≈ S914 > S985> S924 and the Pt aqua chloro- and hydroxo-complexes can be recovered by the sorbent via anion exchange mechanism, the tendency of the PGEs chloro-complexes to form ion pairs with anion 2− 2− 3− extractants according to the following [MCl6] > [MCl4] ≫ [MCl6] > aqua species as it 2− 2− 3− would be for typical metal ion complexes [PtCl6] , [PdCl4] , and [RhCl6] under this condition[205, 430]. The results of MPX-310 and MPX-317 are inconsistent with results published in 2021 [411], where low uptake for Pt and Pd is observed. In this thesis, we proofed the efficiency of MPX- 317[221] and MPX-310 ([221], Annex I) their sorption power for low Pt and Pd concentration at low concentrations after grinding the resins but when we applied the resins at high acidity of 0.5M HCl in their supplied form then they show low uptake. The low uptake of MPX-310

167 and MPX-317 cannot be explained by the acidity as those resins proved to work in very high acidity conditions up to 5.47M HCl [189, 201, 232]. The efficacy of ion exchange resins mainly depends upon their physical properties including their particle size [174], and for some extent to the concentration of the target element [112, 205]. In a study conducted by [205] has shown that Lewatit M+ MP 600 resin will only be selective for Pd given a sufficiently high concentration of Pd in the solution. Such phenomenon of enhanced affinity of the anion exchanger for traces of exchanged PGE ions was also observed by [112], whereby convex segments of the isotherms are reported only when high concentration of Pd is used [205]. In the previous studies on MPX-310 in their original form [189, 201, 232], they were conducted in such high-level concentrations where they were several order of magnitude comparing this study. Therefore, it might be the reason for such behaviour of MPX-310 and MPX-317 that their sorption power in the huge particle size is only for high concentration level while when they are grinded can be used for trace level as the physical properties has influence on resins performance. It is good to mention as well in the article [431], the authors have shown that the difference in the performance of grinded Chelex-100 compared to suspended particulate reagent-iminodiacetate. Considering the complexation of Pt, Pd and Rh with organic matter, some authors they evaluate the effect of the DOC on the resins uptake [110, 373, 400] and they show that the uptake can be reduced due to the strong complexation of PGEs with the DOC, but due to the time limitation of the research we did not consider the DOC effect on the resin uptake for PGEs. Overall, Purolite S920 and Purolite S914 show the best performance for the Rh, Pd and Pt pre- concentration under the examined conditions of 0.5M HCl (pH = 0.3±0.1) but increasing the amount of the resin is necessary to achieve complete recovery. Of all tested resins, Purolite S920 the nest results and we continued our tests using this resin.

168

% uptake 120

100

80 Rh Pd 60 Pt 40

20

0 S920 S985 S924 S914 MPX 317 MPX 310 Figure 6. 4. Uptake % (sorbet fraction by the resins using the static mode (Batch process) for 48h) (±S.D.%, n = 3) in filtrated river water using borosilicate glass bottles (200 ml, adjusted to pH 0.3 ± 0.1 with HCl) of Rh (6802 ng), Pd (3967ng), Pt (7847ng), on 1 gr of each resins of S920, S914, S924, MPX317 and MPX 310 determined by ICPMS. Filtrated River water in borosilicate (BS) glass bottles spiked with Rh, Pd and Pt and adjusted to pH 0.3 ± 0.1 with HCl. After an overnight equilibration, the concentration of analytes in spiked samples was directly determined by ICPMS. The resins were soaked in 1M NaCl for few hours. Then, the resins washed with MQ water for several times to remove any salts for 30min. In the next step, the resins conditioned with 0.5M HCl. 200 ml of the spiked FRW transferred into BS bottle and 1 gr of the resin was added and shake over 48h. The uptake (%) is calculated as the difference between the original concentration and the remaining concentration.

6.3.4. Rh, Pd and Pt pre-concentration using column procedure and S920 chelating resin Figure (6.5) shows the column pre-concertation procedure using Purolite S920 in 4 replicates. Our results show that, when test solution is passed through the columns packed with S920 resin under consideration, Pt is sorbet completely Pt= (99.9±0.1) %, while only (94.0±0.1) % of Pd is recovered. Despite the higher uptake of Rh using static mode and lower mass of the resin Figure (7.4), the use of the higher resin mass in the dynamic process Figure (7.5) shows lower Rh uptake (26.1± 3.3) %. The lower Rh uptake can be explained by the slower kinetics of Purolite S920 for Rh and the use of 1 ml min-1 is not sufficient. Hence, increasing the contact time is necessary to improve the Rh sorption uptake. The reproducibility of results expressed as RSD of 4 replicate determinations of analytes Rh, Pd and Pt in river water 13.6, 1.11, and 0.111, respectively. After loading the sample, the column was washed with 3ml of 0.5M HCl and the mass % (mass loss) was determined in this wash. Mass lost % during the wash the column with 3ml of 0.5M

169

HCl are Rh = 10.0 ±0.744, Pd=0.248 ±0.035, Pt=0.024±0.010. The high percentage loss of Rh comparing to Pd and Pt, which represents the slow sorption kinetics of Rh under these conditions and reflects the necessity to use slower flow rate to improve the Rh uptake by the resin. The average S920 uptake in % for non-PGEs using column procedure is shown in Figure (7.5). Of all interfering elements, only Cu shows high and reproducible uptake about 45%. In our previous work [356], we have shown that the Cu and the other interferents uptake could be removed using deferent soft eluents such 0.05M H2SO4. Hence, interferents uptake on Purolite S920 could be removed using softer eluents and not considered an issue in this section. Rhodium uptake is influenced by the extraction temperature, acid concentration, aging of the solution, contact time and the presence of Sn(II) [426, 427]. Improvement in the sorption of the Rh by the tested resins could be achieved by the use of higher temperature during the adsorption process [425]. The improvement could be explained by reducing the formation of cationic and neutral Rh chloride complexes at higher temperature, which in turn reduces the so-called “ageing” of the PGE solution[424]. These tests were not evaluated due the time of the thesis. Considering the complexation of Pt, Pd and Rh with organic matter, some authors they evaluate the effect of the DOC on the resins uptake [110, 373, 400] and they show that the uptake can be reduced due to the strong complexation of PGEs with the DOC, but due to the time limitation of the research we did not consider the DOC effect on the resin uptake for PGEs. Overall, Purolite S920 and Purolite S914 show the best performance for the Rh, Pd and Pt pre- concentration under the examined conditions of 0.5M HCl (pH = 0.3±0.1) but increasing the amount of the resin is necessary to achieve complete recovery.

170

FRW_ S920 120 Uptake %

100

80

60

40 column 20 number 0 C1 C2 C3 C4

Rh Pd Pt Sr Rb Cu Zn Y Hf Mo Cd Zr Pb

FRW_ S920 120,0 Uptake %

100,0

80,0

60,0

40,0

20,0 Element 0,0

Rh Pd Pt Sr Rb Cu Zn Y Hf Mo Cd Zr Pb

Figure 6. 5. Uptake % (retained fraction in the column) (±S.D.%, n = 4) in filtrated river water (200 ml) of Rh (1027 ng), Pd (1060 ng), Pt (1013 ng), Sr (42991 ng), Rb (13215 ng), Cu (142 ng), Zn (6125 ng), Y (5 ng), Hf (3.8 ng), Mo (1500ng), Cd (3 ng), Zr (12 ng), and Pb (12 ng) on Purolite S920 determined by ICPMS. Mass uptake (%) is calculated from the difference between the initial and effluent concentration in %. The first graph: C1, C2, C3 and C4 are four replicates of the column separation (uncertainty is the standard deviation of the C1, C2, C3 and C4 calculated in %). Dilution factor is 2 for ICPMS measurements. Mean Mass sorbet in % in pH = 0.3± 0.1 HCl, Rh = (26.1 ±3.3) %, Pd= (94.0±0.1) %, Pt= (99.9±0.1) %. Mass lost % during the wash the column with 3ml of 0.5M HCl are Rh = 10.0 ±0.744, Pd=0.248 ±0.035, Pt=0.024±0.010.

171

The suitability of resins for practical application is governed not only by their sorption power and selectivity, but also by the feasibility of desorption of the sorbet component. The difficulty associated with the elution of noble metals from highly selective sorbents has been investigated by a number of researchers [186, 205, 215, 220]. It is known that the elution of noble metals from highly selective ion exchangers is difficult to achieve because of strong retention of adsorbed metal ions by functional groups on the resins[432]. Thus, to achieve successful desorption of PGE-loaded resins, it is necessary to use eluent reagents that form more stable complexes with the PGE ions than the complexes of these metal ions existing in the resin phase or direct incineration/ digestion/ ignition of the PGEs-loaded resin. The two main eluents considered to date and found to be effective are acidic Thiourea solution and aqua regia [221]. Thiourea (1M) in 2M HCl and aqua regia at 60°C were tested in this study in order to assess their relative effectiveness and to select a preferred one for further study of the elution process. The tested eluents are described in Section 2. The elution of PGE from PGE-loaded S920 resin was tested using a dynamic equilibration (column) method for Thiourea 1M in 2M HCl and results are given in Figure 7.6. The elution of PGE from PGE-loaded S920 resin was tested using a batch equilibration method for hot aqua regia at 60°C for 24 h and results are given in Figure 6.7. Figures 6.6 and 6.7 show that neither acidified Thiourea nor hot aqua regia are capable of desorbing Rh, Pd and Pt from the resin under consideration but this can be explained and could be optimised. Figure 7.6 shows the higher recovery for Pt and Pd comparing to Rh and this can be explained by the stronger affinity of acidic Thiourea for Pt and Pd than for Rh [205], leading to slower rate and lower overall elution recovery of Rh from the resin. Rh and Pd desorption mechanism using acidic Thiourea is suggested/explained in [205], [186, 215] respectively. Desorption of loaded Purolite S920 was investigated by Blokhin et al. by means of aqueous ammonia, hydrochloric acid and acidified aqueous Thiourea [220], and the acidified Thiourea (0.66TU in 0.3M HCl) acts as the best eluent and 98% recovery obtained for Pd[220]. High recovery also reported for Pt and Pd of 98.7% and 99.5%, respectively when using acidified Thiourea as eluent for PGEs-loaded XUS 43600.00 resin, which has the same functional group as S920 [205]. Rh recovery reported using acidified Thiourea as eluent for PGEs-loaded XUS 43600.00 resin, which has the same functional group as S920 [48] was less than 10%, whereas the elution percentage of Rh in my study is less than 10%. Hence, the low recovery for Pt and Pd in our results Figure 6.6 can be explained small volume that is used for elution as compared with previous studies on XUS 43600.00 used the same functional group resin as S920. Rh recovery in all previous studies on resins holding similar functional groups have shown difficulties to

172 achieve which can be attributed to the kinetic inertness of the Rh complexes as well as to the stereochemistry of these complexes [205, 433]. In other words, Rh ions are more strongly retained to the ion exchangers than to the eluent considered in this paper. Hence, quantitive recovery of PGEs complexes from the used resin might be achieved by ignition only [186]. It has been reported that the desorption of Rh from a resin can be improved by the addition of oxidising agents such as NaClO, NaClO3 [434] or using HNO3, KMnO4, H2O2[426], hot Aqua regia [221] to the eluent solution taking into account the resin can be used only for single use after applying such harsh conditions. The use of high concentrated acids HCl or HNO3 may help to improve the recovery as suggested in [426]. Another suggested approach to improve desorption of Rh from ion exchange resin involves the use of higher temperature during the 3− adsorption process. As a result, more Rh would be adsorbed on the resin in [RhCl6] form, by means of anion exchange reaction mechanism, rather than in the form of oxidised Rh complexes by chelating or coordination mechanisms. Consequently, the anionic complex of 3− the type [RhCl6] can be more effectively released by the resin when the negatively charged chloride ions are replaced by the neutral thiourea ligand, forming a positively charged Rh(III) Thiourea complex[205].

S920_Thiourea (1M) in 2M HCl

120 Recovery % 100

80

60

40

20 Replicate 0 C1 C2 C3 C4 Mean

Rh Pd Pt

Figure 6. 6. Recovery % of Rh, Pd and Pt from S920 resins under dynamic conditions using Thiourea (1M) in 2M HCl eluent. Mean recovery (Mean elution) % represent the average of the recovery of 4 replicates. Error bars represent the uncertainty (±S.D.%, n = 4) of 4 replicates. PGEs-loaded resins mass are as follows: Rh=267.02 ng, Pd= 966.4 ng, Pt = 1011.98 ng. Reproducibility of Rh, Pd and Pt results expressed as RSD of 4 replicate determinations of analytes in river water and they are 17.7, 15.4 and 15.7, respectively.

173

Figure 6.7 shows the elution recovery using hot aqua regia at 60°C for 24 h in batch method. After the resins is loaded with PGEs, then the resin was washed with 0.5 M HCl, then with MQ water then the resin is collected in 5 ml Teflon bottle and then 3ml of freshly prepared aqua regia is added and the bottle is paced in oven at 60°C for 24h. Despite the application of the high harsh conditions but low mean recovery (n =4) is obtained (49±9) %, (72±9) % and (72±6) % for Rh, Pd and Pt, respectively, which represent of recovery factor fR(0.1274, 0.677, 0.72), respectively. The reason of this low recovery might be explained by the floating the resins over the eluent, which causes low recovery and low reproducibility of Rh, Pd and Pt results expressed as RSD of 4 replicate of 17.52, 12.38 and 7.94, respectively Figure 6.7. The hot aqua regia is showing promising results and increasing volume of the eluent may improve the obtained recoveries.

S920_ aqua regia 100 Recovery % 90 80 70 60 50 40 30 20 10 0 Replicate B1 B2 B3 B4 B_Mean Rh Pd Pt

Figure 6. 7. Recovery % of Rh, Pd and Pt from S920 resins under static mode using hot aqua regia 60°C eluent for 24 h. Mean recovery (Mean elution) % represent the average of the recovery of 4 replicates. Error bars represent the uncertainty (±S.D.%, n = 4) of 4 replicates. PGEs-loaded resins mass are as follows: Rh=267.02 ng, Pd= 966.4 ng, Pt = 1011.98 ng. Reproducibility of Rh, Pd and Pt results expressed as RSD of 4 replicate determinations of analytes in river water and they are 17.51, 12.38 and 7.94, respectively.

The problem of the low recoveries of PGE-loaded Purolite S920 resin using either acidic Thiourea in dynamic mode or the hot aqua regia in static mode is definitely because of the low volume and the low temperature of the eluent. In order to prove this, we re-tested the elution steps ( aqua regia at 60°C or using acidified HCl Thiourea at 60°C) using S920 resin gels – made as described in [221]- instead of using S920 resins beads because it is easier to make identical replicates of PGE-loaded Purolite S920 resin gels disks compared of making PGE-

174 loaded Purolite S920 resin. Furthermore, the use of resin gel requires small amount of the resins (about 0.1 gr resin S920/ desk) hence, the use of 1 ml of hot aqua regia or 1 ml of acidified Thiourea represent higher amount of the eluent comparing the use of resin beads. The results (section 6.3.5.) have shown that total recoveries of PGEs could be achieved when using hot aqua regia or hot acidic Thiourea section. 6.3.5. Eluent and elution condition effect Low elution recovery of PGEs-loaded Purolite S920 is obtained during the use kinetic mode and hot 1M Thiourea acidified with 2M HCl or the use static mode and hot aqua regia in Annex III. The low recovery may have explained due to the low volume of the eluent comparing to huge resin. Hence, to verify this hypothesis and to evaluate the temperature effect and the use of oxidative conditions (HNO3) on the elution, fifteen DGT units with diffusive layers (agarose diffusive gel (0.75 mm) and filter (0.17 mm)) overlaid on binding resin gel of AGA impregnated with binding gels made from S920 as described in [221] of each were deployed in 7 litters filtrated river collected from the study site and spiked with 10-15 μg L−1 of PGEs. The test solution was left for 24 h for equilibration before the deployment. The use of the S920 resin gels instead of using S920 resins beads because it is easier to make identical replicates of PGE-loaded Purolite S920 resin gels disks compared of making PGE- loaded Purolite S920 resin. Furthermore, the use of resin gel requires small amount of the resins (about 0.1 gr resin S920/ desk) hence, the use of 1 ml of hot aqua regia or 1 ml of acidified Thiourea represents higher amount of the eluent comparing the use of resin beads. Master variables (T = 21 °C, pH = 8.6, salinity = 0.4, DOC = 4.3 mg L −1) were measured before the deployment and at the end of the test. The DGT probes were exposed for 24 h in the stirred- solutions. Then the pistons retrieved from the solution and washed with MQ water. Then, the DGT assemblies divided into five groups. The first group (AR) were eluted immediately using hot aqua regia procedure at 60°C as described in section 2.5.3. in the article [221]. The second and third group (TU_ HCl) were eluted using 1 ml of 1M Thiourea in 2M HCl at different temperatures 25°C and at 60°C for 24 h. The third group (TU_ HNO3) were eluted using 1 ml of 1M Thiourea in 2M HNO3 at different temperatures 25°C and at 60°C for 24 h. After the 24h, the gels were cooled down and diluted 10 times, the resin separated from the solution and measured with the ICPMS, and the results are shown in Figure 6.8. Hot aqua regia procedure is previously reported to be the most efficient eluent for PGEs from resin gel using Purolite S920 [221, 411], with recovery factor over 90% for all PGEs so the AR is going to be the reference for comparison the other eluents as shown in Figure 6.8. Rh:

175

Hot aqua regia procedure was previously [221] reported to elute over 95% of the accumulated Rh using S920 resin gel. Figure (6.8) shows higher recovery is obtained when using hot temperature T60°C and Thiourea eluent than room temperature T25°C. The improvement of recovery of acidified Thiourea using higher temperature was also reported in [411] and it is reported in the same study that the elution efficiency was highly temperature-dependent with optimal performance at around 60◦C. The elution efficiency of Thiourea acidified with HNO3 is higher than using

Thiourea acidified with HCl for both cases of T60°C or T25°C because HNO3 acts both as stripping and oxidizing agent [426]. It has been reported that the desorption of Rh from a resin can be improved by the addition of oxidising agents such as NaClO, NaClO3 [434] or using

HNO3, KMnO4, H2O2[426], hot aqua regia [221] to the eluent solution taking into account the resin can be used only for single time after applying such harsh conditions. The use of high concentrated acids HCl or HNO3 may help to improve the recovery as suggested in [426].

Hence, the use of Thiourea acidified with HNO3 at 60°C or aqua regia at 60°C for 24h gives the best Rh elution recoveries. However, no significant improvement in Rh the elution from S920 is reported [411] with increasing the HCl concentration from 1 to 3 mol L− 1.

Pd: Pd elution efficiency with Thiourea acidified with HCl is higher than Thiourea acidified with

HNO3 Figure (6.8), this might due to the higher stability of Pd with chloride than nitrate. Temperature shows insignificant effect for eluting Pd loaded- S920 using Thiourea acidified with HCl. The negative temperature impact on Pd elution form S920 at 80°C is reported [411] and that might explain the slightly lower recovery of Pd at 60°C TU_HCl than the 25°C but the difference is insignificant Figure 6.8. Desorption of Pd loaded Purolite S920 [186, 220] or other resins holding the same functional group [205] have been investigated using different eluents and all the investigations no hot temperature was implanted but the acidic Thiourea with HCl shows the best desorption efficiency over 98%[220], 93.3%[205], and It is concluded that quantitive recovery of palladium (II) complexes from the used resins can be only achieved by ignition [186]. Overall, hot aqua regia is the most efficient procedure for Pd- loaded S920 elution. Pt: Figure (6.8), Pt elution from Pt-loaded S920 shows high elution recovery for all eluents applied and temperature. High recovery for Pt of 97.9 % using Thiourea 1M in HCl 2M

176 eluent when applied for XUS 43600.00 holding similar functional group. Temperature shows no effect on Pt elution using S920 resin gel and acidic Thiourea. Overall, the use of hot temperature condition at 60°C for 24 h is critical to achieve high Rh recovery using acidic Thiourea, and the use of HCl instead HNO3 shows better recovery for Pd while Pt shows the possibility for elution at all conditions. From a practical point of view, after long working with both eluents acidic Thiourea and aqua regia, I recommend the use hot aqua regia eluent instead using acidic Thiourea. The use of acidic Thiourea may arise several problem during the measurement with ICPMS such destabilizing the plasma, increase the blank, blocking the introduction system of the ICPMS, which makes the necessity to run huge amount of Thiourea blank to stabilize the plasma and decrease the blanks reads. Diluting the Thiourea is important as the high concentration of the Thiourea causes blocking for the introduction system of the ICPMS but diluting the sample decreases the sensitivity of the method. Thiourea does not show long stability for PGEs in the solution sample. Hence, hot aqua regia is the more preferable procedure as it shows higher stability for PGEs, less effect on plasma stability, less dilution required.

T25⁰C T25⁰C Pt Pd Rh Mass 300 500 T25⁰C Mass T60⁰C 60 T60⁰C (ng) Mass 250 (ng) 50 (ng) 400 T60⁰C 200 40 300 150 30 200 100 20

100 50 10

0 0 0 AR TU_HCl TU_HNO3 AR TU_HCl TU_HNO3 AR TU_HCl TU_HNO3 Figure 6. 8. Rh, Pd and Pt elution efficiency (ng/ disk) using different eluents, the reference is (AR) hot aqua regia at 60°C, Thiourea 1M in 2M HCl (TU_HCl) at room temperature T25°C and T60°C and Thiourea 1M in 2M HNO3 (TU_HNO3) at room temperature T25°C and T60°C. Error bars represent standard deviation of 3 replicates. Rh, Pd and Pt data are normalized to 115In, 209Bi and 185Re.

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6.4. Conclusions: In this chapter we present incomplete work due to time constrains and further optimisation are still possible. We tried to offer a simplified method for pre-concertation/ separation of Pt, Pd and Rh in surface water matrix using Purolite S914, S920, S985, Italmatch MPX-317 and MPX-310. The evaluated methods are still requiring further work but I present here the obtained results and possible optimisation in the future. Cation exchange combined with non-boiling evaporation method showed good results but it suffers from increasing the blank and ultra-clean environment is mandatory for such test. Purolite S920 offers good separation pre-concertation for Pt and Pd but slow uptake for Rh which requires longer contact time. Hot aqua regia or acidic Thiourea are excellent eluents for Purolite S920 but they must be used at higher temperature and higher volume. In the future, it might possible the use of the combination of Cation exchange followed with S920 extraction and hot aqua regia elution.

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Chapter 7: Distribution of platinum (Pt), palladium (Pd), and rhodium (Rh) in urban tributaries of the Scheldt River assessed by Diffusive Gradients in Thin- Films Technique (DGT)

Abdulbur-Alfakhoury, Ehab Guillaume Trommeter, Natacha Brion, David Dumoulin, Marek Reichstädter, G. Billon, Martine Leermakers, W. Baeyens. “Distribution of platinum (Pt), palladium (Pd), and rhodium (Rh) in urban tributaries of the Scheldt River assessed by Diffusive Gradients in Thin-Films Technique (DGT)” (2021), Science of the Total Environment 784.

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Abstract:

The performance of the newly developed DGT technique for the platinum group elements (PGEs) rhodium (Rh), platinum (Pt) and palladium (Pd) was evaluated in two tributaries of the Scheldt River, the Marque River close to the city of Lille (France), and the Zenne River which flows through the city of Brussels (Belgium). In the Marque River, an interlaboratory comparison was performed between the two laboratories where the DGT techniques dedicated to PGEs were developed (AMGC, VUB & LASIRE, U-Lille). PGEs were also analysed in an effluent of a Brussels hospital and monthly grab sampling was performed at the wastewater treatments plants (WWTPs) of Brussels. The concentrations of the 3 elements are higher in the Zenne River than in the Marque River and much higher Pt concentrations are found in the hospital effluent. Good agreement for Pt was observed between the three selected chelating resins and a relatively good agreement was observed between the two laboratories using the same chelating resin, whereas lower results were observed with the anion-exchange resin. Larger discrepancies between the two laboratories were observed for Pd and no comparison could be made for Rh due to the low natural concentrations. The results show that in small urban rivers with high impact of urbanisation, WWTPs are an important source of Pt, resulting from the use of anticancer drugs in hospitals and households. The limited retention of PGEs in WWTPs results in increased concentrations in urban rivers downstream. For Pd and Rh, similar trends were found with other traffic related elements such as Cu, Zn and Pb, showing the highest concentrations in waters collecting runoff from a highway. The data show that these elements, together with Gd, can be useful to trace specific pollution sources and their dispersion.

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7.1. Introduction

The platinum group elements (PGEs), particularly platinum (Pt), palladium (Pd) and rhodium (Rh), are nowadays increasingly emitted into the environment due to the worldwide usage in vehicle catalysts, industrial applications, the manufacture of jewellery, as alloys in dentistry and especially for Pt, in cancer therapy [23, 142]. Catalytic converters are one of the main sources of PGE emissions, containing Pd, Pt and Rh as active components in various ratios [20, 353, 435, 436]. Since mid ‘90s, because of the increased price of Pt, Pd has partially substituted Pt whereas the proportions of Rh have always remained low and relatively constant [141]. PGEs occur in particle sizes ranging from <1 µm (nanoparticles) to >63 μm in automobile exhaust and in the urban environment [56], supporting that the emission is a combination of processes such as chemical and thermal aging [34]. During rain events, the deposited particles can then be washed into sewers, rivers and water bodies, where they accumulate in sediments [437]. In addition, chemical transformation is suggested by the occurrence of soluble PGE in automobile exhaust. Whereas soluble Pt represents less than 10% of total Pt emissions, soluble Pd and Rh fractions might be greater than 50% of total emissions. Significant quantities of PGEs are converted into bioavailable forms, mainly as chloro or organic complexes [23]. Accumulation of PGEs in freshwater macroinvertebrates indicates a mobility gradient of Pd>>Rh≥Pt [354]. PGEs are also widely used in medicine. For example, the cancerostatic platinum compounds (CPC), predominantly cisplatin, carboplatin and oxaliplatin, are very successfully used in anti- cancer chemotherapy [438] and account for about 5% of total Pt demand. Their emission to the environment occurs through the excretion of administered drugs [27]. Although these substances represent a minor fraction of platinum compared to emissions from car catalytic converters, they have a significantly higher toxicological and cancerogenic impact than catalyst-born inorganic platinum [48]. A recent study showed that platinum-based drugs could have toxic effects on aquatic organisms in urban sewers and receiving water bodies [50]. Besides, Pt-based drugs in hospital effluents are only moderately removed (51%–63%) by WWTPs in their sludge [48, 439]. Furthermore, due to the rise of anticancer home treatments, hospital effluents are no longer the main expected entry route of CPC into the aquatic environment [47]. Elevated PGE concentrations have been reported in sewage and waste [52, 440, 441]. Aquatic ecosystems can be considered as an important sink of platinum group elements. It is further evident, that there is still a lack of data concerning the bioavailability of PGEs for

181 organisms and that there is still a need of well performed and documented field studies. Even though research on platinum group elements is performed for more than 2 decades by now, their distribution and their concentration ranges in natural aquatic ecosystems remain unclear [103]. Quantification of PGEs in natural aquatic samples is indeed challenging due to their very low concentrations (often sub ng L-1 levels) and the bias introduced by possible interferences during the measurement, which requires the need for selective extraction and preconcentration procedures. Diffusive gradient in thin film (DGT) is a technique proposed by Davison and Zhang [13], and that has proven to be an accurate tool for in situ sampling of free ions and labile complexes of dissolved metals [265, 442, 443]. For PGEs, the DGT technique was recently developed independently by two research groups [221, 411]. Abdulbur-Alfakoury et al. [221] (AMGC, VUB, Brussels) used various chelating and ion exchange resins among which Purolite S914, Purolite S920 and Italmatch Chemicals IONQUEST® MPX-317, (named S914, S920 and MPX 317 hereafter) have showed the most promising results, whereas Trommetter et al. [411] (LASIRE, U-Lille, Villeneuve d’Ascq) also evaluated both a chelating resin (Purolite S920) and an anion exchange resin (AG MP-1). In the present study, field deployments evaluating the performance of the chelating resins Purolite S914 and S920; Italmatch Chemicals IONQUEST® MPX-317 were evaluated in urban rivers and an inter-laboratory comparison was carried out on the Marque River. Additional parameters and elements were also measured on the Marque River, allowing us to obtain a better understanding of the distribution and sources of PGEs and their use as potential tracers of pollution sources. 7.2.Materials and methods 7.2.1. Description of the study area The location of the three tributaries of the Scheldt River is shown in Figure 7.1.

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Figure 7. 1. A map of the Scheldt River Basin and the sampling stations in the tributaries Zenne, Marque Rivers

7.2.1.1. The Zenne River, Belgium The Zenne River (Figure 7.1) flowing through Brussels (Belgium) has an important impact on the water quality of the Scheldt river and estuary [444]. The population density in the Zenne River watershed is very high (on average 1,260 inhabitants km2), which makes it the most densely populated sub-basin of the Scheldt watershed [445], and most of the inhabitants (80%) live in Brussels city and suburbs. The Zenne River has a total length of 103 km, crosses the city of Brussels from south to north over a distance of about 20 km and receives the sewage from a combined sewer system draining a 238 km2 urban area [446]. Before the 90’s, most of the sewage generated by the population in the Zenne basin was discharged into the river without any treatment. Nowadays, the Brussels sewage is treated by 2 WWTPs: the Brussels South WWTP (360,000 inhabitant equivalent (inh.eq.), active since 2000) and the Brussels North WWTP (1,1 million inh. eq, active since 2007). Comparison of the present situation with that of the 90s (before the implementation of Brussels WWTPs) showed a sharp improvement of the river quality [447]. However, during rain events, combined sewer overflows (CSOs) are

183 responsible of an increase in contamination in the river downstream from Brussels, and such CSOs occur frequently in Brussels [448]. In addition, illicit sewer connections discharging untreated sewage water directly to the river, even in dry weather conditions, are reported [448]. The Zenne River has an annual average discharge upstream from Brussels of 4 m3 s−1 and the average flow of wastewater released by the two wastewater treatment plants (WWTPs) is in the same order of magnitude (3.9 m3 s−1). Roughly half of the Zenne river water downstream Brussels is thus originating from treated wastewaters, this proportion of wastewater being even higher during low flow periods of the river. Three sampling stations were selected for DGT deployment (Figure7.1): Z5 is located downstream of the WWTP Brussels South, Z7 is the location where the Zenne submerges from its coverage and is located upstream of WWTP Brussels North and Z9 is downstream of WWTP Brussels North. Deployments were performed in September 2018 under average river flow conditions. A preliminary investigation was carried out at station Z5 to determine whether a two-weeks deployment would result in a deterioration of the DGT performance due to saturation, competition effects or biofouling. Longer deployment times for Rh and Pd are necessary, as these elements have higher resin gel blanks and higher detection limits. DGT units containing the resins S920, S914 and MPX-317 were deployed for 5, 7 and 14 days. At each sampling interval, grab samples were also taken and filtered on 0.45 µm filters on site. In this study, only Pt was analysed, because (1) the resin blanks for Pt are very low, allowing shorter deployment times and (2) preliminary measurements showed that dissolved Hf concentrations are low so that mathematical corrections for HfO interference during measurement of Pt are <1% [66]. 7.2.1.2. UZ Brussels hospital and WWTP Brussels North and South A 3 days DGT deployment was performed in the sewer at the outlet of the UZ Brussels Hospital. This hospital has 721 beds, 30,000 hospitalizations and 360,000 consultations per year (https://www.uzbrussel.be/). The sewer of the UZ Brussels Hospital discharges directly in the municipal sewer collectors leading to Brussels North WWTP [449]. Data concerning the platinum-based anticancer drugs used in the hospital are not available, but the hospital has an important oncologic centre. From May 2019 to April 2020, samples were collected every month at the inlet (untreated sewage) and outlet (treated sewage) of WWTP Brussels North and South. After filtration at 0.45 µm and acidification with HCl to pH < 2, samples were analysed for dissolved Pt and Pd.

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7.2.1.3. Marque River, France The Marque River is a 32-km long tributary of the Deûle River, which is a tributary of the Lys River and finally of the Scheldt River. The study site is located in Villeneuve d’Ascq. Uptsream of Villeneuve d’Ascq the river passes through a sub-urban area whereas further downstream the river is heavily urbanized before flowing in the Deûle River. The annual average flow rate is 0.4 to 1 m3 s-1 and is highly affected by rain events. Sampling sites were selected in the river and in connected storm water ponds with regard to possible sources of pollution. The first sampling site is at the entrance of the Heron Lake (M1), an artificial storm water retention pond (with possible wastewater inputs) connected to the Marque River, collecting rainwater from parks, streets and the motorway to limit the impact of the urbanization in case of storm water runoff and also to partly manage the flow of the Marque River. The input of the lake to the Marque River represents on average 7% of the river flow. The second sampling site is a small storm water retention pond located adjacent to the highway and discharging in the Saint Jean Lake (M2). The third sampling station is located in the Marque River (M3), downstream of the WWTP of Villeneuve d’Ascq, which has a capacity of 170,000 inh. eq, including the nearby industries. This WWTP can have a significant impact on the water quality as its flow represents on average 23% of the river flow. During this sampling expedition, an interlaboratory comparison of PGEs using DGT was performed: each laboratory used its own prepared DGT units (described below) and applied its own analysis procedure. LASIRE performed grab samplings on 5 occasions during the two-week deployment (18/04/2019 – 02/05/2019) and daily amount of rainfall was recorded (Figure 7.2).

Figure 7. 2. Rainfall (mm) during sampling period of the Marque River in April-May 2019. Red bars indicate days when grab sampling was performed. Black line is the DGT deployment period.

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7.2.2. DGT functioning The device basically comprises a filter membrane and two gel layers (diffusive gel + binding agent), which are enclosed in a piston like plastic casing. The gel set is composed of a high- selective resin wrapped in gel (polyacrylamide hydrogel or agarose), which is separated from the environmental solution by a polyacrylamide or agarose diffusive gel and a 0.45-μm membrane filter designed to exclude the particulate elements. The flux of labile metal ions into the DGT device occurs - according to Fick’s first law - when a concentration gradient is established between the binding resin and the sample solution [31]. The mass of metal ions adsorbed (M) by the resin gel, after immersion of the DGT device for a time t in the water sample, is related to the free and labile analyte concentrations in the aqueous system (CDGT) as follows (equation 1.7.). The DBL (diffusive boundary layer) considered negligible because for most of the water flow conditions in rivers. DGT passive sampling allows to access a time- weighted average concentration (TWAC) of the labile dissolved concentration, while grab sampling provides an instant, and more fluctuating total dissolved concentration [32]. Moreover, as analyte species are continuously and selectively accumulated within the passive sampling device from the sample matrix and eluted into a small volume, the concentration in the measured extract will usually be higher than grab-sample concentrations, reducing the uncertainty of analytical determination and improving detection limits as well as reducing interferences in the analytical step due to matrix removal [33]. 7.2.3. DGT preparation, deployment, treatment and water sampling The DGT units for PGEs were prepared as described in Abdulbur-Alfakhoury et al. [221] for AMGC and for the Marque River also as described in Trommetter et al. [411] for LASIRE. The basic features of both DGT preparation, handling and analysis procedures are summarized in Table 7.1. Three chelating resins were used by the AMGC group (S914, S920 and MPX- 317) and one chelating resin (S920) and an anion exchange resin (AG MP-1) was used by the LASIRE group and presented in the manuscript. Additional tests using MPX-310 are shown in the Annex. In both laboratories, agarose was used as diffusive layer but with different thickness. The DGT devices were mounted on open Perspex plates by the AMGC group and in a high-density polyethylene (HDPE) cage by LASIRE. Different elution procedures were used: an aqua regia elution and a thiourea elution were used by AMGC and LASIRE, respectively. Elemental analyses were performed by ICP-SFMS and ICP-QMS with He collision cell by AMGC and LASIRE, respectively. Diffusion coefficients used in the calculations were determined experimentally by AMGC and LASIRE and are reported in Table 7.1.

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DGT deployment was conducted for 14 days, except in the hospital effluents, where a deployment time of 3 days was used due to the higher expected Pt concentrations. 4 replicates of each resin gel were used. Water temperature was recorded using a submersible probe (Aquatic 2, Tinytag). After deployment, the probes were thoroughly washed with ultrapure water. DGT field blanks (in triplicate) underwent all processes except for deployment.

Physicochemical parameters (pH, O2, conductivity) were recorded using a multiparameter probe (WTW GmbH, Germany). AMGC took grab samples were taken at the beginning and end of the deployment period. 10 mL samples were filtered on site using 0.45 μm disposable mixed cellulose ester syringe filters (Chromafil), stored in polypropylene (PP) tubes and acidified immediately on the site at 1% HCl and 2% HNO3. This acid mixture gave best results regarding stability of PGEs in solution and compatibility with the analytical procedure at AMGC. The grab samples by LASIRE were collected in 500 mL HDPE bottles, filtered on site using 0.45 µm cellulose acetate filters and stabilized with HCl (2%) and thiourea (6.5 mmol -1 L ) for PGEs and 2% HNO3 for major elements and trace metal analysis. Major elements, elements interfering in PGE analysis and tracers of specific sources (e.g. Gd for urban discharge because of its use in Magnetic Resonance Imaging) were analysed. Table 7. 1. DGT preparation, characteristics and handling procedures. AMGC LASIRE

Purolite S920, S914, Ionquest MPX- Purolite S920 (chelating resin), AG Resins used 317 (chelating resins) MP-1 (anion exchange resin)

Grinding S920, no treatment AG MP- Resin pretreatment Grinding, sieving on 50 µm 1

Resin gel Agarose + resin 0.5 mm Agarose + resin 0.75 mm Diffusive gel Agarose 0.75 mm Agarose 0.50 mm DGT holders Open plate In cage

DGT pretreatment prior to elution of 12 h deionised water PGE

1 mL 0.25 mol L-1 thiourea/0.25 mol L-1 HCl (AG MP-1); 1 mL 0.5 mL Elution 1 mL Aqua Regia 60°C 24 h thiourea/1 mol L-1 HCl (S920) 60 °C 48h

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S914: Pt: 0.91, Pd: 0.94, Rh: 0.99 S920: Pt: 0.81, Pd: 0.92, Rh: 0.95 AG MP-1: Pt: 0.98, Rh: 1.1 Elution factor MPX-317: Pt: 0.94, Pd: 0.87, Rh: S920: Pt: 1.0, Pd: 1.0, Rh: 1.0 0.90

Dilution factor 10 10 (AG MP-1), 20 (S920) Analysis ICP-SFMQ ICP-QMS

Diffusion coefficients S914: Pt: 5.19, Pd: 4.2, Rh: 2.7 AG MP1: Pt: 4.8, Rh: 5.4 used S920: Pt: 5.0, Pd: 4.2, Rh: 3.3 S920: Pt: 5.9, Pd: 5.4, Rh: 3.7 (10-6 cm2 s-1) MPX-317: Pt: 5.7, Pd: 2.8, Rh: 3.3

7.2.4. Sample analysis Trace metals analysis was performed using inductively coupled plasma sector field mass spectrometry (ICP-SF-MS, Element II, Thermo Fisher Scientific Bremen GmbH, Germany), at AMGC [26] and inductively coupled plasma quadrupole mass spectrometry (ICP-QMS, 7900, Agilent Technologies) at LASIRE [27].

7.3. Results and Discussion 7.3.1. Evaluation of DGT performance 7.3.1.1. Method detection limits (MDL) The resin gel blanks will determine the Method detection limits (MDL) and must therefore be kept as low as possible. The resins S920, S914 and MPX-317 require grinding (and sieving) prior to application in DGT. This may induce contamination and increase the blank level as well as variability between batches. Method detection limits (MDL) were calculated based on the standard deviations of the resin blanks and for a 14 days deployment period using equation 1.7. (Table 7.2). MDL are comparable between the two labs for Pt and Pd but are lower for Rh for LASIRE due to the softer elution method. The recently proposed improvement of the aqua regia elution, consisting of selective leaching of interfering elements prior to elution [356], was not yet applied during these experiments.

Table 7. 2. Method detection limits (MDL) for the resins used by the AMGC and LASIRE research groups. Abbreviation: N.D.: not determined.

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Pt Pd Rh ng L-1 ng L-1 ng L-1 S920 (AMGC) 0.007 0.05 0.010 S920 (LASIRE) 0.010 0.04 0.003 S914 (AMGC) 0.007 0.02 0.100 MPX-317 (AMGC) 0.010 0.2 0.050 AG MP-1 (LASIRE) 0.008 N.D. 0.0009

7.3.1.2. Influence of deployment time on the accumulation of Pt The influence of the deployment time (5-14 days) on DGT Pt concentrations has been studied in the Zenne River. The dissolved Pt concentrations varied between 3.5 and 3.8 ng L-1and no significant difference was found between exposure periods (Figure 7.7). The DGT calculated concentrations for the different resin gels and different exposure periods varied from 3.1 to 3.7 ng L-1: hence, no significant difference in accumulation rate was observed between the resin gels and between the exposure periods. A 14-day deployment period, which is the longest exposure period tested was selected for further deployments to accumulate a maximum amount of PGEs.

7.3.2. PGE concentrations in rivers 7.3.2.1. Marque River, France In the Marque River, Pt concentrations range from 0.2 to 0.6 ng L-1 (Figure 7.3). The highest concentration of Pt was found at the outflow of the WWTP at Villeneuve d’Ascq (M3). The lowest concentrations were found in the tributary draining the highway (M2) ; whereas intermediate values are found at the entrance of the Heron Lake (M1), which collects both rainwater and sewage water. This indicates that the main sources of Pt are the sewage effluents and that Pt derived from anticancer drugs are not fully trapped during the water treatment process. The Pt concentrations obtained by the three chelating resins by AMGC are in good agreement with each other, but are a factor 1.1 to 2 times higher than the results of LASIRE obtained by the chelating resin, depending on the sampling site. In addition, much lower results were obtained using the anion exchange resin AG MP1, indicating that not all Pt species are trapped on this resin. Best agreement between AMGC and LASIRE for S920 was found in Lake Saint Jean (M2), close to the highway. Pt emissions from catalytic converters will be in the form of nanoparticles (metallic or oxides), which slowly form dissolved complexes in natural waters whereas in the effluents of

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WWTPS Pt will be in the form of aquation and hydrolysis products of the cancerostatic drugs. Both the colloidal particles as the organoplatinum compounds can diffuse into the diffusive gel- and the uptake kinetics will depend on the affinity of the complexes towards the resins as well as the dissociation kinetics of the complexes [450-452]. The thicker diffusive gel used by AMGC may results in a larger fraction of Pt complexes dissociating during the time of diffusion. In addition, the effect of the thickness of the diffusive boundary layer (DBL) was not considered. A typical DBL of 0.2 mm, results in a 20% thicker diffusive domain in the DGT device of AMGC than that of LASIRE, which can result in systematic differences between the groups when the DBL has been neglected. The differences observed by LASIRE between the anion exchange resin and the chelating resin may indicate that the anion exchange resin has very little affinity towards the colloidal particles and thus maybe an interesting speciation tool.

Pt Pd Rh

0,80 1,40 0,015 1 - 0,60 0,90 0,01 0,40 0,40 0,005 concng L 0,20

0,00 0 -0,10 M1 M2 M3 M1 M2 M3 M1 M2 M3

Figure 7. 3. DGT PGE concentrations obtained in the Marque River using different resins: inter-laboratory study. Uncertainty bars of DGT concentrations represent standard deviations of 4 replicates.

Pd and Rh show a distribution similar to each other, but different from that of Pt, indicating different sources of the PGEs (catalytic converters vs. anticancer drugs). The concentrations of Pd and Rh (Figure 7.3) are highest in the lake draining the highway (M2) and the lowest near the WWTP (M3). Consequently, the input of PGE by road traffic and medical drugs results in a Pd/Pt ratio decreasing according the following order: M2 > M1 > M3, with logically M1 in between because it is a mix of M2 and M3 waters. In the emissions from catalytic converters Pd and Rh will also be present as nanoparticles, slowly forming soluble complexes in natural waters, with a higher solubility of the Pd compared to Rh [2, 11]. Quite large differences were observed between the two groups for Pd using the S920 resin In contrast to Pt, the best agreement was found in the Marque River (M3) and the lowest near the

190 highway (M2). For Rh no comparison could be made as the concentrations were below the detection limits of the AMGC method. Future inter-comparisons should include the exchange of extracts between the different labs to check instrumental differences, the deployment of the resins of the two labs in the same deployment setup, the use of DGTs with different diffusion gel thickness to evaluate the DBL, the use of double resin gels, etc. The chelating resins S920, S914, MPX 317 give comparable results but S920 has lower blank values offers the advantage that PGEs can also be extracted using acidic thiourea, making it the most promising resin for future studies. The amount of rainfall occurring during the sampling period is shown in Figure 7.2. Grab samples were taken on 5 occasions during the DGT deployment. Physicochemical parameters, major ions and a number of trace metals were analysed as well as direct measurements of total dissolved PGEs [411]. The concentration of Pt was lower than the quantification limit for direct analysis (0.7 ng L-1) at M1 and M2 and way around 1 ng L-1 at M3. Pd and Rh could not be quantified due to the high concentration of Sr (1 mg L-1) causing spectral interferences, which were not completely removed with the collision cell. The measured concentrations of Cu, Zn, Fe, Pb and Gd are shown in Table 7.3. A significant amount of rainfall was observed between April 23 and 29. The trace metal concentrations at M1 and M3 did not change significantly, but at M2 (entrance of Lake Saint Jean), collecting drainage water from the highway, a significant increase in traffic related elements such as Cu and Zn was observed [453]. The DGT concentration of Pd and Rh are highest at this station, which is the most affected by traffic. The DGT concentration of Pt is highest in the Marque River downstream of the WWTP (M3), and lowest in Lake Saint Jean (M2) and shows a similar trend as Gd. Surface water and, in some cases, groundwaters in industrialised and highly populated areas show REE patterns with outstanding, positive Gd anomalies [454]. The Gd anomaly results from the use of very stable Gd complexes (e.g. Gd- DOTA), used in magnetic resonance imaging in hospitals. After excretion from the body, these complexes pass through the sewage plants and are redistributed in surface waters. These Gd complexes are very stable under natural conditions and allow the study of the origin of certain waters and their fate during mixing processes.

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Table 7. 3. Concentrations of dissolved Cu, Zn, Pb, Fe and Gd in surface waters in the Marque River system. date Cu Zn Pb Fe Gd µg L-1 µg L-1 µg L-1 mg L-1 ng L-1 M1 (Lake Herron) 18/04 0.52 2.95 0.15 67 42 23/04 0.31 0.83 0.08 68 40 25/04 0.21 0.57 0.05 39 44 29/04 0.41 2.80 0.11 62 48 02/05 0.29 1.70 0.12 127 45 M2 (Lac Saint Jean) 18/04 0.86 13.0 0.48 47 9.3 23/04 0.63 8.60 0.19 52 8.6 25/04 5.87 29.7 0.71 199 13 29/04 3.25 39.5 0.41 80 7.3 02/05 1.10 8.90 0.17 133 14 M3 (Marque River) 18/04 0.22 13.4 0.11 91 360 23/04 0.29 12.0 0.13 120 340 25/04 0.20 11.8 0.14 134 330 29/04 0.46 17.0 0.17 146 240 02/05 0.33 17.8 0.12 150 300

7.3.2.2. Zenne River, Belgium The DGT concentrations of Pt, Pd and Rh obtained at the 3 stations (Z5, Z7 and Z9) for a 14- day deployment are shown in Figure 7.4. The concentrations of Pt increase from 3 ng L-1 at Z5 to 13 ng L-1 at Z9. A good agreement was observed between the different resin gels, as well as a good reproducibility between 6 replicates of the same resin gel (RSD <10%). The concentrations are also in good agreement with the total dissolved concentrations measured by direct ICP-MS measurement. Likewise, Pd concentrations increase from 2 ng/L at Z5 to 15 ng L-1 in Z9 and a good agreement was observed between the resin gels and between replicates (RSD < 10%) as well as with the total dissolved concentrations obtained by direct analysis and mathematical corrections for unresolved interferences [34]. The comparison between total dissolved Pt and DGT labile Pt in the Zenne River suggests that all Pt dissolved species are DGT labile. However, this was not observed by Trommetter et al [411] in effluents of WWTPs in France and their laboratory experiments have shown significantly lower diffusion coefficients for cancerostatic platinum compounds (CPCs) and spiked natural waters. Thus, this requires further investigation. Also, for Pd, in the Zenne River total dissolved Pd was in good agreement with the concentrations obtained by DGT but this could not be confirmed in the Marque River due the lower concentrations and high concentrations of interfering elements. DGT concentrations of Rh in the Zenne River range from 0.05 to 0.3 ng L-1 and total dissolved concentrations could not be quantified. It has however been shown that inert Rh compounds

192 can be formed due to ageing of solutions [455]. Thus, DGT may serve as a useful tool for risk assessment related to PGEs in urban rivers as has been performed for rare earth metals [456] .

Pd Rh Pt 20 20 0,6

1 15 - 15 0,4 10 10 0,2

5 5 concng L 0 0 0 Z5 Z7 Z9 Z5 Z7 Z9 Z5 Z7 Z9

Figure 7. 4. DGT concentrations of Pt, Pd and Rh obtained in the Zenne River in September 2018.

7.3.3. PGE concentrations in hospital effluents and in treated and untreated sewers 7.3.3.1.UZ Brussels hospital In order to investigate the importance of hospital effluents on the Pt concentrations in surface waters, a DGT deployment was performed in the main sewage drain of the UZ Brussels Hospital. The time averaged PGE concentrations in the UZ hospital effluent using different DGT types over 3 days deployment are displayed in Figure 7.5. For the resins S920, S914 and MPX-317 respectively, the obtained Rh concentrations (ng L-1) were (n = 4 replicates): 1.9 ± 1.0, 2.9 ± 0.5, <0.5; Pd concentrations (ng L-1) found were 10.3 ± 0.4, 12.9 ± 2.1 and 8.2 ± 3.3, whereas much higher concentrations were found for Pt (ng L-1): 163 ± 11, 143 ± 14 and 156 ± 11. The UZ hospital drain is receiving water directly from the hospital and runoff waters from surrounding areas such as parking cars zones. The very high concentration of Pt suggests direct input originating from Pt-based anti-cancer drugs. The concentrations in the hospital effluents are expected to show very high temporal variability, depending on the activities in the oncology department. In a UK hospital, Pt concentrations in the drain of the oncology department varied from 0.2 to 138 µg L-1 over a 11-day period [45].

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200 ngL -1 180 160 140 120 100 80 60 40 20

0

S914 S920 S914 S920 S914 S920

MPX-317 MPX-317 MPX-317 Rh Pd Pt

Figure 7. 5. DGT PGE concentrations in combined sewage drain of UZ Brussels Hospital, Jette using 3 different resins (S920, S914 and MPX317).

7.3.4. Concentration of dissolved Pt and Pd in untreated and treated wastewaters Dissolved Pt and Pd and particulate Pt (particulate Pd could not be quantified due to the important Sr interference) were determined monthly during one year at the WWTPs Brussels North and South. No seasonal trends could be observed. Yearly average concentrations of dissolved Pt and Pd in untreated and treated wastewaters of both Brussels WWTPs are presented in table 7.4. Both elements were poorly reduced in the dissolved phase after treatment: between 20 and 36% for Pd and between 16 and 38% for Pt, in North and South respectively. Considering the total Pt, the reduction after treatment is higher: 45 to 59% as the result of the efficient removal of particulate Pt. Note that there is no suspended matter in the clarified waters of the South station (particulate Pt concentration is almost zero) which is equipped with a nanofiltration system. This may explain the better efficiency in retaining Pd and Pt compared to the North station.

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Table 7. 4. Yearly average dissolved Pt (Pt-diss) and Pd (Pd-diss) and total Pt (Pt-tot) concentrations in the treated (outflow) and untreated (inflow) wastewaters of WWTP Brussels North and South, and % reduction after treatment (n=12). Pd-diss Pt-diss Pt-tot ng L-1 ng L-1 ng L-1 Untreated WWTP North 17.8 ± 3.2 13.8 ± 3.1 21.9 ±6.1 WWTP South 15.7 ± 6.0 14.0 ± 8.1 21.1 ±9.3

Treated WWTP North 14.0 ± 3.5 11.6 ± 2.2 11.8 ± 2.2 WWTP South 10.0 ± 4.8 8.7 ± 4.0 8.7 ± 4.0

%Reduction WWTP North 22±7 16±5 45±15 WWTP South 36±22 38±28 59±38

Between stations Z5 and Z9, the Zenne River receives the effluents of wastewater treatment plant Brussels North and minor inputs of untreated sewage. Knowing the river discharge at all Zenne stations (https://www.waterinfo.be/ and https://www.flowbru.be/fr), and the discharge of treated wastewaters at WWTP Brussels-North on the sampling day (AQUIRIS pers. Comm.) as well as the concentrations in all contributors (Table 7.5), it is possible to estimate daily loads of Pt and Pd at all Zenne stations and at the outlet of Brussels Nord WWTP as L = Q x C, with Q, daily water discharge in m³ day-1; C, concentration in mg L-1. Table 7. 5. Water discharge (Q), concentrations and daily loads (L) of dissolved Pt and Pd in Zenne stations Z5, Z7, Z9 and in the treated waters (outflow) of WWTP-North. Note that at Z7, the water discharge is not exactly known because of unknown contributions of untreated sewage, which should, however, be small (SBGE pers. Comm.). Q (m³ s-1) Pd (ng L-1) Pt (ng L-1) L_Pd (g d-1) L_Pt (g d-1) Z5 2.0 1.3 ± 0.4 2.9 ± 1.3 0.2 ± 0.1 0.5 ± 0.2 Z7 2.0+? 4.2 ± 1.4 7.5 ± 0.4 0.7 ± 0.2 1.3 ± 0.1 Z9 5.0 14.6 ± 3.4 13.4 ± 0.9 6.3 ± 1.4 5.8 ± 0.4 WWTP North 3.0 14.2 ± 3.6 11.6 ± 2.2 3.7 ± 0.9 3.0 ± 0.6

Results (Table 7.5 and Figure 7.6) show that the Pt and Pd loads increase between Z5 and Z7 by 0.8 and 0.5 g day-1 respectively, probably due to the contribution of untreated sewage release. Knowing the increase of load and the average concentration of untreated sewage (see table 6.4) we can calculate that the unknown discharge contribution of untreated sewage between Z4 and Z7 should be around 0.3 m³ s-1 which is a realistic estimate. Between Z7 and Z9, the loads of Pt and Pd again increase very strongly (Table 7.5 and Figure 7.8) by 4.5 and 5.4 g day-1, respectively. This increase can mainly be attributed to the release

195 of treated wastewaters by WWTP North. Indeed, the loads of WWTP North explain 67 % of the increase. Given the large uncertainties on load calculations, we can say that the mass balance is obtained (outlet at Z9 = inlet at Z7 + WWTP-North) for Pd and almost obtained for Pt (Figure 7.8).

Figure 7. 6. Cumulated Pt and Pd loads at station Z5 (Zenne River downstream WWTP Brussels South), at station Z7 (Zenne river upstream WWTP Brussels North) and at station Z9 (Zenne river downstream WWTP Brussels North). At Z9 the loads in the Zenne river are compared to the loads in the effluents of WWTP Brussels North

7.3.4.1.Contribution of hospital effluents to the Pt and Pd loads of Brussels untreated sewage Hospital effluents are released to the combined sewers of Brussels. They are particularly enriched in Pt (154 ng L-1) as evidenced above, and might therefore be significant contributors to the sewage loads carried by the city. Effluents at UZ-Brussels hospitals have a daily discharge of 260 m³ day-1 (0.003 m³ s-1) (Swartz, 2016), which can be used to compute daily Pt and Pd loads based on our concentration measurements. As no data is available on the use of anticancer drugs in the hospitals in Brussels, the number of beds was used. UZ Brussels has a total number of beds of 721 while all the hospitals in Brussels have a total of 8492 beds (data from 2016[457]). The loads of the 721 beds of UZ may thus be extrapolated to all the Brussels hospitals to compute a total load from hospital effluents which can then be compared to total loads in untreated sewage (WWTP-North and South). Untreated sewage Pt and Pd loads are computed from the average of measured concentrations and the discharge at WWTP Brussels- North and South on the sampling day (SBGE and AQUIRIS pers. Comm.).

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Table 7. 6. Water discharge (Q), number of hospital beds (bds), concentrations and loads (L) of Pt and Pd in the untreated sewage of WWTP Brussels North (WWTP-N), South (WWTP- S), and all Brussels WWTP; and in the effluents of hospitals UZ-Brussels (UZ) and all hospitals in Brussels. Contribution of Pt and Pd loads in all hospital effluents to their total loads in Brussels sewage (H/Swg).

Q (m³/s) beds Pd (ng L-1) Pt (ng L-1) L_Pd (g/d) L_Pt (g/d) Pd_H/Swg (%) Pt_H/Swg (%)

WWTP-N 3 - 17.8 ± 3.2 13.8 ± 3.1 4.6 ± 0.8 3.6 ± 0.8 - -

WWTP-S 0.5 - 15.7 ± 6.0 14.0 ± 8.1 0.7 ± 0.3 0.6 ± 0.3 - - all WWTP 3.5 - - - 5.3 ± 1.3 4.2 ± 0.9 - -

UZ 0.003 721 10.5 ± 2.4 154 ± 10 0.003 ± 0.001 0.04 ± 0.003 - - all Hospitals - 8492 - - 0.03 ± 0.01 0.47 ± 0.03 0.61 ± 0.17 11.3 ± 2.5

Results are summarized in Table 7.6. We can see that the contribution of hospitals to Pd loads in Brussels sewage is of minor importance (0.6 ± 0.2%) while for Pt is quite significant (11.3 ± 2.5%). Given that this Pt is only poorly retained in WWTP, we can conclude that hospital effluents strongly contribute to the Pt loads in the Zenne River. However, the main Pt inputs are expected to derive from home based anticancer treatments. 7.4. Conclusion The first field applications in urban rivers and hospital effluents show that DGT can successfully be used to measure time averaged concentrations of labile PGE species. Highest concentrations were observed in the highly impacted Zenne River. The selected sampling stations on the Marque River indicate that Pd and Rh are predominantly influenced by traffic inputs whereas Pt in urban rivers is mainly influenced by sewage inputs. Although a good agreement was found between total dissolved Pt and Pd and DGT obtained concentrations in the Zenne River, further validation is still needed for example to evaluate the DGT lability of cancerostatic platinum compounds and their degradation products and to explain the differences observed between the two laboratories. Deployment of DGTs in hospital effluents provides time averaged concentrations over several days, which are much more representative than grab samplings where high variability is expected. In small urban rivers, with high anthropogenic impacts, WWTPs are an important source of Pt to the rivers, as the retention of dissolved PGEs in the WWTPs is very low. In the Zenne River, Pt and Pd are show similar trends, conversely to the Marque River, where Pd and Rh show the same trends and differ from

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Pt. These elements, together with Gd, can thus be useful to trace specific pollution sources and their dispersion.

7.5. Acknowledgments ICP-AES and ICP-MS measurements at LASIRE (U-Lille) were performed on the Chevreul Institute Platform (U-Lille / CNRS). The Region Hauts de France and the French government are warmly acknowledged for the co-funding of these apparatus and for the founding of the CPERs Climibio and ECRIN. The French Water Agency Artois-Picardie is warmly thanked for cofounding the scientific project on the Selle River. The ICP-MS at AMGC (VUB) was funded by the Hercules project UABR/11/010. The authors thank Vincent Perrot for drawing the maps. 7.6. Supplementary Information

Pt

4,5 ng. L-1 4,0 3,5 3,0 2,5 2,0 1,5 1,0 0,5 0,0 S914 S920 MPX-317 0.45 µm C_DGT

5 days 7 days 14 days

Table 7. 7. Total dissolved Pt concentrations (0.45 µm filter) and DGT Pt concentrations with S914, S920 and MPX-317 binding phases for 3 deployment times (5, 7 and 14 days) at station Z5 in Zenne river in September 2018. The measurements named “0.45 µm” represent the direct total dissolved Pt determination. Uncertainty bars of DGT concentrations represent standard deviation of 4 replicates. Uncertainty bars of total dissolved concentrations (< 0.45 µm) represent standard deviations of duplicate measurements at the beginning and at the end of the DGT deployment.

198

199

Chapter 8: Conclusions and perspectives In this study the Diffusive Gradients in Thin-Films technique was successfully developed for Pt, Pd and Rh and applied in field studies in surface water of the Zenne and Marque rivers and in the UZ Brussels hospital effluent.

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8.1. General Conclusions

8.1.1. Method development The DGT method using five different binding phases, all chelating resins: Purolite S914, Purolite S920, Purolite S985, IONQUEST®MPX-317 and IONQUEST®MP-102, was investigated in detail under laboratory conditions. The linear relationship between mass of PGE accumulated on each adsorbent over the deployment time was positively verified. Rapid PGEs uptake on each binding layer was confirmed, thereby proving the validity of those binding gels as appropriate DGT adsorbents for PGE determination over time scale (up to 120h lab deployment in filtrated river or seawater). The influence of pH, ionic strength and organic matter on the DGT uptake was evaluated and diffusion coefficients were determined in spiked river water and seawater. All tested binding gel exhibited high selectivity towards PGEs and high binding capacity, but also accumulate other transition metals. The determining factors for the selection of those binding gels are the resin gels blanks and the accumulation of non-PGEs that may cause interfering effect on the ICPMS analysis of PGEs. All tested binding gels are used for industrial application and this causes the high resin gel blanks for some PGE and variability in blanc levels from batch to batch. Resin gels of MP-102 and S985 show higher binding for some undesired elements such Sr, Rb, Cu, Zn, etc. that limits their application in our study. The binding selectivity order of the binding phase is the imperative factor influencing the performance of the DGT method. The high selectivity of the resins towards PGEs result in a strong binding for the PGEs, which have positive and negative consequences. The positive consequence is the possibility of removing the other elements is possible using selective leaching. The negative consequence is the necessary of using strong acids such as aqua regia to obtain high recovery. The choice of most suitable resin for the different elements is thus based on the blank values of the resin gel, selectivity towards interfering elements and uptake kinetics. Hence, Pt- resin gel selection follows the order: MPX-317 > S920 ≥ S914> MP-102 > S985. Pd-resin gel performance follows the order S914 ≥ S920> MPX-317 > S985> MP-102 and the Rh- resin gel selection follows the order S920 > S914 > MPX-317 > S985 > MP-102. The high concentrations of Sr in aquatic samples forms the most important interference in the analysis of Rh and Pd. The resins S920, S914 and MPX-317 show very little affinity towards Sr and Sr accumulation in the resin gel can easily be removed by three consecutive Milli-Q washes. The resins Purolite S920, S914 and Italmatch Chemicals IONQUEST® MPX-317 were thus selected for further field studies.

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8.1.2. Selective removal of interfering elements from resin gels Further improvement of the method was made by selectively leaching the interfering elements from the resins without leaching the PGE. The success of this selective leaching was resin -1 dependant. As a common eluents for most interfering elements, 0.05 mol L H2SO4 could successfully be used to leach interferents from S920 and S914 without loss of PGEs but this is not the case for MPX-317, where a 15% loss of of PGEs was observed. For S920, over 50% of Cu, Mo and Zr can be removed, 70% of the accumulated Hf and over 95% of Zn, Cd, Pb and -1 Y. For the resin gel S914, the 0.05 mol L H2SO4 wash step is capable removing over 95% of Sr, Rb, Zn, Y and Mo and a 20-30% of Cu, Cd, Pb and Zr and Hf. The MPX-317 resin is particularly interesting for the determination of Pt when high Hf concentrations are expected as this resin accumulates 3-5 times less Hf compared to S920 and S914 and the residual Hf can be removed by a rinse with a salt solution such as NaNO2. The use of optimized rinsing procedures prior to the extraction of PGEs results in an improved reliability of DGT measurements of PGEs, especially in polluted aquatic systems with high concentrations of interfering elements.

8.1.3. Importance of ageing of solutions As aquation and hydrolysis reactions as well as complexation reactions with organic ligands will change the speciation of PGEs in solution in spiked natural waters and the kinetics of these reactions may be slow for Pt and Rh, the DGT technique was evaluated in aged solutions. This demonstrated however, that DGT inert Rh species Rh(OH)3 was formed in near neutral solution at the concentration level used in the experiments, resulting a lower apparent diffusion coefficient as the inert species were also included in the dissolved Rh measurements. Addition of organic matter to the aged solutions results in the solubilisation of these inert species. Thus, freshly spiked solutions need to be used to determine effective diffsuion coefficients of labile species. For Pt, no significant difference was found between freshly spiked and aged solutions indicating that all Pt species formed were DGT labile.

8.1.4. Field deployments and interlaboratory comparison The DGT field assessments exposed the importance of the comprehension of the binding affinity concept and its influence on the performance of the DGT technique under environmental conditions. This study demonstrated that in aquatic ecosystems, where the composition of the matrix is much more complex than in the laboratory deployment solution, as the PGEs in the field are in ultra-trace level (ppt) and the interferences are in ppb or ppm level. The different DGT non-PGEs specific binding affinity orders displayed by a different binding DGT layers is crucial for accurate PGEs determination by ICPMS. Therefore, the

202 knowledge of the affinity of each binding phase towards the analyte of interest and the identity of co-accumulating solutes is absolutely crucial in order to establish working parameters of the specific DGT method. The performance of the DGT technique with S914, S920 and MPX-317 binding layers only were evaluated in the environments of surface water influenced by the heavy traffic or direct effluent discharge of WWTPs and in a hospital effluent. As the concentrations of PGEs are often too low for direct measurements and the direct analysis of Pd and Rh is, we investigated preconcentration methods for water samples on resins. Both a cation exchange method, selectively removing Sr and other cations, followed by preconcentration by evaporation as well a preconcentration method using the chelating resins used in this study using either batch or column experiments was tested. Although satisfactory results were obtained for spiked solutions, due to lack of time, the method could not be sufficiently optimized to obtain reproducible results at low concentrations.

Up to now, only one other research group has focussed on the DGT method development for PGEs and during the course of our work, we performed the first interlaboratory comparison for these elements. Although there was a general good agreement between the two labs for Pt, differences were higher for Pd and concentrations of Rh were too low to allow comparison. These types of experiments are very important in DGT development. In contrast to the trace metals using Chelex binding resin, where the DGT preparation is well established and standardized, this is not the case for new developed elements, where both diffusive gel preparation and resin gel preparation are optimized by the different groups. Effect of other parameters, such as diffusion boundary layer, was not yet investigated.

This is the first study that has verified the feasibility of the DGT technique as a water monitoring and speciation tool for Pt, Pd and Rh determination in surface water and hospital effluent. The development and the application of the DGT approach provided more integrated view on the Pt, Pd and Rh speciation and allowed a better comprehension of the geochemical cycling and bioavailability of PGEs in the surface water and hospital effluents.

The first field applications of DGT in urban rivers and hospital effluents show that DGT can successfully be used to measure TWAC on labile PGE species. Although a good agreement was found between total dissolved Pt and Pd and DGT obtained concentrations in the Zenne River, further validation still needs to be performed to evaluate the DGT lability of CPCs and their degradation products and evaluate possible causes of differences observed between the two labs. Deployment of DGT in hospital effluents provides a TWAC over several days, which

203 is much is much more representative then grab samplings where high variability is expected. indicate that PGEs-First field results of PGEs by DGT. In small urban rivers, with high anthropogenic impacts, WWTPs are an important source of Pt to the rivers, as the retention of PGEs in the WWTPs are very low. In the Zenne River, Pt and Pd are correlated with each which was not the case in the sampling sites of the Marque River, where Pd and Rh were correlated as well as Pt and Gd. These elements can thus be useful tools as tracers of anthropogenic inputs and transport.

8.1.5. Overall evaluation of the resins used Based on the all the laboratory studies performed as well as the field work we can rank the resins in terms of suitable for simultaneous analyse of Pt, Pd and Rh in the order S920 > S914 > MPX 317. For S920, the lower blanc levels, possibility of removing transition metals using

0.05M H2SO4, possibility of using this resin water preconcentration methods as well as the possibility of leaching with PGEs with acidified thiourea instead of aqua regia are interesting properties. The binding strength of Pt and Pd on S914 and MPX 317 are higher, requiring the need of aqua regia digestion for quantitative recoveries. And these resins often show higher blanc values. If only Pt is to be measured, MPX 317 may be favourable due to the low accumulation of Hf on the resin.

8.2. Future perspectives

1. Using the technique in the work, a thorough study of the distribution, partitioning and speciation of PGEs in impacted rivers such as the Zenne River; including seasonal variations; influence of rainfall and sewer overflow events would be interesting to gain more insight on the biogeochemical cycle of PGEs. Partitioning between dissolved and particulate phases should be determined. In addition, the downstream transport, estuarine behaviour and input to the marine environment should be investigated.

2. The transformations of cytostatic platinum compounds released to surface water should be investigated as well as their binding on the DGT resins as there is still no clarity on their binding affinity for the resins.

3. Further intercomparison studies should be performed to investigate the reasons for the observed differences. This also includes evaluation of DBL, possible use of mixed binding gels, exchange of resins between the labs to investigate the analytical step, etc. The introduction of the standardized operating procedures, quality control, interlaboratory comparisons, and quality assurance is a

204 prerequisite to pave the way of the DGT technique into routine regulatory environmental monitoring programmes, and risk assessments, as described under the Water Framework Directive.

4. The importance of binding of PGEs to dissolved organic matter seawater and the lability of these complexes is unknown. Laboratory DGT deployments in seawater with and without UV irradiation may shed light on the DGT labile versus non labile fraction.

5. Besides Rh, Pd and Pt, it would also be interesting to study Ru and Ir. Many studies have mentioned about using the Ru as anticancer drug and some products are currently available but no study is available about their fate in the environment. Some other studies are reporting the use Ir in car catalysts. Trommetter et al [411] have already demonstrated that these elements can quantitatively be trapped on S920. Ru can be eluted by acidified thiourea but Ir would need an aqua regia digestion.

6. The resins used for PGEs also have a high affinity for Hg and other metals such as Cd and Pb. S924 (see annex) has successfully been sued for the multielement analysis of Cd, Pb and Hg in fish sauce Reichstadter el al, 2020 [217], The use of these resins for multielement analysis in environmental samples should further be explored.

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Annex I

Supplementary information for chapter 4: MPX-310 Uptake kinetics of PGEs by binding gels The uptake kinetics of Pt and Pd from old spiked 0.01M NaCl solution at pH 6.1± 0.2 on MPX-310. MPX-310 shows better performance for Pd uptake than MPX-317.

Figure A.I.1. Uptake kinetics of Pt and Pd on different resin gels and aged solutions (n= 3) at the same conditions (T = 24⁰C, pH = 6.1±0.2).

Resin gel blank The average DGT blank value based on 4 replicates measurement of is represented as (AVG (ng/ desk) + 3σ standard deviation of the handling blank; n = 4). MPX-310_blank_(ng)_n=4_ Rh(0.131±0.040) , 105Pd(0.125± 0.015); 106Pd(0.289± 0.0245); 108Pd(0.164± 0.083); Pt(0.115± 0.022). The obtained resin gel blank is similar to that of MPX-317 (Chapter 7. Section 7.3.1.1.). Effective diffusive coefficient The obtained effective diffusive coefficient for Pt, Pd and Rh using MPX-310 is similar to the reported one in our pervious study [221].

Figure A.I.2. The Diffusion coefficients (× 10−6 cm2 s−1, at 25 °C) with different binding gel obtained in time series experiments in filtrated seawater (FSW) (Rh= 2.9302±0.1920, Pd= 3.2304±0.933, Pt= 5.3902±0.601), and filtrated river water FRW (Rh= 3.052±0.573, Pd= 3.358±0.354, Pt= 5.3272±0.3629). The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution.

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Ageing effect The obtained effective diffusive coefficients for Pt and Rh using MPX-310 are similar to the reported one in our pervious study [455]. No salinity effect on the Pt measurement. The DOC causes the decrease in the Pt measurement.

Figure A.I.3. The Diffusion coefficients (× 10−6 cm2 s−1, at 25 °C) with different binding gel obtained in time series experiments in filtrated seawater (FSW) (Rh = 2.159±0.099, Pt=4.86±0.324), and filtrated river water FRW (Rh= 0.858±0.021, Pt= 4.223±0.506). The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution.

Pt_ Ageing effect C /C 1,50DGT sol 1,40 S920 1,30 S914 1,20 MPX 317 1,10 1,00 0,90 0,80 0,70 0,60 0,50 0,40 0,30 0,20 0,10 0,00 0.03 0.1 0.5 4 8 12

M (NaCl) mg/L (DOC)

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Rh_ aged solutions 0,50

0,40

0,30 S920 S914 0,20 MPX 317

0,10

0,00 0.03 0.1 0.5 4 8 12 M (NaCl) mg/L (DOC)

Figure A.I.4. Effect of ionic strength and DOC on the ratio of DGT Pt, Pd and Rh concentrations, CDGT, to their concentrations in the bulk solution, Csol Using MPX-310 binding gel in comparison with other resins. Mean values and standard deviation (error bars) of triplicate measurements are given (n = 3). "NaCl 0.03M, pH =6.1, T 19.6 ⁰C, Sal=1.2 ", "NaCl 0.1 M, pH=6.01, T 20.1 ⁰C, Sal=6.3" "NaCl 0.5M, T19.8 ⁰C; pH=6.06, Sal=33.5 ", "DOC4 mg L-1, pH=5.95, T 19.7 ⁰C, Sal=1.3"; “DOC 8mg L-1, T =19.5 ⁰C, pH=6.14, Sal=1.4”; " DOC 12 mg L-1, pH=6.07, T 19.2 ⁰C, Sal=1.6”. Pt DDGT for each binding gel is adopted from our previous work [44] taking into account temperature correction. Rh DDGT for each binding gel is adopted from table 7.2 the filtrated river water.

Field application The measured concentration during the field expedition is similar to other resin gels (Chapter 7)

Figure A.I.5. Field application of MPX-310 in Lille France together with the other resins.

Purolite S924 Resin gel blank The average DGT blank value based on 4 replicates measurement of is represented as (AVG (ng/ desk) + 3σ standard deviation of the handling blank; n = 4). S924_blank_(ng)_n=4_ Rh (0.0125±0.01), 105Pd (0.0140±0.0051); 106Pd (0.1725± 0.01439); 108Pd (0.0622± 0.00811); Pt (0.0059± 0.0009). The obtained resin gel blank is lower to that of S924 (Chapter 7. Section 7.3.1.1.).

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Effective diffusive coefficient The obtained effective diffusive coefficient for Pt and Rh using S924 is lower to the reported one in our pervious study [221], while Pd is similar to that obtained for other resins. Purolite S924 has high selectivity for Pd while it shows lower affinity for Rh or Pt.

Figure A.I.6. The Diffusion coefficients (× 10−6 cm2 s−1, at 25 °C) with different binding gel obtained in time series experiments in filtrated river water FRW (Rh= 0.451±0.046, Pd= 3.17 ± 0.44, Pt= 3.223±0.506). The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution.

MP-317 Resin gel blank The average DGT blank value based on 4 replicates measurement of is represented as (AVG (ng/ desk) + 3σ standard deviation of the handling blank; n = 4). MP-317_blank_(ng)_n=4_ Rh(0.965±0.028) , 105Pd(0.493± 0.011); 106Pd(0.54± 0.008); 108Pd(0.452± 0.067); Pt(0.241± 0.014). The obtained resin gel blanks are extremely for any field application and this resin is just rejected for its blank as first reason. Effective diffusive coefficient The obtained effective diffusive coefficient for Pt, Pd and Rh using MP-317 is lower to the reported one in our pervious study [221]. The accumulations of PGEs in function of the deployment time using MP-317 show non-linear behaviour while the resin gel made by using grounded MPX-317 shows linear accumulation at the same condition (see the article [221]). This type of non-linear behaviour could be explained reach either the capacity or poor selectivity. In this case and definitely, it is reaching the capacity. MPX-317 shows high selectivity for PGEs under similar conditions[221] hence the polymer of MP-317 must shows the similar selectivity. This is case is similar to use the micro chelex macro chelex [431] . MP-317 shows far low capacity comparing to MPX-317 and less than 50 ng of Pt, Pd or Rh is accumulated.

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Figure A.I.7. The Diffusion coefficients (× 10−7 cm2 s−1, at 25 °C) with different binding gel obtained in time series experiments in filtrated seawater (FSW) (Rh= 4.94±0.560, Pd= 3.37±0.382, Pt= 5.33±0.655). The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution. pH= 8.01; T=23 °C.

Conclusion for MP-317 Low accumulation and extremely high blanks so this resin could not be used for any field applications.

MP-101 Effective diffusive coefficient MP-101 shows non-linear accumulation in function of the deployment time for Pd and Pt. The obtained diffusive coefficient for Pd and Pt are lower than the reported for other resins (See the graph below). The non-linear behaviour might be competition effect (unlike the polymer of MP-317) s as this polymer is reported to have high affinity for all transition metals with strong binding for Cu [235]. Surprisingly this resin shows high affinity for Rh even at high salinity condition (filtrated seawater) and the obtained effective diffusive coefficient is similar to the reported for other binding gel [221].

Figure A.I.8. The Diffusion coefficients (× 10−6 cm2 s−1, at 25 °C) with different binding gel obtained in time series experiments in filtrated seawater (FSW) (Rh= 2.033±0.12, Pd= 1.320±0.021, Pt= 0.4766±0.054). The uncertainties of the diffusion coefficient values are combination of the uncertainties of the slope of the plots (95% confidence interval of regression line), the thickness of the diffusive gel and PGEs concentration of the exposure solution. pH= 8.01; T=23°C

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Resin gel blank The average DGT blank value based on 4 replicates measurement of is represented as (AVG (ng/ desk) + 3σ standard deviation of the handling blank; n = 4). MP-101_blank_(ng)_n=4_ Rh (0.215±0.0139), 105Pd (0.265± 0.015); 106Pd (0.023± 0.005); Pt (0.008± 0.00022). The obtained resin gel blank is similar to that of MPX-317[221].

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Curriculum Vitae: Ehab Abdulbur-Alfakhoury obtained a Bachelor Degree in Pure Chemistry at Al Baath University, Homs, Syria in 2007 and a Master Degree in Chemistry at Ghent University, Belgium in 2013 after working as a research assistant at the Atomic Energy Commission of Syria in Damascus from 2010- 2011. In 2015 he started his PhD at the Vrije Universiteit Brussel (VUB) under the supervision of Prof. Dr. Martine Leermakers. His PhD research focuses on the development of the passive sampling technique of diffusive gradients in thin-films (DGT) for platinum group elements, and it’s application in urban rivers. Ehab has presented the results from this work at several national and international scientific conferences. His work has led to the publication of four scientific articles in international peer-reviewed journals. He is currently working as a Technical Laboratory Specialist at the Department of Analytical, Environmental and Geochemistry of the VUB.

List of publications Abdulbur-Alfakhoury, Ehab, Steve Van Zutphen, and Martine Leermakers. "Development of the diffusive gradients in thin-films technique (DGT) for platinum (Pt), palladium (Pd), and rhodium (Rh) in natural waters." Talanta 203 (2019): 34-48. Abdulbur-Alfakhoury, Ehab, and Martine Leermakers. "Elimination of interferences in the determination of platinum, palladium and rhodium by diffusive gradients in thin-films (DGT) and inductively coupled plasma mass spectrometry (ICP MS) using selective elution." Talanta 223 (2020): 121771. Reichstädter, Marek, Pavel Divis, Abdulbur-Alfakhoury, Ehab, Yue Gao, “Simultaneous determination of mercury, cadmium and lead in fish sauce using Diffusive Gradients in Thin-Films technique." Talanta 217 (2020): 121059. Abdulbur-Alfakhoury, Ehab and M. Leermakers, “Evaluation of the effect of solution ageing on the DGT speciation of rhodium (Rh) and platinum (Pt)”. Journal of Analytical Atomic Spectrometry, 2021. DOI: 10.1039/D0JA00442A Abdulbur-Alfakhoury, Ehab Guillaume Trommeter, Natacha Brion, David Dumoulin, Marek Reichstädter, G. Billon, Martine Leermakers, W. Baeyens. “Distribution of platinum (Pt), palladium (Pd), and rhodium (Rh) in urban tributaries of the Scheldt River assessed by Diffusive Gradients in Thin-Films Technique (DGT)” (2021), Science of the Total Environment 784.

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