dissolution of cu-based engineered nanomaterials (enms) in
TRANSCRIPT
Dissolution of Cu-based Engineered Nanomaterials (ENMs) in agricultural soil :
Impacts on metal bioavailability and toxicity
Submitted in partial fulfillment of the requirements for the degree of
Doctor of Philosophy in
Department of Civil and Environmental Engineering
Xiaoyu Gao
B.S., Environmental Science, Nanjing University, China
M.S., Civil and Environmental Engineering, Carnegie Mellon University
Carnegie Mellon University Pittsburgh, PA
August 2019
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Acknowledgement Many people that have helped me during my Ph.D. studies in Civil and Environmental Engineering
Department of Carnegie Mellon University. I would like to express my thanks to five groups of
people, without whom this thesis would not have been possible: My advisors, my thesis committee
members, my collaborators, my lab members, funding agencies, and my family.
My Advisors
First, I would like to thank my two advisors, Dr. Gregory V. Lowry, and Dr. Elizabeth Casman. Dr.
Lowry is one of the reasons that I chose CMU as the school to do my graduate level studies. His
research in Environmental Nanotechnology attracted me here. I came to CMU as a Master student,
but I was able to do some interesting independent studies with Dr. Lowry. I had a great time in lab
and research group meeting, and I finally became a Ph.D. student after finishing my MS study.
During the 4 years of Ph.D. study, Dr. Lowry gave me tremendous help in my research. He guided
me to think independently and helped me fix the problem that I faced in my research. Dr. Elizabeth
Casman became my advisor after I passed the qualification exam. She provides tons of insightful
ideas regarding environmental policy and data analysis. Dr. Casman also provided me help on my
scientific writing. Writing ability was a problem when I entered the Ph.D. program because I am not
a native English speaker. With the help from Liz, now I am more confident about scientific writing.
Together with Dr. Lowry and Dr. Liz, we published five very good stories about the environmental
implications of ENMs, and a few more papers will be published soon. I feel truly grateful to be a
Ph.D. student of Dr. Lowry and Dr. Liz. Without their guidance, I could not finish all of the work.
Thesis Committee Members
Besides my advisor, I would like to thank the rest of my dissertation committee members (Dr. Sonia
M. Rodrigues and Dr. David A. Dzombak) for their support and invaluable advice. Dr. Sonia M.
Rodrigues is also our collaborator. She is an expert in soil chemistry and provided lots of valuable
advice with regards to the behavior and bioavailability of metals in the environment. Dr. David A.
Dzombak is our Department head, who taught me Environmental Chemistry during my 1st year in
CMU. His class was still one of my favorite classes, and he really inspired me to do research in this
field. His comments, like the redox reactions of Cu in soil, really helped me to make my studies
more well-rounded.
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My Lab members
I would like to thank my all lab mates, including Yilin Zhang, Garret Bland, Astrid Avellan, Jiang
Xu, Zimo Lou, Eleanor Spielman-Sun, Stephanie Laughton, Eric McGivney, Joe Moore, Rucha
Vaidya, and John Stegemeier. All of them have provided valuable feedback on my studies. Many of
them are very close collaborators for my studies.
I would also like to thank our lab managers, Ronald Ripper and Brian Belowich. My experiments
could not run successfully without their help. I can’t remember how many times that I was very
worried about my experiments and ask them for help. They really helped me so much, so that I
could successfully collect my experimental data.
CEE faculty and staff
I am also grateful to all the CEE staff. They helped me to make sure my Ph.D. studies can go
smoothly. They have also held lots of CEE events to make my Ph.D. life pleasurable.
Other faculty in CEE have also provided me help during my Ph.D. studies. I especially thank Dr.
Mitchell Small for helping me with the modeling and statistical tests in my study. I also thank Dr.
Kelvin Gregory and Dr. Robert Tilton for their comments in my research.
Funding Supports
I thank the NSF and EPA funding under NSF Cooperative Agreement EF-1266252, Center for the
Environmental Implications of NanoTechnology (CEINT), and CBET-1530563 (NanoFARM) for
their funding support.
I thank Carnegie Mellon University, College of Engineering for providing me College of
Engineering Dean’s Fellowship. I also thank Carnegie Mellon University’s Graduate Student
Assembly and Carnegie Mellon University’s Civil and Environmental Engineering Department
(under Professor Steven J. Fenves’s travel grant) for providing me travel funds.
My Family and Friends
Last but not least, I would like to express my deepest thanks to my family and friends. My mother
and my father supported me all the way to my Ph.D. studies. I feel grateful that they provided me a
chance to let me study abroad. I also thank my wife for staying with me and supporting me all the
time. This dissertation would not have been possible without their continued support.
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Table of Contents Acknowledgement .............................................................................................................................................. i
List of Figures ................................................................................................................................................... vii
List of tables ....................................................................................................................................................... x
ABSTRACT ....................................................................................................................................................... xi
CHAPTER 1: Introduction ......................................................................................................................... - 1 -
1.1 Introduction .................................................................................................................................. - 1 -
1.2 Objectives and overview of this thesis ............................................................................................ - 4 -
1.3 References of Chapter 1 .................................................................................................................... - 6 -
CHAPTER 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil ....................................................................................................................................- 10 -
Abstract: ...................................................................................................................................................- 10 -
2.1 Introduction ......................................................................................................................................- 11 -
2.2 Method and Materials ......................................................................................................................- 13 -
2.2.1 Chemicals ...................................................................................................................................- 13 -
2.2.2Nanoparticle Characterization. ................................................................................................- 13 -
2.2.3 Soils and Characterization of Soil Properties. ......................................................................- 14 -
2.2.4 Soil amendment and incubation. ............................................................................................- 14 -
2.2.5 Total Metal Concentration. .....................................................................................................- 15 -
2.2.6 Extractions to assess the labile Cu in soil samples. .............................................................- 15 -
2.2.7 Determination of Cu speciation in soils. ..............................................................................- 15 -
2.2.8 Dissolution kinetics. .................................................................................................................- 16 -
2.3 Results and Discussion ....................................................................................................................- 16 -
2.3.1 Soil and nanoparticle characterization. ..................................................................................- 16 -
2.3.2 The change of soil pH after amendment. .............................................................................- 17 -
2.3.3 General trends in extractable Cu for CuO NP- and Cu(NO3)2-amended soil. ...............- 18 -
2.3.4 Fractions of dissolved Cu and particulate Cu in extracts. ..................................................- 20 -
2.3.5 Effect of CuO NP concentration on its extractability in soil. ...........................................- 21 -
2.3.6 Dissolution rate of CuO NP in soil. ......................................................................................- 21 -
2.3.7 Effect of aging on speciation of Cu in Cu(NO3)2 and CuO NP amended soil. ..............- 22 -
2.4 Environmental Implications ...........................................................................................................- 24 -
2.5 References of Chapter 2 ..................................................................................................................- 26 -
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CHAPTER 3: Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution kinetics of CuO NP in soil..........................................................- 30 -
Abstract: ...................................................................................................................................................- 30 -
3.1 Introduction ......................................................................................................................................- 31 -
3.2 Method and Materials ......................................................................................................................- 33 -
3.2.1 Chemicals ...................................................................................................................................- 33 -
3.2.2Nanoparticle Characterization .................................................................................................- 34 -
3.2.3 Soil amendment ........................................................................................................................- 34 -
3.2.4 Extraction procedure to measure the fraction of dissolved CuO NP and soil pH. .......- 35 -
3.2.5 Determination of Cu speciation in soils ...............................................................................- 35 -
3.2.6 Dissolution models. .................................................................................................................- 35 -
3.3 Results and Discussion ....................................................................................................................- 38 -
3.3.1 Effect of Soil Organic Matter on dissolution of CuO NP in soil. ....................................- 38 -
3.3.2 Effect of soil pH on dissolution of CuO NP in soil. ..........................................................- 39 -
3.3.3 Effect of soil moisture content on the dissolution rate and solubility of CuO NP in soil. .. - 40 -
3.3.4 Dissolution rate and solubility of CuO NPs in soils with various properties. ................- 41 -
3.3.5 Predicting CuO NP solubility and dissolution rate in two test soils. ................................- 45 -
3.4 Environmental Implications ...........................................................................................................- 46 -
3.5 References of Chapter 3 ..................................................................................................................- 47 -
CHAPTER 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil. ..........................................................................................................................- 52 -
Abstract: ...................................................................................................................................................- 52 -
4.1 Introduction ......................................................................................................................................- 53 -
4.2 Method and Materials ......................................................................................................................- 55 -
4.2.1 Chemicals ...................................................................................................................................- 55 -
4.2.2Nanoparticle Characterization .................................................................................................- 56 -
4.2.3 Soils and Characterization of Soil Properties .......................................................................- 56 -
4.2.4 Soil amendment. .......................................................................................................................- 56 -
4.2.5 Germination and plant growth ...............................................................................................- 57 -
4.2.6 Sampling of soil and plant tissue ............................................................................................- 57 -
4.2.7 Soil extraction ...........................................................................................................................- 58 -
4.2.8 Cytoviva analysis .......................................................................................................................- 58 -
4.3 Results and Discussion ....................................................................................................................- 59 -
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4.3.1 Nanoparticle characterization .................................................................................................- 59 -
4.3.2 Change in extractability of Cu in bulk soil during the plant growth experiment. ............- 59 -
4.3.3 Toxicity of CuSO4 and CuO NP............................................................................................- 61 -
4.3.4 Cu root association and Cu uptake ........................................................................................- 62 -
4.3.5 Effect of near-root environment on Cu availability from CuO NP treatment ...............- 64 -
4.3.6 Soil pH in bulk soil, rhizosphere soil and loosely attached soil .........................................- 65 -
4.4 Discussion .........................................................................................................................................- 67 -
4.4.1 CuO NP dissolution is linked to toxicity. .............................................................................- 67 -
4.4.2 CaCl2 extractable Cu correlates with toxicity of CuO NP to wheat. ................................- 68 -
4.4.3 Root-associated CuO NP modulates toxicity. .....................................................................- 68 -
4.4.4 Root exudates affect CuO NP dissolution and availability. ...............................................- 68 -
4.4.5 Triticum aestivum regulated Cu uptake. ....................................................................................- 70 -
4.5 Agricultural implications .................................................................................................................- 70 -
4.6 References of Chapter 4 ..................................................................................................................- 71 -
CHAPTER 5: Dissolution functional assay improves understanding of metallic nanoparticle toxicity in agricultural soil ........................................................................................................................................- 76 -
Abstract ....................................................................................................................................................- 76 -
5.1 Introduction ......................................................................................................................................- 77 -
5.2 Methods .............................................................................................................................................- 77 -
5.2.1 Dissolution profile measurement assay .................................................................................- 77 -
5.2.2 Plant uptake measurement and toxicity ................................................................................- 78 -
5.3 Results and discussion: ....................................................................................................................- 78 -
5.3.1 Differences in dissolution time scale require different assays ...........................................- 78 -
5.3.2 Dissolution profile measurement assay predicted toxicity of Cu species to Triticum aestivum ..............................................................................................................................................................- 79 -
5.3 Environmental Implications ...........................................................................................................- 80 -
5.4 References for Chapter 5 ................................................................................................................- 81 -
CHAPTER 6: Summary of Major Contributions and Perspective on Future Research ..................- 83 -
6.1 Summary of Major Contribution ...................................................................................................- 83 -
6.1.1. Major Contribution from Objective 1: A test method to measure dissolution of CuO NP in soil was developed. ........................................................................................................................- 83 -
6.1.2. Major Contribution from Objective 2: A model was developed to evaluate the effect of soil pH and organic carbon content on dissolution kinetics of CuO NP in soil. ....................- 83 -
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6.1.3. Major Contributions from Objective 3:Dissolution of CuO NPs in soil was correlated with its toxicity to wheat (Triticum aestivum). Dissolution of CuO NPs under the influence of root activity in rhizosphere soil was quantified. ............................................................................- 84 -
6.1.4. Major Contribution from Objective 4: Dissolution kinetics functional assays were used to estimate exposure to ionic Cu from Cu-based ENMs in soil. This exposure correlated to observed toxicity in wheat. ...............................................................................................................- 84 -
6.2 Perspectives for future research .....................................................................................................- 85 -
6.2.1. Extension of the model to predict the behavior of other metal/metal oxide ENMs in soil ..............................................................................................................................................................- 85 -
6.2.2. Optimize the way to measure toxicity of metal/metal oxide ENMs in soil. ..................- 88 -
6.2.3. Design ENMs that can solve the micronutrient deficiency problem in calcareous soil. - 88 -
6.3 References for Chapter 6 ................................................................................................................- 89 -
Appendices ..............................................................................................................................................- 92 -
Appendix 1- Supporting information for Chapter 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil. .................................................................- 92 -
Appendix 2- Supporting information for chapter 3:Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution kinetics of CuO NP in soil. ......................................................................................................................................................... - 103 -
Appendix 3- Supporting information for Chapter 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil. ................................................ - 115 -
Appendix 4. Explanation on the solubility of CuO NP in Chapter 3 ......................................... - 125 -
Reference for Appendices: ................................................................................................................. - 126 -
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List of Figures Figure 2-1. Extractable Cu and in CuO NP and Cu(NO3)2 amended soils as a function of time and
the first order dissolution fit for CuO NP in soil. ---------------------------------------------------(19)
Figure 2-2. Fraction of small particles and dissolved ions (those passing 3kDa filter) in (a) DTPA
extracts and (b) CaCl2 extracts. -----------------------------------------------------------------------------(20)
Figure 2-3. Change of Cu speciation in amended soils as inferred by XANES.-------------------(23)
Figure 3-1. Schematic of CuO NP dissolution model. -------------------------------------------------(36)
Figure 3-2. Dissolution kinetics of CuO NP in Lufa 2.1 soil without added SOM (100 mg/kg dw
CuO NP treatment, circles) or with added SOM (300mg/kg dw CuO NP treatment, triangles). (39)
Figure 3-3. DTPA extractable Cu in Lufa 2.2 soil dosed with 500 mg/kg CuO NP at pH 5.9 and
pH 6.8. ------------------------------------------------------------------------------------------------------------(40)
Figure 3-4. Effect of moisture content on the dissolution kinetics of CuO NP in soil. -----------(41)
Figure 3-5. Correlation between organic carbon content and solubility (a) and between {H+}and
dissolution rate constant, kd (b). -------------------------------------------------------------------------------(44)
Figure 3-6. Prediction and experimental data of CuO NP dissolution in an Arizona soil (a) and in a
Portugal soil (b). --------------------------------------------------------------------------------------------------(46)
Figure 4-1. Change in DTPA extractable Cu over time for each treatment: a) CuO NP treatment, b)
CuSO4 treatment, and comparison of mean of extractable Cu for each Cu treatments at the end of
the plant growth period: c) DTPA extraction, d) CaCl2 extraction. ------------------------------------(60)
Figure 4-2. a) Root compactness and b) leaf length (leaf growth stage is noted with number, from 1
being the youngest to 3 the oldest) of wheat seedlings grown in freshly amended and aged CuO NP,
CuSO4-amended soil, and control treatments. -------------------------------------------------------------(62)
Figure 4-3. Hyperspectral imaging of plant roots grown in soil with freshly amended CuO NPs (a-e)
or after aging (f-i). ------------------------------------------------------------------------------------------------(63)
Figure 4-4. CaCl2 and DTPA extractable Cu in fresh (left side) and aged (right side) CuO NP
amended rhizosphere soil, loosely attached soil and bulk soils. -----------------------------------------(65)
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Figure 4-5. Mean ± SD of soil pH (measured using CaCl2 extraction) in rhizosphere soil, loosely
attached soil and bulk soil in a) soil freshly amended with CuO NP, b) aged CuO NP treatment c)
control soil, and; d) Comparison of pH of bulk soil among all treatments.--------------------------(66)
Figure 5-1. The dissolution profile of 250mg/kg of CuO NP, Cu(OH)2 NP and CuSO4 in Lufa 2.2
soil. ----------------------------------------------------------------------------------------------------------------(79)
Figure 5-2. Correlations between (a) Cu2+ integrated exposure and toxicity to Triticum aestivum and
(b) Cu2+ concentration at the end of exposure period and toxicity to Triticum aestivum ------------(80)
Figure A1-1. A) Primary particle size distribution determined from counting primary particles from
10 TEM imagines. B-K) Ten TEM images of CuO NP.--------------------------------------------------(95)
Figure A1-2. Size distribution of 80mg/kg CuO NP in pH=7, 5mM NaHCO3 buffer determined by
dynamic light scattering: (a) Number averaged size distribution, (b) intensity averaged size
distribution and (c) and autocorrelation function-----------------------------------------------------------(96)
Figure A1-3. Zeta potential of 80mg/kg CuO NP as a function of pH measured in (a) 5mM
NaHCO3 buffer and (b) 5mM NaNO3. Error bars indicate ± 1standard error. -----------------------(97)
Figure A1-4. X-ray diffraction spectrum of CuO NP. The CuO NPs used here are identified as
tenorite. ------------------------------------------------------------------------------------------------------------(98)
Figure A1-5. pH of CaCl2 extracts in different amended and blank soils-----------------------------(99)
Figure A1-6. Extractable Cu and in wet and air dried amended soils as a function of time: (a)DTPA
extraction for 10 mg/kg amendment, (b) CaCl2 extraction for 10 mg/kg amendment,(c) DTPA
extraction for 100 mg/kg amendment and (d) CaCl2 extraction for 100 mg/kg amendment. (101)
Figure A1-7. XANES spectra for model compounds----------------------------------------------------(102)
Figure A2-1. Cu EXAFS spectra (black) and linear combination fits (red) for CuO NP and CuSO4
exposed soil. (a) Arizona soil, (b) Lufa 2.2 soil. -----------------------------------------------------------(111)
Figure A2-2. DTPA extractable Cu in Lufa 2.1 soils dosed with 100mg/kg CuO NPs at pH 5.0
(squares) and pH 7.4 (triangles). Bars are standard deviation of the measurements. ----------------(112)
Figure A2-3: Cross validation of the correlation between kd and {H+}. -----------------------------(113)
Figure A3-1. Change in DTPA extractable Cu and CaCl2 extractable Cu for 250mg/kg CuO NP
treatment, 250mg/kg and 500mg/kg CuSO4 treatments (without growing plants) over 30 days. (116)
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Figure A3-2. Different soil regions defined in Chapter 4. ----------------------------------------------(117)
Figure A3-3. (A) Spectral library of the CuO-NPs. The spectral library has been built using
datacubes of CuO mixed with hydrated soil. (B) Example of SAM (Spectral Angle Mapping) results
to test for the specificity of the spectral library using positive controls (soil containing CuO NPs) or
negative controls (soil without CuO NPs or control root) images. The pixels containing the spectral
signal of CuO NP are highlighted in red (bottom line).---------------------------------------------------(120)
Figure A3-4. DTPA extractable Cu on bulk soil and bulk bottom soil in different Cu treatments
In all treatments, no significant differences (P<0.05, unpaired t-test) were found between DTPA
extractable Cu in bulk soil and bulk bottom soil, suggesting no vertical transport of Cu in all
treatments. ------------------------------------------------------------------------------------------------------(121)
Figure A3-5. Representative photos showing Cu toxicity led to shortened root and/or root
compactness in fresh CuSO4 treatment---------------------------------------------------------------------(123)
Figure A3-6. Hyperspectral imaging of plant roots grown in soil with CuO-NP, CuSO4 or Na2SO4
(control) freshly amended or after aging. Roots exposed to CuSO4 (both after soil aging or not)
showed a brown-damaged (necrotic) zone, that was not found on any of the CuO NP exposed
roots. -------------------------------------------------------------------------------------------------------------(124)
Figure A3-7. Mean concentration of Cu (mg/kg) in wheat tissue (dry weight): a) Cu concentration
in shoots, b) Cu concentration in roots. -------------------------------------------------------------------(125)
Figure A4-1. Conceptual of Cu speciation in Lufa 2.2 soil a) CuO NP treatment and b) CuSO4
treatment. -------------------------------------------------------------------------------------------------------(126)
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List of tables Table 2-1. Modeled first-order dissolution parameters for CuO NP amended soil. ---------------(22)
Table 3-1. Dissolution rate and solubility of CuO NP in a range of soils with various properties(42)
Table A1-1. Calibration ranges used for ICP-MS measurement used in Chapter 2 ------------------(93)
Table A1-2. Total Cu measured in amended soils (4 replicates) -----------------------------------------(94)
Table A1-3. Extractable Cu in unamended soil (3 replicates) ----------------------------------------------(94)
Table A1-4. The results of the LCF analysis of the X-Ray Absorption Near Edge Structure
(XANES) region of the samples------------------------------------------------------------------------------(103)
Table A2-1: Properties of sampled soils--------------------------------------------------------------------(106)
Table A2-2: Mass balance and experimental conditions for each treatment-------------------------(106)
Table A2-3. Comparison between CuO NP dissolution measured by XANES and chemical
extraction. --------------------------------------------------------------------------------------------------------(109)
Table A2-4. Linear combination fitting results of k3-weighted Cu EXAFS spectra (Figure S3-1) for
Arizona soil exposed to 300mg/kg of CuO NP or CuSO4. --------------------------------------------(110)
Table A2-5.Multivariate regression between dissolution rate constant and soil organic matter content and hydrogen ion activity. -------------------------------------------------------------------------(114)
Table A2-6.Multivariate regression between solubility and soil organic matter content and hydrogen ion activity ------------------------------------------------------------------------------------------------------(114)
Table A2-7.Multivariate regression between reprecipitation rate constant and soil organic matter content and hydrogen ion activity. --------------------------------------------------------------------------(114)
Table A3-1: Total Cu concentration (mean (SD), mg/kg) in soil for each treatment--------------(115)
Table A3-2: DTPA extractable Cu (mean (SD), mg/kg) in the control treatment before and after plant growth ----------------------------------------------------------------------------------------------------(115)
Table A3-3: Samples that provided sufficient soil for DTPA extraction for rhizosphere soil and loosely attached soil--------------------------------------------------------------------------------------------(122)
Table A3-4: Samples that provided sufficient soils for CaCl2 extraction for rhizosphere soil and loosely attached soil--------------------------------------------------------------------------------------------(122)
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ABSTRACT Metal and metal oxide (Me/MeO) engineered nanomaterials (ENMs) are being used in
agriculture as fertilizers and fungicides. A better understanding of how ENMs behave in agricultural
soil, interact with plants (through both soil application and foliar application), and become
bioavailable to plants can guide us to design safer, and more sustainable ENM enabled
agrochemicals. To that end, this thesis aims at better understanding the fate and behavior of ENMs
in agricultural soil and their bioavailability to plants.
The first objective of this work was to develop a method to measure the dissolution kinetics
of CuO NP in soil. Chemical extractions, CaCl2 and Diethylenetriamine Pentaacetic Acid extraction
over time have been developed to achieve this goal. This method was then applied in the second
objective to investigate how soil properties (pH, moisture content and organic carbon content)
influence the dissolution of CuO NPs in different standard soils and a natural soil. In this study, the
solubility of CuO NP was found to correlate well with soil organic matter content (R2 = 0.89),
independent of soil pH. In contrast, the dissolution rate constant of CuO NP in soil correlated with
soil pH for pH<6.3 (R2 = 0.89), independent of soil organic matter. Moisture content, on the other
hand, showed no impact on the dissolution kinetics of CuO NP in soil. These relationships
predicted the solubility and dissolution rate constants of CuO NP in two non-standard test soils
(pH=5 and pH=7.6).
The third objective was to investigate the bioavailability of CuO NP to plants, and their
effects compared to Cu salts. The third study quantified the influence of time and near-root
chemical conditions on dissolution of CuO NP to investigate influence of such dissolution on its
toxicity to Triticum aestivum. Readily available Cu (as reflected by CaCl2 extraction) increased in
rhizosphere soil, whereas the overall dissolution of CuO NP (as reflected by DTPA extraction)
decreased in rhizosphere soil. On the other hand, aging of CuO NPs increased the toxicity to
Triticum aestivum (reduction in root maximal length).This study stressed the importance of CuO NP
dissolution on its toxicity, and showed that plant-induced changes in rhizosphere conditions should
be considered when measuring the dissolution of CuO NP near roots.
The fourth objective was to develop a functional assay that used CuP NP dissolution
kinetics to estimate the Cu2+ ion exposure, and to determine if this assay can predict toxicity to
plants and to soil organisms. For different Cu based ENMs, dissolved Cu was plotted over time to
get the dissolution profile. Different Cu-based ENM had distinct dissolution rates. The integrated
xii
Cu2+ exposure (the area under the dissolution curve in the dissolution profile) was correlated with
selected biological endpoints. The integrated exposure of CuO NP correlated well with its toxicity to
Triticum aestivum. This study suggested that the dissolution profile of Cu-based ENMs may be a
better measure of exposure to Cu ion than a single measurement of extractable Cu2+ ion at the end
of an experiment.
Overall, meeting these objectives provided fundamental knowledge on how different ENMs
behave in soils, and how they interact with plants. It also provides a test method to measure the
dissolution rate of Cu-based ENMs directly in soils. These methods and knowledge will provide
guidance on the design and application of nano-enabled agrochemicals for improving the
sustainability of agriculture, especially in marginal soils where agriculture productivity is typically
low.
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CHAPTER 1: Introduction 1.1 Introduction Metal and metal oxide (Me/MeO) engineered nanomaterials (ENMs) are being used in fertilizers
(e.g. ZnDDP ® or CuDDP®) and fungicides (e.g. Kocide ® 3000) in agriculture1–3. The small size of
nanoparticles (NPs) facilitates their uptake by plants through leaves and roots4,5. Also, they can be
tuned to slowly release active ingredients over time or to target the relevant plant tissue6,7 by
modifying their surface coatings. The overuse of non-nano metal-containing agrochemicals has
resulted in toxicity to plants8, microbial communities9 and rhizosphere and soil invertebrates10. If
properly tuned, nanoparticles could deliver metals more efficiently to plants than non-nano
agrichemicals, reducing the attendant environmental contamination. But, to fulfil the promise of
nano-enabled agrichemicals, we need to understand their behavior in soil, particularly the factors
that influence their dissolution, because information on ENM dissolution in soil will be necessary
for developing guidelines on the appropriate dosage of nano-enabled agrichemicals and for
predicting their impact on human health and the environment.
Previous studies have investigated the toxicity of metal and metal and metal oxide (Me/MeO)
ENMs in aqueous systems. These studies attributed the toxicity of Ag NPs11–13, ZnO NPs14 and
CuO NPs15,16 to the ions released upon their dissolution. By separating the dissolved ions from the
NPs using either filtration or centrifugation methods, previous studies have shown that the
dissolution of Me/MeO ENMs in aqueous systems generally follows a first-order process17–19.
It is more difficult to measure the dissolution of Me/MeO ENMs in soil. In soils, metal ions
released from Me/MeO ENMs associate with soil solid surfaces, e.g. soil organic matter (SOM) and
clay20,21. In soil, the released metal ions need to be chemically extracted from soil in order to measure
the dissolution kinetics.
The lack of characterization of dissolution behavior of Me/MeO ENMs in soil has resulted in
contradictory conclusions on whether the toxicity of Me/MeO ENMs in soil is due to metal ions or
due to a particle-specific effect. Some studies attributed the toxic effect of Me/MeO ENMs in soil
to ion release (dissolution) 22–24, while others concluded the opposite25,26. A few examples that
illustrates the problem follow.
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Servin et al. assumed that 10% of the CuO NP would dissolve in soil, the same fraction that
dissolved in pure sand, and used that estimate as the Cu2+ concentration of their ionic control, rather
than measuring actual CuO dissolution in soil. They concluded that dissolution of CuO NPs could
not fully explain the plant toxicity because the plant responses differed from those of their Cu2+ ion
control .25 Much more than 10% CuO NP could have dissolved in soil, compared to sand, because
SOM is a Cu sink that increases the amount of CuO NP that can be dissolved27. Not measuring
dissolution in soil undermined their conclusions about a NP-specific effect. Similar problems
occurred in other studies 22,23,28–30.
Dimkpa et al. (2013) evaluated the total CuO NP dissolved in soil using a water-extraction method. 31 The water-extraction method does not extract Cu bound to the soil solid matrix (which includes
SOM), the location of >90% of the dissolved Cu in soils32–34. Because they underestimated the
actual dissolved Cu in the soil, their conclusion of CuO NP-specific toxicity in soil was false. (There
are other problems with the water extraction method that are discussed later, in Chapter 4.)
Qiu et al. observed that the toxicity of CuO NP, CuO bulk particles and soluble Cu (Cu(AC)2)
depends on their solubility in soil, and that the distinction between NP and bulk particles diminished
after a 90-day aging period. However, the dissolution profile over time (a graph of changing metal
ion concentration over time during an exposure experiment) was not provided (they measured
dissolution only on day 0 and day 90). They successfully correlated the toxicity of CuO NP to roots
of Hordeum vulgare L. (in a 5-day root elongation experiment) with free Cu ions in soil pore water
measured at a single time point before seeding.24 This is a major step in the right direction, but an
even better correlation could have been obtained if dissolution were measured over the 5-day
exposure period. Had the dissolution profile been known, the ionic control could have reflected the
change in Cu2+ over time. This would be necessary to determine whether a NP specific effect
existed. If the dissolution of metal or metal oxide ENMs in soil is not tracked, the attribution of an
observed toxic effect to the NPs or the ion-release process is usually inconclusive.
Chemical extraction methods have been developed to measure the speciation and bioavailability of
metals in soil. The chemical extraction methods can be generally classified as two types: (1) the pore
water extraction method, and (2) the labile metal extraction method. The first method extracts free
mobile metal ions in soil pore water. Such species are usually considered to be readily available to
plants. This method uses a dilute salt as extractant, such as 0.01M calcium chloride (CaCl2) or
0.005M calcium nitrate (Ca(NO3)2). Such extractions predict the bioavailability of metals by
- 3 -
mimicking the chemistry of soil pore water (e.g. similar ionic strength) and targeting the
exchangeable metal ions in soil pore water 35–37. The second method uses strong chelators, such as
0.05M diethylenetriamine pentaacetic acid (DTPA) or 0.05M ethylenediaminetetraacetic acid
(EDTA) as an extracting agent. Those strong chelating agents mimic the chelating effect of root
exudates which enhance the availability of nutrients from the soil rhizosphere for plant uptake38.
This extraction method targets the total “labile” metal in soil that not only includes the free metal
ions in soil pore water, but also the metal ions associated with SOM. In this thesis I show how these
two extraction methods, with modifications, can be used to measure the released ions from ENMs
directly in soil. The behavior of metal ions in soil is a rapid partitioning process between soil pore
water and soil solid surfaces.21 In contrast, NPs slowly releases metal ions into the soil porewater39.
Essentially, this makes the bioavailability of Me/MeO ENMs time-dependent. Thus, to capture the
dissolution kinetics of Me/MeO NPs in soil, chemical extractions need to be performed repeatedly
over time.
When measuring the dissolution of ENMs in soil, the time scale of the dissolution process could
also be system-dependent. In aqueous systems, studies have shown that the main factors affecting
the dissolution kinetics of Me/MeO ENMs include pH,17,40,41 organic matter (OM) content1,42,43,
oxygen concentration44 (if the dissolution is a oxidative dissolution, e.g. AgNPs) and complexing
ions (e.g. Cl-)45. The dissolution of Me/MeO NPs in soil systems is expected to be affected by the
soil properties, e.g. soil pore water pH and organic matter content. These parameters affect either
the ion release rate or the partitioning of the metal ions with soil solid surfaces. It is the partitioning
in this multiphase system (soil, water, ENM) that makes the time scale of dissolution in soil different
from that in water.
The time scale of dissolution and the time scale of exposure are both important when trying to
correlate the dissolution of ENMs in environmental media to biological endpoints
(toxicity/bioavailability). In aqueous systems, the time scale of dissolution is usually much shorter
than the time scale of exposure. A measurement of (equilibrium) solubility of the Me/MeO ENMs
in the relevant aqueous medium should correlate with NP toxicity24,46. However, in soil systems,
where the dissolution kinetics are more complex 47,48, a measurement of the overall dissolution
profile is likely needed to establish the connection between dissolution of ENMs and their biological
endpoints.
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In several regions around the world, soils are deficient in copper, especially regions with calcareous
soils49. There, additional Cu is added to soil to supply the micronutrient for optimal plant growth.
Copper-based ENMs can be applied to soil as a ‘slow-release’ source to deliver Cu to plants. The
Cu-based ENMs have a higher efficiency and lower risk compared to Cu ion based fertilizers.27,50,51.
The other reason that Cu-based ENMs are applied to agricultural soil is due to their anti-microbial
properties52, e.g. Kocide ® 3000 (whose main component is Cu(OH)2 NPs) is a registered fungicide
in U.S. agriculture. Although the agricultural application of Cu-based fertilizers and pesticides are
regulated (maximal application rate is 75 kg ha-1 yr-1) , the regulations were based on toxicity of
soluble species such as CuSO4 or Cu(NO3)227. However, the fate and behavior of Cu based ENMs is
different from the Cu soluble species27,39 leading to potentially different bioavailability and toxicity to
plants and soil organisms compared to more soluble species. Questions remain about how to
accurately measure the toxicity of Cu-based ENMs in soil and about how they should be regulated.
Thus, Cu-based ENMs, especially CuO NP was selected as the main Me/MeO ENMs to investigate
in this thesis. Specifically, in this thesis, dissolution of Cu-based ENMs, which release the active
ingredient- Cu2+, is considered the main transformation process that affect its bioavailability in soil.
Other transformation processes, e.g. redox reactions, were not considered. For aerated agricultural
soil (topsoil to which agrochemicals are applied), the redox potential is usually above 400mv53. At
such a redox condition, Cu(II) is the major Cu valence state54.
1.2 Objectives and overview of this thesis
In order to tackle the challenge of understanding and quantifying metallic ENMs' behavior in
aerobic soil and plant systems, four objectives were pursued, each comprising a separate chapter of
this thesis. Objective 1 was to develop a method for quantifying the dissolution kinetics of CuO
NPs in soil. Objectives 2 and 3 used the new dissolution measurement assay to investigate 1) how
soil properties affect the dissolution behavior of CuO NPs in soil (Objective 2) and how plant
exudates affect the dissolution of CuO NP in soil, and how dissolution of CuO NP affects the plant
health (Objective 3). Objective 4 was to determine the best assay of Cu-based ENM dissolution in
soil to predict their toxicity to plants and soil isopods. This knowledge will provide guidance on the
design and application of nano-enabled agrochemicals for improving the sustainability of agriculture,
especially in marginal soils where agriculture productivity is typically low.
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Objective 1: Develop a functional assay to measure the dissolution kinetics of metal-based
nanoparticles in soil.
I developed a protocol for a time-dependent series of extractions to determine the CuO NP
dissolution rate constant and reveal the dissolution profile in soils. This work has been published in
Environmental Science & Technology. 27
Objective 2: Quantify the effect of soil properties, including soil moisture content, organic
carbon content and pH, on the dissolution kinetics of CuO NP in soil.
Chemical extractions were applied to measure CuO NP dissolution kinetics in soils with different
properties, providing data for a model to predict the dissolution kinetics of CuO NPs in soil. The
model successfully predicted the dissolution kinetics of CuO NPs in two unknown soils. This study
showed that soil pH and organic matter content affect the dissolution behavior of CuO NP in soil in
a predictable manner. This work has been published in Environmental Science & Technology55.
Objective 3: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum
aestivum) in rhizosphere and bulk soil.
To quantify the influence of time and near-root chemical conditions on dissolution and lability of
CuO NPs in rhizosphere soil, and to determine the influence of this dissolution on the toxicity of
CuO NPs to plants, we measured the rate of dissolution of CuO NPs in bulk soil, and in soil in
which wheat plants (Triticum aestivum) were grown. At the end of the plant growth period (14 days),
available Cu was measured in three different soil compartments: bulk (not associated with roots),
loosely attached to roots, and rhizosphere (soil firmly attached to roots). Root length shoot length
and biomass were also measured as indicators of toxicity. This study correlated CuO NP dissolution
and the resulting Cu ion exposure profile to phytotoxicity and showed that plant-induced changes in
rhizosphere conditions are the most important determinants of ENM toxicity to roots. This work
has been published in Environmental Science & Technology47.
Objective 4: Use the dissolution kinetics functional assays to predict Cu-based ENMs
toxicity in agricultural soil.
Dissolution profiles of various Cu-based ENMs in Lufa 2.2 soil were measured. Toxicity of different
Cu species to Triticum aestivum (as evidenced by shortened leaves length and root length, reduction in
biomass) was measured. The preliminary analysis of this approach suggested that dissolution profiles
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of different Cu species could be predictors of the different biological endpoints than the status quo
of measuring extractable Cu only at the end of the exposure period.
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CHAPTER 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil
Abstract: The effect of CuO nanoparticle (NP) concentration and soil aging time on the extractability of Cu
from a standard aerobic sandy soil (Lufa 2.1) was investigated. The soil was dosed with CuO NP or
Cu(NO3)2 at 10 mg Cu kg-1 soil (mg/kg) or 100 mg/kg total copper, then extracted using either
0.01M CaCl2 or 0.005M DTPA (pH 7.6) extraction fluids at selected times over 31 days. For 100
mg/kg CuO NP, the amount of DTPA-extractable Cu in soil increased from 3 wt% immediately
after mixing to 38 wt% after 31 days. In contrast, the extractability of Cu(NO3)2 was highest initially,
decreasing with time. The increase in extractability was attributed to CuO NP dissolution in soil.
This was confirmed with synchrotron X-ray absorption near edge structure (XANES)
measurements. The CuO NP dissolution kinetics were modeled by a first-order dissolution model.
Our findings indicate that dissolution, concentration, and aging time are important factors
influencing Cu extractability in CuO NP-amended soil, and suggest that a time dependent series of
extractions could be developed as a functional assay to determine the dissolution rate constant.
This work has been published in Environmental Science & Technology as ‘Time and Nanoparticle Concentration Affect the Extractability of Cu from CuO NP-Amended Soil’ , doi: acs.est.6b04705
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2.1 Introduction Copper based nanoparticles (NP) including metallic copper (Cu NP), copper oxides (Cu2O NP and
CuO NP), and copper hydroxides (Cu(OH)2 NP) are manufactured nanomaterials that have been used
as pesticides and fungicides because of their antimicrobial properties1, 2 .They can also be used as
fertilizers to deliver micronutrient-Cu to plants, which can improve fertilizer efficiency and crop
yield3,4. Copper salt (mainly as Cu(NO3)2 or CuSO4) based micronutrients and pesticides have
historically been widely used. Excessive use of Cu containing fertilizers and pesticides may lead to
negative impacts on ecosystems, soil microorganisms, microbial processes5, plants6 and soil
invertebrates7.
In the U.S., Cu containing fertilizers and pesticides are regulated, with the maximum
application rate of 75 kg/ha/year (USEPA, 1993). However, these regulations were determined
using highly soluble Cu salts (e.g. Cu(NO3)2 and CuSO4) in soil. Dynamic processes including
aggregation, oxidation, and dissolution will likely make the available pool of Cu derived from Cu
based NP time-dependent8, 9. While the importance of time on the fate and bioavailability of Cu salts
is documented10-12, aging effects for Cu based NP has not been elucidated. In order to assess the
impact of Cu based NP to agroecosystems, it is important to determine the factors controlling their
bioavailability in soils.
Chemical extraction methods are used to predict the bioavailability of metal in soil13. Several
single extraction methods, originally developed to determine the fraction of metals in soil involved
in geochemical equilibrium processes including sorption and precipitation, can predict the leaching
of soil metals to groundwater, their impact on ecosystems, and their bioavailability for soil organisms
or plants11-22. Two extraction methods, 0.01M CaCl2 extraction and 0.005M
diethylenetriaminepentaacetic acid (DTPA) extraction (pH 7.3~7.6) are commonly used for
predicting the bioavailability or lability of metals such as Cu, Zn and Cd, in soil13-20. CaCl2 extraction
(0.01M) predicts metal bioavailability by mimicking the chemistry of soil pore water and targets the
exchangeable metal ions in soil pore water which are ‘readily available’ to plants in soils14, 20, 21. DTPA
is a strong chelating agent that mimics the chelating effect of root exudates to enhance the nutrient
availability from soil for subsequent uptake 15. The DTPA extraction not only targets the free ions in
soil pore water, but also the carbonate-bound and the organic-bound fractions of metal in soil,
which could be ‘potentially available’ to plants14, 22. While these extraction methods for assessing the
lability of Cu in Cu salt (CuSO4 and Cu(NO3)2) amended soil or for metal contaminated soils are
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well-developed, there are only a few reports using such methods with Cu-based NP or other
metal/metal oxide nanoparticles in soil23-25.
Recently, a few studies have used single time point CaCl2 extraction and DTPA extractions
to predict the lability of metal/metal oxide nanoparticles in soil. Judy et al.25 used CaCl2 and DTPA
extractions to estimate the bioavailability of ZnO-NP, TiO2-NP and Ag-NP in soil and concluded
that these extraction methods could not predict their bioavailability to plants (Medicago truncatula) in
Woburn sandy soil. Pradas del Real et al.23 used DTPA and CaCl2 extractions to assess the labile
pool of Ag in Ag NP amended soil, and concluded that the low extractability of Ag in soil was
consistent with the low bioavailability of Ag to plants (wheat and rape) in a loamy soil. Xu et al.24
used CaCl2, EDTA and DTPA extractions to estimate the bioavailability of CuO NP and TiO2 NP
to soil microbes and their community structures in a typical paddy soil. They observed that DTPA
and EDTA extractable Cu in CuO NP amended soil correlated well with microbial activity
(microbial biomass, soil enzyme activity, and total phospholipid fatty acids) in a CuO NP amended
soil. So far, results from studies on the use of chemical extraction methods to predict the
bioavailability of metals from nanoparticles in soils are contradictory and often inconclusive. One
reason for this may be the fact that these studies did not assess the rates of transformations of NP in
those soils and the corresponding effect on metal extractability. We hypothesize that aging time and
concentration will be important factors influencing these particles’ transformation and bioavailability
in soil, which may explain the absence of a correlation between extractability and bioavailability
using a single time-point extractions23-25.
The dissolution and transformation of some metal and metal oxide NP in soil have been
determined. The dissolution of copper oxide nanoparticles over time in three soils was reported by
McShane et al26. In their study, they measured an increase in free Cu2+ activity in soil pore water over
time and concluded that CuO NPs were dissolving. However, the rate of dissolution of CuO NP
was not modeled or reported. The present study extends this work by McShane et al. by measuring
pore water and SOM-associated Cu species using well-established extraction methods designed to
assess bioavailable fractions of Cu, by synchrotron X-ray analysis to confirm changes in copper
speciation, and by determining the effect of NP concentration on the dissolution behavior. The
transformations of metal and metal oxide nanoparticles in soil have been monitored using
synchrotron X-ray absorption spectroscopy (XAS)27-32 to measure changes in metal speciation over
time. Recently, Sekine et al.27 used XAS to monitor the change of speciation of Ag-NP, AgCl-NP
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and Ag2S-NP in soil over time. They observed that an increase in S-bound Ag species, including
Ag2S-NP, Ag-cysteine and Ag-cysteine, correlates with the decrease in labile Ag determined using
diffusive gradients thin films. However, the Ag NP transformation kinetics were not studied.
The dissolution of a number of metal and metal oxide nanoparticles in water has been
reported33-39. Most studies use empirical first-order dissolution models to describe their dissolution 35-
38, and evidence suggests that the measured dissolution rate constants are concentration dependent36.
However, the dissolution rate of metal and metal oxide nanoparticles in soils, where water content
and SOM can greatly affect the dissolution, is less well-understood. These rates are needed to
understand the dynamic nature of nanoparticulate metals relative to soluble metals added to soils
and to parameterize fate and transport models for engineered nanomaterials. 40
The objectives of the present study were to (a) compare the extractability of CuO NP with
the extractability of Cu(NO3)2 in soil, (b) quantify the extractability of CuO NP as a function of time
and nanoparticle concentration in a sandy (Lufa 2.1) soil (c) determine the fate processes influencing
the extractability of CuO NP in soil and (d) to model the dissolution kinetics of CuO NP in soil
from extraction experiments. We used 0.01 M CaCl2 and 0.005M DTPA (pH=7.6) extraction
methods to study the extractability of Cu(NO3)2 and CuO NP in aerated soils over a one-month
period at two different total added Cu concentrations (10 and 100 mg Cu kg-1 (mg/kg) dried soil).
Changes in speciation of Cu in soil were monitored using XAS to infer the dissolution of CuO NP.
2.2 Method and Materials
2.2.1 Chemicals
CuO NP (50 nm), DTPA, (>99% (titration)) and triethanolamine (TEA, ≥99.0% (GC)) were
purchased from Sigma-Aldrich. Cu(NO3)2 (>98% ACS grade), calcium chloride (≥99.0%, (ACS
grade)) and sodium bicarbonate (≥99.7%, (ACS grade)) were purchased from Fisher Scientific.
2.2.2Nanoparticle Characterization.
Primary particle size distribution of the CuO NP was characterized by transmission electron
microscopy (TEM, Hitachi H-9000 TEM microscope operating at 300 kV). The hydrodynamic
diameter and zeta potential of CuO NP in suspension (80 mg/kg as Cu in 5mM pH=7 NaHCO3
buffer) were determined by dynamic light scattering (Zetasizer Nano, Malvern). The isoelectric
points of 80mM CuO NP in 5mM NaHCO3 buffer and in 5mM NaNO3 were calculated from
- 14 -
measurements of the zeta potential of the particles in suspension over a range of pH. The crystal
structure of CuO NP was determined by X-ray powder diffraction (XRD, Panalytical X’Pert Pro
MPD X-Ray Diffractometer).
2.2.3 Soils and Characterization of Soil Properties. Standard soil (2.1-sandy soil) was purchased from Lufa, Germany. The standard soil (Lufa 2.1) was
used because it is commonly used in bioavailability studies and therefore can enhance comparison
from different studies. Lufa 2.1 soil also contains very little extractable Cu and total Cu (as discussed
later in ‘Soil and nanoparticle characterization’ section), making background interference minimal.
Lufa soil was air dried and sieved < 2mm before shipping. The soil was further air-dried for 12
hours before all experiments. Soil pH was determined according to the standard procedures
recommended by the USDA41. Specifically, 5 g of air-dried soil was mixed by hand for 10s with 5ml
of deionized water. The pH of the solution was measured after allowing the mixture to settle for 10
minutes. To determine the soil moisture content, 2 g of the air dried soil were dried in an oven at
105 ºC for 24 h 42. The moisture content was then determined gravimetrically. Soil field moisture
capacity was determined using a modified cylinder method in which air-dried soil was added to a
15ml-graduated cylinder. Deionized water was then added into the cylinder to wet the top 2 cm of
soil. After 24h, the wetting front in the soil moved downward. After removing the top 2 cm of soil,
the moisture content of soil above the wetting front (which was assumed to be at soil’s field
capacity) was determined.
2.2.4 Soil amendment and incubation. Two doses of CuO NP and Cu(NO3)2 were used in our study: 10mg Cu/kg dry soil for the low dose
amendment, and 100 Cu mg/kg dry soil for the high dose amendment. These two doses were
selected to investigate the influence of concentration on extractability of CuO NP in soil. While the
low does is more realistic, the high dose provided sufficient Cu concentration for XAS study. Soils
were amended with CuO NPs or Cu(NO3)2. All amended soil samples were incubated in 50ml
centrifuge tubes under aerobic conditions between 0 and 31 days before being extracted and
digested. Holes were made in the caps of centrifuge tubes for air exchange. This aerobic condition
was chosen to prevent CuO NP from being reduced to Cu(I) species or Cu(0). Cu speciation
analysis from XAS confirmed the absence of significant amounts of reduced Cu species in the
samples (<a few wt%) by. The experiments described in the following chapters of this thesis applied
the same incubation conditions for ENM soil amendments. Additional details of the amendment
procedure can be found in Appendix 1.
- 15 -
2.2.5 Total Metal Concentration. Soil total metal concentration was determined using acid digestion according to USEPA Method
3050B (1996). According to the procedure, 1g of air-dried soil was digested with concentrated nitric
acid and 30% hydrogen peroxide at 95 ºC using a hot plate. After digestion, the samples were
centrifuged at 3000 rpm for 10 min, followed by filtration using 0.45um filter to remove fine
particles in the supernatant. The filtered supernatant was diluted with Milli-Q water and acidified
with 20% HNO3 (final HNO3 concentration was 2%) for analysis by ICP-MS (Agilent 7700x). The
instrument was calibrated with a mixed calibration standard (purchased from Agilent Technologies)
every time before measurement. The calibration ranges used for different samples can be found in
table A1-1 (Appendix 1).
2.2.6 Extractions to assess the labile Cu in soil samples. After different incubation periods, 2.0 g of air-dried soils or 2.3 g of wet soils were extracted with
two standard extractants: The first one (termed DTPA) uses a 4 mL mixture of 0.01M CaCl2,
0.005M DTPA and 0.1M triethanolamine (TEA) (pH=7.6). The second one (termed CaCl2) (pH=5)
uses 20 mL of 0.01M CaCl2. All extractions were done using a reciprocal shaker at 180 rpm for 2
hours. Sample bottles were laid horizontally in the shaker. Both wet soil and air dried soil were used
to study the effect of air drying. After extraction, all samples were centrifuged at 3000 rpm for 10
min, and the supernatants were filtered with using a 0.2 um PTFE filter. In order to monitor the
impact of CuO NP suspension or Cu(NO3)2 solution on pH of soil, the pH of CaCl2 extracts for air-
dried amended soil and a unamended soil (no nanoparticle or Cu(NO3)2 added) were also measured
to estimate the soil pore water pH. The samples collected were further filtered with a 3kda filter to
separate the dissolved and nanoparticulate fraction of Cu in extracts. All samples were acidified with
20% HNO3 (final HNO3 concentration was 2%) and Milli-Q-water and analyzed by ICP-MS. Due to
the large difference between Cu concentrations from CaCl2 extracts and Cu concentration from
DTPA extracts, different calibration ranges were used. The different calibration ranges used for
different samples can be found in Table A1-1 in Appendix 1.
2.2.7 Determination of Cu speciation in soils. Cu speciation in soils (Lufa 2.1) on 1, 4, 7 and 19 days after amendment was analyzed by Cu K-edge
XAS at the Stanford Synchrotron Radiation Lightsource (SSRL) on Beamline 11-2. Spectra for both
100mg/kg and 10mg/kg amended soils were collected. However, the signal-to-noise ratio for the
10mg/kg amended soils was too poor for adequate speciation. Specifically, samples were lyophilized,
- 16 -
ground with a mortar and pestle to achieve uniformity, pressed into pellets, and placed between
Kapton tape. A double crystal Si (220) monochromator was calibrated by setting the first inflection
of the K-edge of a metallic Cu foil to 8979 eV. Harmonic rejection was achieved by detuning the
monochromator crystal by 25%. Spectra of soil samples were recorded in fluorescence mode at
room temperature using a 100-element germanium detector. The scans were averaged, energy
corrected using a metallic Cu foil standard, deadtime-corrected, background subtracted with E0
defined at 8988 eV, and de-glitched using SIXPack data analysis software43. Spectra were analyzed by
linear combination fitting (LCF) using the following reference spectra: CuO NP, metallic Cu,
CuSO4, Cu(NO3)2, CuPO4, Cu-cysteine, Cu2S (chalcocite mineral sample) ,CuS (covellite mineral
sample), Cu- iron oxide, Cu+ sorbed to humic acid (Cu(I)-HA) and Cu2+ sorbed to humic acid
(Cu(II)-HA). Inclusion of a reference spectrum into the combination fit required at least a 10%
decrease in the R-value, indicating a significant change to the quality of the fit.
2.2.8 Dissolution kinetics. For both extraction methods, the extractable Cu (either in pore water (CaCl2), or pore water plus soil
bound Cu (DTPA)) is assumed to increase proportionally as the CuO NPs dissolve.
The increase in the extractability of Cu over time is modeled using equation 2-1,
𝑑𝑑𝑑𝑑𝑑𝑑𝑑𝑑
= 𝑘𝑘�𝐸𝐸𝑓𝑓𝑓𝑓𝑓𝑓𝑓𝑓𝑓𝑓 − 𝐸𝐸� (2-1)
where E is the concentration of extractable Cu at time t, k is the empirical 1st order extraction rate
constant, and Efinal is the concentration of extractable Cu at the end of experiment. If the dissolution
of the CuO NP is the rate limiting step, i.e. the Cu-soil organic matter interaction is much faster
than the dissolution of CuO NP in soil, then the measured extraction rate constants from both
extractions should be similar, and equal to the CuO NP dissolution rate constant.
2.3 Results and Discussion 2.3.1 Soil and nanoparticle characterization. Lufa 2.1 soil is a sandy soil, containing 3 wt% clay, 11 wt% silt and 86 wt% sand (as provided by
Lufa). It has low organic matter content (organic carbon content is 0.7 wt% as provided by Lufa).
After air-drying, Lufa soil had 1.2 wt% moisture content. The soil pH was 5.6 and the field capacity
- 17 -
was 16 wt%. The total Cu concentration of the unamended soil was 2.95±0.11mg/kg. Total Cu
concentration measured in each of the amended soils is presented in Table A1-2 (Appendix). The
DTPA extractable Cu in unamended soils ranged from 0.37 to 0.53 mg/kg dried soil while the CaCl2
extractable Cu in unamended soils ranged from 0.005 to 0.024mg/kg (Table A1-3 in Appendix).
The primary particle size of CuO NP (measured from TEM) was 38nm (s.d. =14nm, 278 particles
were counted). The hydrodynamic diameter and zeta potential of 80mg/kg CuO NP in pH=7, 5mM
NaHCO3 buffer were 557nm (s.d. =56nm, 3 replicates, polydispersity index<0.297) and -16.1mv
(s.d. =0.8mV, 3 replicates), respectively. TEM images of the particles, along with the number of the
primary particles used to determine the size distribution are provided in the supporting information
(Figure S1-1, Appendix 1). Hydrodynamic size distribution (intensity averaged and number
averaged) and autocorrelation functions of CuO NP are also provided in supporting information
(Figure S1-2, Appendix 1). The isoelectric point of CuO NP shifted from pH=5.8 (in 5mM
NaHCO3 buffer) to pH=8.8 (in 5mM NaNO3), indicating a specific interaction between carbonate
and/or bicarbonate with the CuO NPs. The zeta potential of CuO NP measured over a range of pH
in NaHCO3 buffer and NaNO3 is provided in the supporting information (Figure A1-3 in Appendix
1). XRD results (Figure A1-4 in Appendix 1) indicate that the CuO NP we used is tenorite.
2.3.2 The change of soil pH after amendment. The pH of amended soils and unamended soils were stable over time. The pH of soil pore water
(measured in CaCl2 extracts) ranged from 4.9 to 5.1, except for the high dose CuO NP amended
soil, whose pH ranged from 5.2 to 5.5. The relatively higher pH in the high dose CuO NP amended
soils may be due to the acid-promoted dissolution of the CuO NP (eqn. 2-2 and 2-3). This
dissolution is described in more detail later in the manuscript.
CuO(s) + H2O(l) ↔ Cu(OH)2 (s) (2-2)
Cu(OH)2 (s)+2H+(aq) ↔ Cu2+
(aq) +2H2O(l) (2-3)
In our system, the pH rose from 5.1 to 5.4. This increase is less than expected for consumption of
40 mg Cu/kg soil of according to eqn 2-3, suggesting that the soil’s buffering capacity limited the
- 18 -
increase in soil pH44. The pH measured in the CaCl2 extracts can be found in supporting information
(Figure A1-5 in Appendix).
2.3.3 General trends in extractable Cu for CuO NP- and Cu(NO3)2-amended soil. The CaCl2 (0.01M) and DTPA extractable (0.005M, pH=7.6) Cu for CuO NP and Cu(NO3)2
amended soils over time are shown in Figure 2-1. Both the low dose (10 mg/kg) and high dose
(100mg/kg) scenarios are included for comparison. For both the low and high doses of added Cu,
there are clear differences in the trends of extractable Cu for CuO NP compared to Cu(NO3)2.
Initially, the extractable Cu for Cu(NO3)2 amended soils was higher than for the CuO NP amended
soil for both CaCl2 and DTPA extractions. Over time, the Cu(NO3)2 amended soils showed a
decrease in extractable Cu, whereas the CuO NP amended soils showed an increase in extractable
Cu with time. This is consistent with the findings of McShane et al.26 who found that ionic Cu in
pore water increased with time for CuO NP addition, but decreased with time for the Cu(NO3)2
addition. The trend of lower extractability of Cu in Cu(NO3)2 amended soil over time is well
documented; the lower extractability over time is a result of micro pore diffusion and the
complexation of ionic Cu by SOM10-12 as well as possible irreversible binding between ionic Cu and
SOM 45. The increasing extractability of Cu from CuO NP with time suggests that the NPs were
transforming to become more extractable. This is a result of dissolution (as suggested by McShane et
al.26), which was confirmed in this study with XAS, as discussed later in the paper. Note that the low
extractable Cu in CuO NP amended soil on day 0 (immediately after amendment) indicates that the
extractions did not induce nanoparticle dissolution.
For all extractions and time points, the DTPA extractable Cu was higher than CaCl2 extractable Cu
for both CuO NP amended soil and Cu(NO3)2 amended soil. This is consistent with former studies 17, 24, 25. This is because the CaCl2 extraction only extracts the dissolved metal and small particulate
metal in pore water, while DTPA extracts both the metal in pore water and the carbonate mineral-
bound and organic-bound metal15, 20-22. Importantly, this result suggests that dissolution of CuO NPs
is followed by an interaction between released copper and soil organic matter, which is known to
affect the amount of ionic Cu in soil. In the high dose addition, there was a relatively rapid change
occurring in the first ~7 days, followed by a period of slower change as the system approached an
apparent steady state with respect to CuO(s) dissolution. In the low dose addition, the extractability
of Cu for the CuO NP addition is the same as that for Cu(NO3)2 for t > 10 d (p > 0.05,
Kolmogorov-Smirnov test). This suggests that the Cu may be fully dissolved and “aging” similarly
- 19 -
to the Cu(NO3)2. However, the slight downward trend in extractability for t > 10d is not statistically
significant (P > 0.05, one-way ANOVA test).
Although extraction procedures generally used air dried soils21,46 , we used both the air dried soils
(after incubation) and wet soils for extractions to investigate the influence of air drying on
extractability of CuO NP in soil. Our results indicated that air drying has no significant effect
(P>0.05, Kolmogorov-Smirnov test) on extractability of Cu in both CuO NP amended and
Cu(NO3)2 amended soils. Thus, only the results from air dried soil is shown in Figure 2-1 for clarity.
Additional discussion on the effect of air drying can be found in the supporting information.
0 1 0 2 0 3 0 4 00
2
4
6
8
1 0
I n c u b a t i o n t i m e ( d a y s )
C u ( N O 3 ) 2
C u O N P
( a )
Ex
tra
cta
ble
Cu
(mg
/k
g d
rie
d s
oil
)
1 0 m g / k g , D T P A e x t r a c t i o n
0 1 0 2 0 3 0 4 00 . 0
0 . 1
0 . 2
0 . 3
I n c u b a t i o n t i m e ( d a y s )
C u ( N O 3 ) 2
C u O N P
( b )E
xtr
ac
tab
le C
u
(mg
/k
g d
rie
d s
oil
)
1 0 m g / k g , C a C l 2 e x t r a c t i o n
0 1 0 2 0 3 0 4 00
2 5
5 0
7 5
1 0 0
I n c u b a t i o n t i m e ( d a y s )
C u ( N O 3 ) 2
C u O N P
( c )
Ex
tra
cta
ble
Cu
(mg
/k
g d
rie
d s
oil
)
1 0 0 m g / k g D T P A e x t r a c t i o n
0 1 0 2 0 3 0 4 00
4
8
1 2
I n c u b a t i o n t i m e ( d a y s )
C u ( N O 3 ) 2
C u O N P
( d )
Ex
tra
cta
bl e
Cu
( mg
/k
g d
r ie
d s
oi l
)
1 0 0 m g / k g , C a C l 2 e x t r a c t i o n
Figure 2-1. Extractable Cu and in CuO NP and Cu(NO3)2 amended soils as a function of time and
the first order dissolution fit for CuO NP in soil: (a) DTPA extraction for 10 mg/kg amendment, (b)
CaCl2 extraction for 10 mg/kg amendment, (c) DTPA extraction for 100 mg/kg amendment and (d)
CaCl2 extraction for 100 mg/kg amendment. Error bars indicate ± 1 standard error. Dashed lines
indicate model fits using equation 1. For the low dose amendment, because CuO NPs were fully
dissolved after the 7-day sampling time, we modeled only the first 7 days. represents extractable
- 20 -
Cu in CuO NP amended soils air dried after incubation and represents extractable Cu in
Cu(NO3)2 amended soils air dried after incubation.
Figure 2-2. Fraction of small particles and dissolved ions (those passing 3kDa filter) in (a) DTPA
extracts and (b) CaCl2 extracts. D1, D2, D31 stand for 1 day, 2 days and 31 days after dosing. Error
bars indicate ± 1 standard error.
2.3.4 Fractions of dissolved Cu and particulate Cu in extracts. Bioavailability of Cu depends on its speciation, e.g. free ions, complexed ions and particulate
species47. We used filtration (first a 0.2-micron filter followed with a 3kDa filter) to distinguish
between dissolved and particulate species of Cu in each of the extracts. Figure 2-2 shows the
fraction of Cu that passes the 3 kDa filter (considered dissolved) in CaCl2 and DTPA extracts. For
DTPA extraction, nearly all extractable Cu (from 90% to 100%) was dissolved. This is because most
Cu in DTPA extracts bound with the chelating agent (DTPA) and the Cu-DTPA complex can pass
through the 3kDa filter. In contrast, filtration of the CaCl2 extract indicated the presence of Cu-
containing particles compared to the DTPA extracts (P<0.05, Kolmogorov-Smirnov test). These
small particles may include Cu2+ ion complexed with SOM or potentially small CuO NP (in CaCl2
extracts for CuO NP amended soil). The species of Cu in CaCl2 extracts was not analyzed in this
study. However, there were no effects of concentration, type of Cu added, or time on the amount
of dissolved vs. particulate Cu (p>0.05, one way ANOVA test).
- 21 -
2.3.5 Effect of CuO NP concentration on its extractability in soil. The concentration of added Cu influences the extraction behavior for CuO NP compared to
Cu(NO3)2. For the low Cu dose, the extractability of Cu in CuO NP amended soil was the same as
for the Cu(NO3)2 amended soil after ~10 days. No statistically significant difference (p>0.05,
Kolmogorov-Smirnov test) is found for extractable Cu for both CaCl2 extractions and DTPA
extractions between CuO NP amended soil and Cu(NO3)2 amended soil on day 13, 19 and 31,
suggesting that the CuO NP were fully dissolved before 13 days in soil at the lower dose. The
behavior was quite different at the high Cu dose. For the high dose of added Cu, extractable Cu in
Cu(NO3)2 amended soils was always higher than the extractable Cu in CuO NP amended soil. The
extractability of Cu from the CuO NP amended soil increased over the entire 31day period,
suggesting that CuO NP was dissolving over 31 days, but the dissolution of CuO NP in soil was not
complete. One possible explanation on the persistence of CuO NP and the slower dissolution rate
after ~7 days in the high dose soil (100mg/kg Cu) is that the free Cu2+ in soil pore water approached
saturation with respect to CuO(s). Conversely, the lower dose system (10 mg/kg) was not
oversaturated with respect to the CuO(s) phase. While CaCl2 extraction is a well-established method
to assess the pore water concentration of dissolved Cu, the potential for artefact during the
extraction and uncertainty in the complexation constants for Cu and the NOM in our system
prevents an accurate determination of the degree of saturation in the pore water. .
2.3.6 Dissolution rate of CuO NP in soil. For the high dose of CuO NP (100 mg/kg), the first-order extraction model describes the change of
extractable Cu over time well (R2>0.995) (dashed lines in Figure 2-1). However, we should note that
Cu2+ ions dissolved from CuO NP can become irreversibly bound with soil organic matter, making
it unextractable by DTPA, as indicated in former sections. This irreversible interaction is about 20%
for our soils, and has a minimal effect on the calculated dissolution rate constant. This is in part
because it is a small fraction of the total, and in part because the time scale for partitioning into this
irreversible fraction is short, i.e. less than 1d compared to the dissolution processes being
investigated, i.e. many days to weeks. For CaCl2 extraction, the fraction of extractable ionic Cu was
significantly less, with only 2% to 10% of the ionic Cu being extractable because it targeted only Cu
in soil pore water. Despite the differences in the extractable amount of Cu, the modeled dissolution
rate constants for DTPA extractable Cu and CaCl2 extractable Cu are similar (Table 2-1). This
indicates that the extractable amount of Cu by either the DTPA or CaCl2 extraction can be used to
monitor the CuO NP dissolution in the soils. This is a natural consequence of a first-order
- 22 -
dissolution process, which scale with the ratio of the final and initial concentration (C/Co) so any
process that reduced C and Co by the same constant fraction will not affect the calculated rate.
Moreover, it suggests that Cu2+ binding to SOM is rapid enough, such that dissolution of the CuO
NP is the rate-limiting process controlling both DTPA extractable Cu and CaCl2 extractable Cu in
soil.
2.3.7 Effect of aging on speciation of Cu in Cu(NO3)2 and CuO NP amended soil. Speciation of Cu in the 100mg/kg CuO NP and 100mg/kg Cu(NO3)2 amended soils were
determined at selected time points using XANES (Figure 2-3). Details regarding the spectra for
model compounds and fitting result can be found in the supporting information (Figure A1-7 and
Table A1-4 in Appendix 1). The speciation of Cu in Cu(NO3)2-amended soils can be adequately
modeled using only the Cu(II)-HA model compound, indicating that the Cu has predominantly Cu-
O character, i.e. associated with humic acids or potentially (but less likely) with clay or metal oxide
surfaces of the solids. This is consistent with prior speciation studies indicating that the main species
of Cu in soil is Cu(II)-HA using experimental approachs48-49 and with results of equilibrium
partitioning modeling50. This also suggests that Lufa 2.1 soil has the capacity to sorb up to
100mg/kg of added Cu, because our data showed that all Cu in the 100mg/kg Cu(NO3)2 amended
soil was Cu(II)-HA. In contrast, the Cu speciation in CuO NP amended soil required both Cu(II)-
HA and CuO NP model compounds. In the high dose CuO NP amended soil, linear combination
fitting indicates that the presence of CuO decreases over time, with a subsequent increase in the
Cu(II)-HA. This suggests that the CuO NPs were dissolving relatively fast in the first 7 days and
then more slowly after that as the pore water becomes saturated with respect to CuO(s). The rapid
dissolution in the first 7 days in consistent with the DTPA and CaCl2 extractability data, which
increased most rapidly in the first 7 days, followed by a slower increase. The dissolution of CuO NP
slowed down after 7 days even though the soil has not reached its capacity to adsorb Cu, which
confirms our former assumption that dissolution of CuO NP is the limiting factor controlling the
extractability of Cu from soil. For both the CuO NP and Cu(NO3)2 treatments, XANES analysis
showed no indication of Cu reduction in the soil, confirming our assumption that the experimental
condition was aerobic. Note that we also analyzed the 10mg/kg soils and the unamended soil
samples, but the signal-to-noise ratio was too poor for adequate speciation.
Table 2-1. Modeled first-order dissolution parameters for CuO NP amended soil.
- 23 -
Extraction type k (day-1) 95% confidence intervals
for k (day-1)
Half-life
(days)
E0a
(mg/kg)
Efinal
(mg/kg)
R2
High dose amendment
DTPA extraction
(dry soil)
0.15 0.11-0.19 4.6 3.35 37.4 0.995
CaCl2 extraction
(dry soil)
0.13 0.12-0.18 5.2 0.05 1.0 0.998
Low dose amendment
DTPA extraction
(dry soil)
0.16 0.06-0.25 4.5 0.36 6.71 0.936
CaCl2 extraction
(dry soil)
0.11 0.07-0.14 6.6 0.03 0.16 0.975
a: E0= initial extractable Cu at day 0 (intercept at y axis)
Figure 2-3. Change of Cu speciation in amended soils as inferred by XANES: in (a) Cu(NO3)2
amended soil and (b) CuO NP amended soil dosed at 100 mg/kg total Cu. The red dash lines are
- 24 -
fitted data while the black lines are experimental data. Model compounds used for the fits are below
the experimental spectra. The pie charts represent linear combination fits of the various model
compounds.
2.4 Environmental Implications The extractability of Cu from CuO NP-amended soils is different from that in soils dosed with Cu
ions as Cu(NO3)2, suggesting that the lability of CuO NP may be different from the lability of the
highly soluble Cu salts used as pesticides in soils. CuO NP was much less labile than Cu(NO3)2 in
soil immediately after they were added to the soil, but its lability increased over time. The differences
in lability between CuO NP and Cu(NO3)2 became negligible at low Cu doses (10 mg/kg) after
about 7 days, but differences in lability remained over 31 days for the high dose. The increase of the
labile pool of CuO NP over time was a result of their slow dissolution. Thus, our research shows
that dissolution is an important process controlling the extractability of CuO NP in soil, but the
dissolution rate and CuO NP persistence will be concentration dependent. Moreover, the aging time
in soil must be considered when assessing the lability or bioavailability of CuO NP in soils as was
also previously suggested by Sekine et al for Ag NP, Ag2S NP and AgCl NP and McShane et al. for
CuO NP (dosed at 500 mg/kg)26, 27, along with the total applied dose. If toxicity is purely the result
of the release of copper ion, the regulatory limit for applying nano CuO in agriculture could be
adjusted to consider its “slow release” behavior and concentration-dependent persistence. Because
of the relatively slow dissolution behavior of CuO NP, the regulatory limit for CuO NP could be
higher than that set for Cu salts. This is especially true if, with some additional surface modification,
the dissolution rate of Cu-based nanoparticles could be further reduced. Compared with a direct
spray application of Cu salt, a slow sustained release of ions from CuO NP may have lower
environmental impact to groundwater and rivers because particles have lower leachability and
mobility. On the other hand, if CuO NPs exhibit nanoparticle specific toxicity51,52, for higher doses
where CuO NPs persist, regulations will need to consider this persistence if CuO NPs show greater
toxicity than the Cu salts. Overall, the regulation of nano enhanced particles might be better based
on their dissolution rate at the applied dose, which could be easily determined with the methods
used in this study.
This work advances our understanding of the fate of CuO NP in several important ways. First, we
found CuO NP dissolution is the rate limiting step in controlling the increase of CaCl2 extractable
Cu and DTPA extractable Cu in CuO NP amended soil, indicating the dissolution process of CuO
- 25 -
NP in soil is much slower than the Cu-SOM interaction. Thus, we can monitor the dissolution of
CuO NP in soil from either the increase in dissolved Cu in soil pore water (as indicated by CaCl2
extraction) or increase in extractable Cu by DTPA extraction (dissolved Cu plus Cu bound to SOM
and carbonates). While McShane et al.26 suggested that soil pH is an important factor controlling the
dissolution of CuO NP in soil, we also suggest that the amount of SOM in soil may be as or more
important because it provided the sink for the released Cu in the soils used here. Secondly, our
research indicates that the concentration of soluble nanoparticles added to the soils can affect
temporal changes in Cu speciation, which in turn can affect the interpretation of exposure or toxicity
testing. At a low dose (10 mg/kg dried soil), CuO NPs became fully dissolved within 10 days. Thus,
at low doses, exposures to nanoparticles after ~10 days are not occurring and exposures and toxicity
testing would be expected to be consistent with a dissolved Cu species. Moreover, the Cu species
present was similar to Cu(II)-HA found in the natural soil so responses to CuO NP amended soils at
these low doses would likely be similar to exposures to native soils with the same Cu concentration.
However, using a higher CuO NP dose (100 mg/kg dried soil), about 40% of CuO NPs remained
undissolved after 31 days, potentially because the dissolution was limited by the solubility with
respect to CuO(s). In experiments using this high concentration, exposures and effects may be a
result of interactions with CuO NP and therefore different than for added ions or native soils.
Our research suggests that a single time point extraction after dosing soil may not be adequate for
predicting bioavailability unless that extraction is made at the same time as the end point of interest
(e.g. plant uptake). Rather, a time series of extractions after dosing may be more appropriate for
predicting the bioavailability of metal/metal oxide nanoparticles in soil. The time series of
extractions used here could be developed as a functional assay for studying the dissolution kinetics
of metal/metal oxide nanoparticles in soil. The functional assay approach has recently been
proposed as a means to empirically predict nanomaterial behaviors in complex media53. The method
that we developed is simple, and highly reproducible among the three replicates in our experiments.
The dissolution rate constant could be used for nanomaterial risk forecasting in soil system, as
suggested by Hendren et al53. Further studies need to confirm this method using different
metal/metal oxide nanoparticles in different soil systems. For example, several well-known
limitations of soil extractions methods, e.g. dilution effects, and the presence of an “irreversibly
bound” fraction of metal, exist. In the current study, the irreversibly bound fraction was relatively
low (<20%) and was achieved quickly such that is remained constant during the extraction process.
- 26 -
This enabled calculation of a first-order dissolution rate constant because it is independent of the
extractable amount as long as the extractable percentage is not changing over time. For soils where
the irreversibly bound metal fraction is changing at time scales similar to the NP dissolution, this will
complicate the analysis. However, a time-series extraction using the ionic salt can be used to control
for this “aging” effect. Here, we used XANES analysis to monitor Cu speciation. In our paper, we
determined that the two extractions that we used did not significantly influence the dissolution
process (there was low Cu extraction at Day 0), but this is not necessary true for all nanoparticles. A
preliminary experiment is required to prove that the extraction procedures do not induce significant
particle transformation. Regardless of these limitations, the excellent correlation between
extractability and XANES analysis showed that soil extraction methods are indeed good proxies for
Cu dissolution studies in aerobic soil.
Importantly, for CuO NP, we found that dissolution is the main processes controlling its lability in
aerobic soil. However, for other particles such as metallic Cu-NP, Ag-NP, or other reactive or
redox-sensitive nanomaterials, or different soil conditions (e.g. anaerobic) different processes may
also affect bioavailability. For example, sulfidation of Ag NP and Cu NP have been shown to affect
its properties, fate in soil54, and toxicity55-57. Oxidation may also be an important determinant of
lability in soil28.Future research is needed to better relate these different transformation processes
with nanoparticle lability or bioavailability.
2.5 References of Chapter 2 (1) Tegenaw, A.; Tolaymat, T.; Al-Abed, S.; El Badawy, A.; Luxton, T.; Sorial, G.; Genaidy, A.,
Characterization and potential environmental implications of select Cu-based fungicides and bactericides employed in U.S. markets. Environ. Sci. Technol. 2015, 49, (3), 1294-302.
(2) Giannousi, K.; Avramidis, I.; Dendrinou-Samara, C., Synthesis, characterization and evaluation of copper based nanoparticles as agrochemicals against Phytophthora infestans. RSC Adv. 2013, 3, (44), 21743-21752.
(3) Liu, R.; Lal, R., Potentials of engineered nanoparticles as fertilizers for increasing agronomic productions. Sci. Total Environ. 2015, 514, 131-139.
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(6) Nagajyoti, P.; Lee, K.; Sreekanth, T., Heavy metals, occurrence and toxicity for plants: a review. Environ. Chem. Lett. 2010, 8, (3), 199-216.
(7) Posthuma, L.; Van Straalen, N. M., Heavy-metal adaptation in terrestrial invertebrates: a review of occurrence, genetics, physiology and ecological consequences. Comp. Biochem. Physiol. C: Pharmacol. Toxicol. 1993, 106, (1), 11-38.
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(8) Rodrigues, S.; Trindade, T.; Duarte, A.; Pereira, E.; Koopmans, G.; Römkens, P., A framework to measure the availability of engineered nanoparticles in soils: Trends in soil tests and analytical tools. TrAC, Trends Anal. Chem.2016, 75, 129-140.
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(13) Rao, C.; Sahuquillo, A.; Sanchez, J. L., A review of the different methods applied in environmental geochemistry for single and sequential extraction of trace elements in soils and related materials. Water Air Soil Pollut. 2008, 189, (1-4), 291-333.
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nodulation, shift microbial community composition, and result in increased metal uptake relative to bulk/dissolved metals. Environ. Sci. Technol. 2015, 49, (14), 8751-8758.
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(27) Sekine, R.; Brunetti, G.; Donner, E.; Khaksar, M.; Vasilev, K.; Jamting, A. K.; Scheckel, K. G.; Kappen, P.; Zhang, H.; Lombi, E., Speciation and Lability of Ag-, AgCl-, and AgS-Nanoparticles in Soil Determined by X-ray Absorption Spectroscopy and Diffusive Gradients in Thin Films. Environ. Sci. Technol. 2014, 49, (2), 897-905.
(28) Gomes, S. I.; Murphy, M.; Nielsen, M. T.; Kristiansen, S. M.; Amorim, M. J.; Scott-Fordsmand, J. J., Cu-nanoparticles ecotoxicity–Explored and explained? Chemosphere 2015, 139, 240-245.
(29) Collins, D.; Luxton, T.; Kumar, N.; Shah, S.; Walker, V. K.; Shah, V., Assessing the impact of copper and zinc oxide nanoparticles on soil: a field study. PLoS One 2012, 7, (8), e42663.
(30) Unrine, J. M.; Tsyusko, O. V.; Hunyadi, S. E.; Judy, J. D.; Bertsch, P. M., Effects of Particle Size on Chemical Speciation and Bioavailability of Copper to Earthworms (Eisenia fetida) Exposed to Copper Nanoparticles. J. Environ. Qual. 2010, 39, (6), 1942-1953.
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(35) Bian, S.-W.; Mudunkotuwa, I. A.; Rupasinghe, T.; Grassian, V. H., Aggregation and dissolution of 4 nm ZnO nanoparticles in aqueous environments: influence of pH, ionic strength, size, and adsorption of humic acid. Langmuir 2011, 27, (10), 6059-6068.
(36) Zhang, W.; Yao, Y.; Sullivan, N.; Chen, Y., Modeling the primary size effects of citrate-coated silver nanoparticles on their ion release kinetics. Environ. Sci. Technol. 2011, 45, (10), 4422-4428.
(37) Mudunkotuwa, I. A.; Rupasinghe, T.; Wu, C.-M.; Grassian, V. H., Dissolution of ZnO nanoparticles at circumneutral pH: a study of size effects in the presence and absence of citric acid. Langmuir 2011, 28, (1), 396-403.
(38) Peretyazhko, T. S.; Zhang, Q.; Colvin, V. L., Size-controlled dissolution of silver nanoparticles at neutral and acidic pH conditions: kinetics and size changes. Environ. Sci. Technol. 2014, 48, (20), 11954-11961.
(39) Jiang, C.; Aiken, G. R.; Hsu-Kim, H., Effects of natural organic matter properties on the dissolution kinetics of zinc oxide nanoparticles. Environ. Sci. Technol. 2015, 49, (19), 11476-11484.
(40) Dale, A. L.; Casman, E. A.; Lowry, G. V.; Lead, J. R.; Viparelli, E.; Baalousha, M., Modeling nanomaterial environmental fate in aquatic systems. Environ. Sci. Technol. 2015, 49, (5), 2587-2593.
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(41) J. B. Peters, M. V. N., C. A. M. Laboski, pH and Lime Requirement. In Recommended Chemical Soil Test Procedures for the North Central Region, North Central Regional Research Publication No. 221 (Revised): 2012, pp 4.1-4.7.
(42) Ghani, A.; Dexter, M.; Perrott, K., Hot-water extractable carbon in soils: a sensitive measurement for determining impacts of fertilisation, grazing and cultivation. Soil Biol. Biochem. 2003, 35, (9), 1231-1243.
(43) Webb, S., SIXpack: a graphical user interface for XAS analysis using IFEFFIT. Phys. Scr. 2005, (T115), 1011.
(44) Eckert, D.; Sims, J. T., Recommended soil pH and lime requirement tests. Recommended soil testing procedures for the northeastern United States. Northeast Regional Bulletin 1995, 493, 11-16.
(45) Mao, L.; Young, S. D.; Bailey, E. H. Lability of copper bound to humic acid. Chemosphere 2015, 131, 201–208.
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(48) Jacobson, A. R.; Dousset, S.; Andreux, F.; Baveye, P. C., Electron microprobe and synchrotron X-ray fluorescence mapping of the heterogeneous distribution of copper in high-copper vineyard soils. Environ. Sci. Technol. 2007, 41, (18), 6343-6349.
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CHAPTER 3: Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution
kinetics of CuO NP in soil. Abstract: The objectives of this research were to quantify the impact of organic matter content, soil pH and
moisture content on the dissolution rate and solubility of copper oxide nanoparticles (CuO NPs) in
an aerobic soil, and to develop an empirical model to predict the dissolution kinetics of CuO NPs in
an aerobic soil. CuO NPs were dosed into standard LUFA soils with various moisture content, pH
and organic carbon content. Chemical extractions were applied to measure the CuO NP dissolution
kinetics. Doubling the reactive organic carbon content in LUFA 2.1 soil increased the solubility of
CuO NP 2.7-fold but did not change the dissolution rate constant. Increasing the soil pH from 5.9
to 6.8 in LUFA 2.2 soil decreased the dissolution rate constant from 0.56 mol1/3·kg1/3·s-1 to 0.17
mol1/3·kg1/3·s-1 without changing the solubility of CuO NP in soil. For six soils, the solubility of CuO
NP correlated well with soil organic matter content (R2 = 0.89) independent of soil pH. In contrast,
the dissolution rate constant correlated with pH for pH<6.3 (R2 = 0.89). These relationships
predicted the solubility and dissolution rate constants of CuO NP in two test soils (pH=5 and
pH=7.6). Moisture content showed negligible impact on the dissolution kinetics of CuO NPs. Our
study suggests that soil pH and organic matter content affect the dissolution behavior of CuO NP in
soil in a predictable manner.
This work has been published in Environmental Science & Technology as ‘Effect of Soil Organic
Matter, Soil pH, and Moisture Content on Solubility and Dissolution Rate of CuO NPs in Soil’ , doi:
acs.est.8b07243
- 31 -
3.1 Introduction
Copper(Cu)-based nanoparticles (NPs) have been used in agriculture as fungicides and have
potential for use as trace element fertilizers1–5. Either through terrestrial application or foliar
application, these NPs will intentionally or inadvertently enter soils, where their fate depends
primarily on dissolution processes 6–8. Given that the release of Cu2+ ions is the primary mode of
toxicity of the copper oxide nanoparticles (CuO NPs)9,10, dissolution will affect both bioavailability
and toxicity of Cu-based NPs to plants and soil organisms11. Thus, dissolution rate and solubility are
properties of practical relevance12.
Solubility is the maximum mass of the CuO NPs that will dissolve and partition into different soil
compartments at equilibrium, and dissolution rate as the speed at which the ion release processes
occurs. These processes are influenced by environmental factors. Although the impact of matrices’
physical-chemical properties (e.g. concentration of organic matter (OM), pH, ionic strength) on the
dissolution rate of metal and metal (hydr)oxide nanoparticles in aquatic systems has been reported13–
16, such impacts in soils have not yet been systematically studied. Finally, while ionic Cu release from
Cu-based NP has been assessed in pore water7,17 and speciation in soil evaluated at a few given time
points6,8, abiotic factors influencing these dissolution behaviors have not been investigated yet,
resulting in the absence of adequate models to predict CuO NP dissolution kinetics in soils.
Measuring the dissolution of NPs in soil is not straightforward and studying their dissolution
kinetics is even more challenging. Recently, two methods have been developed to study the
dissolution of NPs in soil. The first method monitors the change in speciation of the metal in soil
over time using X-ray absorption spectroscopy (XAS)18,19. Transformation and dissolution is inferred
from an observed decrease in the fraction of the original NPs 6,8,18,20. Due to constraints on
synchrotron access and the relatively high cost of measuring each sample, such studies rarely provide
enough data to quantify the dissolution kinetics of NPs in soil, which requires multiple
measurements over time. The high cost also precludes XAS as a routine method for studying NP
dissolution processes.
Two soil extraction methods, developed originally to study the geochemical equilibrium distribution
of metals in soil and to predict their bioavailability to plants21, offer a less costly alternative. The first
method involves extracting pore water metals from soil using either dilute salt (e.g. 0.01M calcium
chloride, CaCl2 )22–25 or water7,26. It is used as a proxy for extracting soil pore water metal from
soil22,27,28. The extracted metal is considered to be ‘readily available’ to plants or ‘highly mobile Cu’.
- 32 -
The second extraction method is for labile metal and extracts ‘potentially available’ metals, i.e., those
that are reversibly bound to the soil solid matrix. Such extractions use chelating agents or dilute
strong acid (e.g., 0.005M Diethylenetriamine pentaacetate, DTPA, or 0.05M
Ethylenediaminetetraacetic acid, EDTA)24,25,29–31. The DTPA extraction method has been
demonstrated to extract most of the ionic Cu released from CuO NP6,10. The change over time in
extractability of metals in soil can be used as a proxy of metal NP dissolution. Both methods have
been assessed and can be used to monitor the dissolution processes of Cu based NPs in soil6,7.
Dissolution of metal and metal oxide NPs (e.g. CuO NPs) is most often an acid-promoted process.
Thus, the concentration of hydrogen ion (pH) plays a role in both the expression of the dissolution
rate law, and the equilibrium constant (Equation 3-1). OM, as well as other natural metal chelators,
(e.g., siderophores, or amino acids), can also affect dissolution by binding with metal ions released
by the NPs or by interacting with the particle (e.g. through coatings32) and affecting the total surface
available for dissolution. The dissolution of NPs also requires the interaction between NP and the
soil pore water. Given that moisture content in agricultural and natural soils varies over time and
space, it is also important to know how moisture content affects the dissolution of CuO NP in soil.
Redox potential is another influence to consider for the dissolution of CuO NP in soil. However,
for most agricultural soils, the plow layer, where agrochemicals are applied, is intended to be aerobic
(the redox potential is usually above 400mv due to the interaction with air33). At this redox
condition, reduction of Cu is not thermodynamically favored34. Thus, the influence of redox
potential on dissolution of CuO NP was not explicitly addressed in this work.
𝐶𝐶𝐶𝐶𝑂𝑂 𝑁𝑁𝑁𝑁𝑁𝑁(𝑁𝑁) + 2𝐻𝐻+(𝑎𝑎𝑎𝑎) ↔ 𝐶𝐶𝐶𝐶2+(𝑎𝑎𝑎𝑎) + 𝐻𝐻2𝑂𝑂 (3 − 1)
Recent efforts have been made to evaluate the environmental factors influencing dissolution kinetics
of CuO NP in soil. McShane et al. showed that the free Cu2+ concentration from CuO NPs
(measured using an ion selective electrode) in solution extracted with water from soil was affected by
soil pH, concluding that pH affected CuO NP dissolution in soil7. The correlation of free Cu2+ in
solution with dissolution is consistent with expectations based on CuO NP solubility in water.
However, previous studies have shown that pH also affects the partitioning of Cu2+ between pore
water and the soil solid surfaces. Higher pH results in more Cu2+ binding to soil organic matter
(SOM)29,35. Therefore, it is necessary to extract the dissolved Cu from the soils to ensure that all of
the dissolved Cu is accounted for in the measurement6. Another study used XAS to track the
- 33 -
changes in speciation of Cu in CuO-NP-amended soil. They also found that lower pH resulted in
higher dissolution in the short term (within 5 days) and that the SOM content slowed the dissolution
process in the short term8. The latter finding still requires more investigation since it contradicts the
findings from studies in water showing that SOM increased CuO NP dissolution13,15. SOM is an
important Cu sink/pool in soil36 because nitrogen and oxygen atoms in SOM can strongly bond
with Cu37, thus one would expect it should increase the solubility of CuO NP in soil. The studies
mentioned above also used a variety of soils to demonstrate the effect of soil pH and SOM on the
dissolution of CuO NP. This introduces potentially confounding variables as environmental factors,
e.g. soil texture and field capacity could also potentially affect the dissolution of CuO NP by
affecting CuO NP-soil aggregation and the distribution of CuO NP between soil pore water and soil
solids.
The objective of this study is to quantify the effect of pH, SOM content, and moisture content on
the dissolution rate and solubility of CuO NP in soil. We used several standard agricultural soils at
different pH and with different moisture and SOM content to investigate the influence of these soil
properties on CuO NP dissolution behavior. Dissolution models were then used to quantify the
effect of soil pH and SOM content on the dissolution kinetics of CuO NP in soil. Finally, the ability
of this model to predict the dissolution kinetics and solubility of CuO NP in soil was evaluated using
two test soils with different properties.
3.2 Method and Materials
3.2.1 Chemicals
Calcium chloride (≥99.0%, ACS grade), calcium oxide (CaO), calcium carbonate (CaCO3) (99%+),
and hydrogen peroxide (30%, certified ACS) were purchased from Fisher Scientific. DTPA (>99%)
and triethanolamine (TEA, ≥99.0% (GC)) were purchased from Sigma-Aldrich. Trace metal grade
nitric acid (65%-70%) was purchased from VWR. Copper sulfate (CuSO4) was purchased from
Fisher Scientific . Lufa Standard soils (2.1, 2.2, 2.4 and 2.4) were purchased from Lufa Speyer,
Germany. A calcareous soil (pH 7.6) was collected in Arizona (termed Arizona soil) and used to test
the model’s ability to predict CuO NP dissolution behavior based on soil pH and SOM content.
Another more acidic soil (pH=5.0) was collected from a grassland in northwestern Portugal (termed
Portugal soil). Detailed properties of all the soils used can be found in appendix ( Table A2-1).
- 34 -
3.2.2Nanoparticle Characterization
CuO NPs (~40 nm primary particle size, zeta potential (ζ) = -16.1 mV ± 1.7mV at pH=7 in 5mM
NaNO3), were purchased from Sigma-Aldrich. The primary size of particles, zeta potential,
isoelectric point and hydrodynamic diameter have been characterized and reported in Chapter 26.
3.2.3 Soil amendment Soil pH, SOM content and moisture content, factors hypothesized to affect dissolution kinetics of
CuO NP in soil, were systematically varied in this study (the soil properties for all treatments can be
find in Table A2-2, Appendix 2). To investigate the effect of pH on the dissolution of CuO NP, a
mixture of CaO and CaCO3 powders were used to increase the soil pH from the original pH of 5 to
~7.5 for Lufa 2.1 soil (0.27g CaO, 0.68g CaCO3 in 270g of Lufa 2.1 soil), and from 5.9 to 6.8 for
Lufa 2.2 soil (0.27g CaCO3 in 270g of Lufa 2.2 soil)38. To investigate the influence of SOM on
dissolution of CuO NP with all other soil properties held constant, the soil total organic carbon
(TOC) content in Lufa 2.1 soil was increased from the original 0.7% to 0.9% by adding SOM
extracted from Lufa 2.1 soil. Note that generally the SOM content is ≈ 1.74 times the soil organic
carbon content, although this can vary between soil types39. SOM was extracted from Lufa 2.1 soil
following a procedure described by van Zomeren et al.40 Additional details on SOM extraction,
recovery, and preliminary characterization are provided in Appendix 2. Only about 23% of organic
carbon in Lufa 2.1 soil was extractable. This 23% is considered to be the ‘reactive organic carbon,’
the SOM fraction that usually controls the Cu sorption behavior. The remaining fraction was mostly
humic substances that have low affinity for metals41. In this study, 161mg extracted fulvic acid, FA,
and 368mg extracted humic acid, HA, was added to 90g Lufa 2.1 soil. In the original soil
(TOC=0.7%), the reactive carbon content was 0.16%. Thus, by adding 0.2% of reactive organic
carbon content in soil, the total reactive carbon in Lufa 2.1 soil was effectively doubled. (Note
carbon content in HA and FA are provided in Appendix 2.) CuO NPs and CuSO4 (control
treatment) were added to different soils to achieve final concentrations of 100 mg/kg, 250 mg/kg
and 500 mg/kg dry weight (dw) (as Cu). To investigate the influence of moisture content with all
other soil properties held constant, we used Lufa 2.2 standard soil at 21% and 10% moisture
content. The two moisture contents were selected because they span relevant moisture conditions,
on one end where the soil is as wet as it could be (field capacity) and the other as dry as it could
reasonably be (wilting point) for an agricultural soil. CuO NPs were also dosed into the Arizona soil
(500mg/kg Cu dw) and Portugal soil (500mg/kg Cu dw) to test our models’ ability to predict
solubility and dissolution rate of CuO NP in natural soils. The concentration of CuO used in each
- 35 -
treatment was selected based on the solubility of the CuO NPs in each soil determined in
preliminary studies (Appendix 2). Enough CuO NPs was added to each treatment to ensure that
some CuO NPs remained undissolved after 30d. Details on the treatment condition and Cu mass
balance are in appendix, Table SA-2. Note that during the 30d incubation period, all soils were
maintained under aerobic conditions (soils were incubated in centrifuge tubes with holes in the caps
allowing air exchange). It was verified in a previous study that these experimental conditions
precluded significant Cu reduction 42. For topsoil in agriculture (Eh>400mv)53,Cu(II) is the major Cu
valence state54. Thus Cu in or released from the CuO NPs are is assumed to remain in the Cu(II)
redox state.
3.2.4 Extraction procedure to measure the fraction of dissolved CuO NP and soil pH. The amount of CuO NP that had dissolved at each incubation time (days 0, 2, 4, 7, 14, 21, 30 after
amendment) and the corresponding soil pH at that time point, were measured using a previously
published extraction method6. Briefly, for each Cu treatment, 2.0 g of air-dried soils were extracted
with two standard extractants: (1) 4 mL of DTPA (0.05 M DTPA, 0.01M CaCl2 and 0.1M TEA at
pH 7.6) and (2) 20 mL of 0.01 M CaCl2 (pH =5). All extractions were done in a reciprocal shaker at
180 rpm for 2 hours. After extraction, samples were centrifuged and filtered with 0.45µm PTFE
filters. Then, the filtered samples were acidified and analyzed by ICP-MS (Agilent technologies
7700). The measurements were made right after each aging period. It should be noted that our
previous studies have demonstrated that such extractions did not induce any CuO NP dissolution6.
The pH of CaCl2 extracts for air-dried amended soils were measured as soil pH using a common
procedure43,44.
3.2.5 Determination of Cu speciation in soils
Cu speciation in soils after amendment was analyzed by Cu K-edge XAS at the Stanford
Synchrotron Radiation Lightsource (SSRL) on Beamline 11-2. Details on sample preparation and
measurements can be found in the appendix.
3.2.6 Dissolution models.
The model used for CuO NP dissolution in soil includes the following steps (Figure 3-1): (1): CuO
NP dissolves (reversibly, with rate constants kd and kr)), releasing free Cu ions into the soil pore
water. (2): Cu2+ attaches to different ligands (e.g. dissolved organic matter (DOM)) and soil surfaces
(e.g. clay, SOM) 45. The second step (Cu ion partitioning between soil pore water and soil solid
- 36 -
surfaces) has been investigated previously 29,46–48. The reversible dissolution of CuO NPs are of
primary interest to this study.
Figure 3-1. Schematic of CuO NP dissolution model. Where 𝒌𝒌𝒅𝒅 is the dissolution rate constant, 𝒌𝒌𝒓𝒓
is the local reverse reaction rate constant for Cu(II) ions precipitating back onto the CuO NP
surface (precipitation). Note that this reverse reaction must be occurring locally near the CuO NP
surfaces if the particles are in local equilibrium with the surrounding water. 𝑲𝑲𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒅𝒅 is the
partitioning constant between Cu associated with natural ligands (including both DOM and soil
surfaces, e.g. SOM, clay, iron oxides) and free Cu2+(aq). 𝒌𝒌𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍𝒍 is the constant to account for
irreversible loss of Cu to the matrix over long time spans. It should be noted that only the CuO NP
dissolution parameters, highlighted in purple, are new additions to the well-known multi-surface
geochemical model47,49.
To model the dissolution kinetics, we define Cu2+Tot as the total concentration of Cu2+ being released
from CuO NP (free 𝐶𝐶𝐶𝐶2+ + 𝐶𝐶𝐶𝐶 𝑎𝑎𝑁𝑁𝑁𝑁𝑜𝑜𝑜𝑜𝑜𝑜𝑎𝑎𝑜𝑜𝑜𝑜𝑑𝑑 𝑤𝑤𝑜𝑜𝑜𝑜ℎ 𝑛𝑛𝑎𝑎𝑜𝑜𝐶𝐶𝑛𝑛𝑎𝑎𝑛𝑛 𝑛𝑛𝑜𝑜𝑙𝑙𝑎𝑎𝑛𝑛𝑑𝑑𝑁𝑁) , which can be extracted by
DTPA. If we assume that Cu2+Tot (t=0) = 0 and that [H+] remains constant during the dissolution
(implying a stable pH during the dissolution process due to the relatively high buffering capacity of
soil6), the rate law can be expressed by equation (3-2).
𝑑𝑑[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑇𝑇𝑇𝑇,𝑇𝑇𝑑𝑑𝑑𝑑
= 𝑘𝑘𝑑𝑑([𝐶𝐶𝐶𝐶𝑂𝑂]𝑜𝑜 − [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑)2/3 − 𝑘𝑘𝑟𝑟[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑1
1+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙([𝐶𝐶𝐶𝐶𝑂𝑂]𝑜𝑜 − [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑)2/3
(3-2)
- 37 -
The derivation of Equation (3-2) can be found in Appendix 2. The key assumptions are:
1: The Cu2+ released by CuO NP is in equilibrium with respect to its partitioning to other soil
components, e.g. DOM and SOM. This equilibrium is fast compared to the rate of dissolution.
2: The solubility of CuO NP(s) is limited by the local dissolution/precipitation equilibrium.
The dissolution of CuO NP(s) in soils is not complete. A reverse reaction, which is a precipitation
process, must occur at the surface of CuO NP(s) to stop CuO NP from completely dissolving.
Dissolution stops when the dissolution rate near the CuO NP surface equals the reverse reaction
rate near the NP surface. The precipitation of Cu2+ preferentially happens near the surface of CuO
NP because of the localized higher Cu2+ concentration on the surface of the NP.
3: We assume that precipitation of Cu phases other than CuO is not significant.
This was corroborated with the facts that (a) ~80% of Cu was still extractable by DTPA in the Lufa
2.2 soil amended with a high concentration of CuSO4 (500 mg/kg), which did not form a solid
phase6; and (b) the Cu X-ray absorption near edge structure (XANES) spectra of Lufa 2.2 soil dosed
with 500mg/kg CuSO4 indicated that 99.6% of the Cu was present as Cu-NOM after 30 days
(appendix, Figure A2-1). It should be noted that the process of Cu2+ sorbed to the soil organic
matter (SOM) is not considered a ‘precipitation’ process, rather, it is a sorption process.
4: We assume the dissolution/precipitation of CuO NP are both surface-controlled process, e.g.
dissolution rate and the reverse reaction rate are both proportional to the total surface area of CuO
NP. Moreover, we assume that the CuO NPs are spherical and that their surface area changes
according to a 2/3 power law as has been previously described with the dissolution of spherical
ZnO NPs15.
At equilibrium, 𝑑𝑑[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑇𝑇𝑇𝑇,𝑇𝑇𝑑𝑑𝑑𝑑
= 0 so the solubility of the CuO NPs in the soil, [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞, is given
by Equation (3).
[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞ =𝑘𝑘𝑑𝑑𝑘𝑘𝑟𝑟
(1 + 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑) (3 − 3)
Equation (2) can be re-written using 𝑘𝑘𝑑𝑑 and [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞:
- 38 -
Equation (3-4) was applied to estimate the unknown constants, 𝑘𝑘𝑑𝑑 , 𝑘𝑘𝑟𝑟 𝑎𝑎𝑛𝑛𝑑𝑑 [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,∞ from fits of
the dissolution data collected for the soils over time. Note that these three parameters are correlated
by Equation (3-3). The Euler method was applied to solve equation 3-3 numerically. 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 was
estimated from the experimental data (Equation 3-5). From control experiments extracting Cu from
CuSO4 dosed soil, the efficiency of DTPA extraction, 𝜂𝜂𝐷𝐷𝑇𝑇𝐷𝐷𝐷𝐷 , was estimated to be 80%.
𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 =[𝐶𝐶𝐶𝐶]𝐷𝐷𝑇𝑇𝑃𝑃𝐴𝐴 𝜂𝜂𝐷𝐷𝑇𝑇𝑃𝑃𝐴𝐴
[𝐶𝐶𝐶𝐶]𝐶𝐶𝑙𝑙𝐶𝐶𝑙𝑙2∙𝑥𝑥𝐶𝐶𝐶𝐶2+ (3-5)
Where as [𝐶𝐶𝐶𝐶]𝐷𝐷𝑇𝑇𝐷𝐷𝐷𝐷 is DTPA extractable Cu, 𝜂𝜂𝐷𝐷𝑇𝑇𝐷𝐷𝐷𝐷 is the extraction efficiency (0.8 in this study),
[𝐶𝐶𝐶𝐶]𝐶𝐶𝑓𝑓𝐶𝐶𝑓𝑓2 is CaCl2 extractable Cu, and 𝑥𝑥𝐶𝐶𝐶𝐶2+ is the fraction of free Cu ions in soil pore water.
3.3 Results and Discussion
3.3.1 Effect of Soil Organic Matter on dissolution of CuO NP in soil.
To investigate the effect of SOM on dissolution of CuO NP in soil, a dissolution test in Lufa 2.1 soil
(100 mg/kg dw CuO NP treatment) and in Lufa 2.1 with added SOM (300 mg/kg dw CuO NP
treatment) was conducted (Figure 3-2). Different concentrations of CuO NP were applied based on
the estimated solubility from preliminary experiments (described in Appendix 2). Using the
dissolution model described in the methods section, the modeled solubility should increase from 95
mg/kg ( 95% CI: 87-108 mg/kg) to 254 mg/kg (95% CI: 234-280 mg/kg) in the amended soil
(Table 1). Doubling the reactive organic carbon content in Lufa 2.1 soil increased the solubility of
CuO NP by 2.7-fold, suggesting reactive organic carbon holds the main Cu pool in soil. Although
the solubility increased by 2.7-fold, the modeled dissolution rate constants between Lufa 2.1 soil and
Lufa 2.1 soil with added SOM are similar (95% confidence intervals are overlapping), suggesting that
SOM mainly affects the solubility of CuO NP in soil, but not its dissolution rate.
- 39 -
0 10 20 30 400
50
100
150
200
250
Time (days)
DTPA
Ext
ract
able
Cu
(mg
/kg
drie
d so
il) Lufa 2.1 soil, SOM added
Lufa 2.1 soil
Figure 3-2. Dissolution kinetics of CuO NP in Lufa 2.1 soil without added SOM (100 mg/kg dw
CuO NP treatment, circles) or with added SOM (300mg/kg dw CuO NP treatment, triangles). Bars
are standard deviation of the extractable Cu measurements (3 replicates). Soil pH in these studies
was 5.0 (unamended Lufa 2.1 soil) and 4.9 (Lufa 2.1. amended with SOM).
3.3.2 Effect of soil pH on dissolution of CuO NP in soil. The effect of soil pH on the dissolution behavior of CuO NP was investigated by modifying the pH
of Lufa 2.1 soil (100 mg/kg dw CuO NP treatment) and Lufa 2.2 soil (500 mg/kg dw CuO NP
treatment) with either CaO or CaCO3. Figure 3 indicates that higher pH significantly slowed down
the dissolution rate of CuO NP in soil in Lufa 2.2 soil. The modeled dissolution rate constant
decreased from 0.56 (CI95: 0.35-0.84)) (mg1/3·kg1/3·s-1) in Lufa 2.2 soil (pH=5.9) to 0.17 (CI95: 0.14-
0.21) (mg1/3·kg1/3·s-1) in Lufa 2.2 soil with pH adjustment (pH=6.8). For Lufa 2.1 soil (Appendix 2,
Figure A2-2), the dissolution of CuO NPs in pH-adjusted soil (pH=7.4) could not be accurately
modeled because of very limited dissolution, but it was clear that it was much slower than the
dissolution in Lufa 2.1 soil without pH adjustment (pH=5.0, kd= 0.83 mg1/3·kg1/3·s-1, with 95% CI:
0.65-1.00) during the 31d aging period. Although the dissolution rate constants are different,
suggesting a different particle lifetime in soil, the modeled solubility of CuO NPs in Lufa 2.2 soil
with and without pH adjustment are similar (Table 3-1). This can be observed from the extended
trend lines (dash lines) from the modeled dissolution kinetics in Figure 3-3. Thus, the soil pH mainly
determines how fast CuO NPs dissolve but has no measurable impact on their solubility. This is because most of
the Cu ions released from CuO NPs are retained by SOM. Carboxylic acid functional groups (pKa
- 40 -
<5) and weak acid groups (phenolics, pKa>9) mainly contribute to the acidity of humic acid (the
main component of SOM)50,51. The binding capacity between Cu and SOM is not sensitive to pH at
agriculture soil relevant pH (5 ~ 7.5)52 because the protonation state of SOM is not susceptible to
pH variation in this range. Thus, for a typical agriculture soil, although an increase in soil pH should
slow down the ion release process from CuO NP, it may have limited impact on the solubility of
CuO NP in that soil.
Figure 3-3. DTPA extractable Cu in Lufa 2.2 soil dosed with 500 mg/kg CuO NP at pH 5.9
(squares) and pH 6.8 (triangles). Dashed lines are model results showing the longer time trend. ‘X’ at
t=300 days is modeled maximum DTPA extractable Cu for each treatment. Bars are standard
deviation of the measurements (3 replicates) or the 95% confident intervals of the modeled
maximum DTPA extractable Cu (t= 300 day).
3.3.3 Effect of soil moisture content on the dissolution rate and solubility of CuO NP in soil.
As suggested from Figure 3-4, moisture content had no impact on the dissolution kinetics of CuO
NP. The modeled dissolution rate constants (kd and kr) and solubility [𝑪𝑪𝑪𝑪𝟐𝟐+]𝑻𝑻𝑻𝑻𝑻𝑻,∞ are the same for
CuO NP dissolving in soil with 10% moisture content or with 21% moisture content (Table 3-1).
This is consistent with the dissolution model that we proposed in which the soil pore water reaches
an equilibrium state with the soil solid matrix, where most dissolved Cu is retained by the soil solid
surfaces, not the soil pore water6,46. Thus, soil moisture should not affect the dissolution rate or
solubility of CuO NPs. It is acknowledged that we did not test extremes of dryness (e.g. moisture
content <<10%) because they do not represent normal agricultural soil conditions. CuO NP
dissolution could potentially be affected by extreme dryness due to the lack of water needed to
dissolve the CuO NPs. It should be noted that we also did not consider the flooded condition,
- 41 -
which would be relevant for crops like rice. In a flooded condition, where microbial processes can
deplete the soil of oxygen and lead to a strongly reducing condition, redox reactions could play a
role, affecting the dissolution of CuO NP.
0 10 20 30 400
100
200
300
400
DTPA extraction, Lufa 2.2 soil
Time (days)
Extra
ctab
le C
u(m
g /k
g dr
ied
soil) 21% moisture content
10% moisture content
a)
0 10 20 30 400
1
2
3
Time (days)
Extra
ctab
le C
u(m
g /k
g dr
ied
soil) 21% moisture content
10% moisture content
b) CaCl2 extraction, Lufa 2.2 Soil
Figure 3-4. Effect of moisture content on the dissolution kinetics of CuO NP in soil. (a) DTPA
Extractable Cu in Lufa 2.2, (b) CaCl2 extractable Cu in Lufa 2.2 soil at field capacity or 10%
moisture content. Bars are standard deviations of the measurements (3 replicates).
3.3.4 Dissolution rate and solubility of CuO NPs in soils with various properties.
To further investigate the effects of soil pH and SOM content, which are the more important
factors controlling the dissolution of CuO NP in soils, CuO NP dissolution tests in soils with
various soil properties, including Lufa 2.1, 2.2, 2.3 and 2.4 soils and pH/organic carbon amended
soils, were measured. The speciation of Cu in selected CuO NP dosed soils was confirmed by
XANES (appendix). The experimental conditions (soil properties, NP concentration) and fitted
dissolution model parameters (kd, kr and solubility) are shown in Table 3-1.
- 42 -
Table 3-1. Dissolution rate and solubility of CuO NP in a range of soils with various properties
Soil Moisture content
Concentration of CuO NP (mg/kg
dw)
Organic carbon content (% C)
pH kd (mg1/3·kg1
/3·s-1)
kr (mg-
2/3·kg-2/3·s-
1)
Solubility (mg/kg)
R2
Lufa 2.1 soil
16% 115±7 0.67 5.0 0.83 (0.65-1.00)
0.19 (0.13-0.26)
95 (87-108)
0.992
Lufa 2.2 soil
21% 503±17 1.71 5.9 0.70 (0.55-0.92)
0.23 (0.15-0.33)
328 (300-370)
0.993
Lufa 2.4 soil
22% 537±89 1.99 7.2 0.12 (0.08-0.17)
0.09 (0.00-0.22)
315 (170-*1)
0.991
Lufa 2.3 Soil
17% 539±3 0.66 6.5 0.15 (0.11-0.21)
0.26 (0.15-0.38)
84 (75-99)
0.987
Lufa 2.2 soil- pH adjusted
21% 501±11 1.71 6.8 0.17 (0.14-0.21)
0.05 (0.02-0.08)
295 (220-550)
0.998
Lufa 2.2 soil
21% 265±13 1.71 5.8 0.56 (0.35-0.84)
0.19 (0.01-0.41)
319 (220-3800)
0.986
Lufa 2.2 soil
10% 481±24 1.71 5.9 0.57 (0.43-0.72)
0.20 (0.13-0.29)
325 (287-405)
0.994
Lufa 2.1 soil- pH
adjusted2
16% 112±5 0.67 7.4 N.A. N.A. N.A. N.A.
Lufa 2.1 soil- OM adjusted
16% 287±16 1.34 4.9 0.99 (0.84-1.18)
0.05 (0.03-0.06)
254 (234-280)
0.996
1: The upper bond of solubility cannot be determined with confidence.
2:Dissolution was too low to be modeled with confidence.
- 43 -
To investigate the correlations between soil properties (soil pH, organic carbon content) and
dissolution kinetics (solubility and dissolution rate constant), we conducted a multivariate regression
(Appendix 2 table A2-5 to-A2-7). The result showed that the solubility was correlated with soil organic
carbon content, but not {H+} whereas dissolution rate constant was correlated with {H+} but not organic carbon
content. A strong correlation between organic carbon content and solubility of CuO NP in soil
(Figure 3-5a, R2 = 0.89, 𝑁𝑁 < 0.0001, Solubility = 1.83 ∗ 102 ∗ 𝑂𝑂𝐶𝐶%) was observed, suggesting
soil organic carbon content is the main driver for the solubility of the CuO NPs in soil across soil
types.. It should be noted that our study did not cover soils with extreme high organic carbon
content (~10%). The SOM in our study ranged from ~0.8% to 3.5% (0.5% to 2% organic carbon
content). Thus, the developed correlation covers the low end to an intermediate range for
agricultural soils, but is likely to be predictive of 10% SOM (2.9x higher than our high end). On the
other hand, hydrogen ion activity, {H+}, is positively correlated with the dissolution rate constant
(Figure 3-5b). According to previous studies in aqueous systems, pH should play an important role
in the dissolution rate of CuO NP13,53,54. However, because most previous studies used empirical first
order dissolution models, the reaction mechanism and reaction order of CuO NP dissolution with
respect to hydrogen concentration remained unknown. In this study, no assumption about the
reaction order with respect to {H+} was made. Instead, reaction order was experimentally
determined by plotting the dissolution rate constant against {H+}. Figure 3-5 b shows the linear
relationship between the dissolution rate constant and {H+} for soil pH below 6.3 (R2 = 0.89,𝑁𝑁 =
0.016 𝑘𝑘𝑑𝑑 = 2.83 ∗ 104 ∗ {𝐻𝐻+} + 0.57 ). However, above pH=6.3 this relationship no longer
holds, and the dissolution rate constant is approximated by a single value (𝑘𝑘𝑑𝑑,𝑓𝑓𝑎𝑎𝑎𝑎𝑟𝑟𝑓𝑓𝑙𝑙𝑎𝑎 =
0.14, 95% 𝐶𝐶𝐼𝐼(0.07 − 0.21). Due to the limited amount of dissolution rate and solubility data, a
cross-validation was done, suggesting the fit was stable among the different soils (i.e., no particular
soil treatment inordinately affected the correlation, see Figure A2-3, Appendix 2). The pH-
dependent correlations suggest that the dissolution of CuO NP in soil with pH below 6.3 may be
governed by a different dissolution mechanism (Equation 3-6) compared to CuO NP dissolution in
soil with pH above 6.3 (Equation 3-7). At higher pH, the activity of H+ decreases, thus Equation 3-7
could be the dominant reaction pathway rather than Equation 3-6. Regarding the reverse reaction
(precipitation) process, no trend was found between kr and {H+} and organic carbon content (Table
A2-7). However, kr was correlated with kd and solubility by Equation (3-2) and is thus affected by
both the pH and SOM simultaneously.
- 44 -
CuO (s)+2H+(aq) ↔ Cu2+
(aq) +H2O(l) (3-6)
CuO(s) + H2O(l) ↔ Cu2+(aq)+ OH- (3-7)
0 . 0 0 . 5 1 . 0 1 . 5 2 . 0 2 . 5
0
2 0 0
4 0 0
6 0 0
O r g a n i c c a r b o n c o n t e n t ( % C )
So
lub
ilit
y (
mg
/kg
)
a )* *
Figure 3-5. Correlation between organic carbon content and solubility (a) and between {H+}and
dissolution rate constant, kd (b). The right figure in b) shows the high soil pH data in the red box. *:
upper 95% CI was high, see Table 3-1.
b)
- 45 -
3.3.5 Predicting CuO NP solubility and dissolution rate in two test soils. Using the correlations derived from Figure 3-5 a and b, we estimated the solubility and dissolution
rate constant of the CuO NPs in soil samples from Arizona and from Portugal (soil properties can
be found in appendix, table A2-1) based on their pH and soil organic carbon content. Note that the
95% CI of the prediction is calculated using uncertainties in estimating kd and solubility from the
correlations shown in Figure 3.5-a) and b). According to the correlations that we developed, the
predicted solubility of CuO NP in the Arizona soil is 96 ±13 mg/kg, with the first order dissolution
rate constant, 0.14±0.07 mg1/3·kg1/3·s-1, whereas the predicted solubility and first order dissolution
rate constant of CuO NP in the Portugal soil are 213±30 mg/kg and 0.68±0.12 mg1/3·kg1/3·s-1
respectively. The pH of the Arizona soil was in the high pH region (where kd is constant), whereas
the pH of the Portugal soil, being in the lower pH region where there is a linear relationship
between kd and pH. To determine the precision of the model predictions, a Monte Carlo simulation
(500 simulations based on the standard deviation and mean of the estimated kd and solubility to
generate a confidence interval) was used to simulate the dissolution kinetics of CuO NP in Arizona
soil and in Portugal soil (Figure 3-6). In both cases, the experimental data fell within 95% CI. This is
evidence that the correlations that we developed can predict the dissolution rate and solubility of
CuO NP in Arizona soil and in Portugal soil based on its pH and organic carbon content. The
experimentally measured CuO NP dissolution was lower than the best fit model prediction in the
Arizona soil and was higher than the best fit model prediction in the Portugal soil. This discrepancy
may be due to our assumption that the composition of SOM and DOM was the same among
different soils. These assumptions contribute to the uncertainties in predicting the solubility and the
dissolution rate constant of CuO NPs in soil. This is because the different chemical composition of
SOM and DOM in various soils may indeed affect the ability of SOM/DOM to complex with Cu or
affect its dissolution rate constant.13
0 1 0 2 0 3 0 4 0
0
5 0
1 0 0
1 5 0
T i m e ( d a y s )
Dis
so
lve
d C
u (
mg
/kg
)
( a )
0 1 0 2 0 3 0 4 0
0
1 0 0
2 0 0
3 0 0
4 0 0
T i m e ( d a y s )
Dis
so
lve
d C
u (
mg
/kg
)
E x p e r i m e n t a l d a t a
P r e d i t i o n h i g h e r 9 5 % C I
P r e d i t i o n l o w e r 9 5 % C I
P r e d i c t e d d i s s o l u t i o n k i n e t i c s
( b )
- 46 -
Figure 3-6. Prediction (predicted dissolution kinetics is the red line, upper 95% CI is the blue line,
lower 95% CI is the purple line) and experimental data (triangle, Cu% dissolved estimated from
DTPA extraction) of CuO NP dissolution in an Arizona soil (a) and in a Portugal soil (b).
3.4 Environmental Implications Previous studies proposed that soil pH is an important factor affecting the dissolution rate of CuO
NP7,8. In this paper, we mechanistically demonstrate that, while soil pH mainly affects how fast the
CuO NPs dissolve, SOM content is the main factor affecting the solubility of CuO NPs in soil (how
much of the NPs can dissolve at equilibrium). The dissolution rate and solubility of CuO NPs
together describe their overall dissolution kinetics in soil. We also found that the soil moisture
content had no impact on the dissolution kinetics of CuO NPs in soil due to the low mass of Cu in
porewater compared to the other soils sinks for Cu (e.g. SOM and mineral surfaces).
With the dissolution model and correlations developed in this study, the dissolution kinetics of the
studied CuO NPs can be predicted from the SOM content and soil pH. This enables modelers to
include the dissolution process of the CuO NP (with certain size, shape and make) in their model
without the need for experimental data or guessing55. Our study also predicted the dissolution rate
and solubility of CuO NPs, which is needed to evaluate their environmental risks56. However, these
results are specific to the CuO NPs studied here. Work is still needed to investigate how the size,
shape and coatings affect the dissolution kinetics of CuO NP in soil. It is possible to incorporate
these properties in the dissolution model. For example, we represent the shape of the particle by the
shape factor, n (Appendix 2, eq A2-3). For a spherical particle or a cube, n=2/3, but for particles
with other shapes, the shape factor would change. The initial size of the particle would affect the
conversion from particle mass concentration to total surface area, which is reflected by C1
(Appendix 2, eq A2-3). The model that we developed in this study for CuO NPs can be potentially
extended to model the dissolution rates of other metal-based ENMs when the time scale for
dissolution is much longer than the time scale of the sorption behavior of ions to soil surfaces. It
should be noted that pH and the amount of SOM were found to be the most important soil
properties affecting the dissolution of CuO NP in soil, but other soil properties could be important
as well. A better understanding of how the composition of SOM and DOM affects the dissolution
of CuO NP could make the model more accurate. With more specific characterization of SOM and
DOM composition, it may indeed be possible to improve the predictive capability of the models.
However, it would require more detailed characterization of SOM and DOM composition as a
- 47 -
trade-off. Our study did not eliminate microbial activity in soil, which could also contribute to the
dissolution of ENMs in soil57,58. However, on a gross level, the impacts of microbial activity seem to
be small relative to the impacts of pH and SOM content. Future models could distinguish between
the microbial and geochemical contributions to dissolution. Our previous study also suggested that
root activity could play a role in dissolution of CuO NP in rhizosphere soil10. Our current model did
not incorporate such influences, but this could be an interesting topic for future research.
Previous studies showed that SOM is the biggest Cu sink in soil, but other soil surfaces, like clay and
iron hydroxide surfaces could be more important sinks for other metals47,49. Our study
demonstrated that, in addition to nanoparticle properties, soil properties should be considered when
predicting the risks or efficiency of ENMs being applied to soil.
This study showed that, unlike the dissolution kinetics in aqueous systems, in soil, pH affects the
dissolution rate constant for CuO NPs, but not their overall solubility . Instead, soil organic matter,
which provides the sink for the dissolved Cu species, was found to control the overall solubility of
CuO NP in soil.
3.5 References of Chapter 3 (1) Tegenaw, A.; Tolaymat, T.; Al-Abed, S.; El Badawy, A.; Luxton, T.; Sorial, G.; Genaidy, A.
Characterization and potential environmental implications of select Cu-based fungicides and bactericides employed in U.S. markets. Env. Sci Technol 2015, 49 (3), 1294–1302.
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(3) Liu, R.; Lal, R. Potentials of engineered nanoparticles as fertilizers for increasing agronomic productions. Sci. Total Environ. 2015, 514, 131–139.
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(5) Simonin, M.; Colman, B. P.; Tang, W.; Judy, J. D.; Anderson, S. M.; Bergemann, C. M.; Rocca, J. D.; Unrine, J.; Cassar, N.; Bernhardt, E. S. Plant and microbial responses to repeated Cu (OH) 2 nanopesticide exposures under different fertilization levels in an agro-ecosystem. Front. Microbiol. 2018, 9, 1769.
(6) Gao, X.; Spielman-Sun, E.; Rodrigues, S. M.; Casman, E. A.; Lowry, G. V. Time and nanoparticle concentration affect the extractability of Cu from CuO NP amended soil. Environ. Sci. Technol. 2017, 51 (4), 2226–2234.
(7) McShane, H. V. A.; Sunahara, G. I.; Whalen, J. K.; Hendershot, W. H. Differences in soil solution chemistry between soils amended with nanosized CuO or Cu reference materials: implications for nanotoxicity tests. Env. Sci Technol 2014, 48 (14), 8135–8142.
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(8) Sekine, R.; Marzouk, E. R.; Khaksar, M.; Scheckel, K. G.; Stegemeier, J. P.; Lowry, G. V; Donner, E.; Lombi, E. Aging of Dissolved Copper and Copper-based Nanoparticles in Five Different Soils: Short-term Kinetics vs. Long-term Fate. J. Environ. Qual. 2017.
(9) Ivask, A.; Juganson, K.; Bondarenko, O.; Mortimer, M.; Aruoja, V.; Kasemets, K.; Blinova, I.; Heinlaan, M.; Slaveykova, V.; Kahru, A. Mechanisms of toxic action of Ag, ZnO and CuO nanoparticles to selected ecotoxicological test organisms and mammalian cells in vitro: a comparative review. Nanotoxicology 2014, 8 (sup1), 57–71.
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(11) McManus, P.; Hortin, J.; Anderson, A. J.; Jacobson, A. R.; Britt, D. W.; Stewart, J.; McLean, J. E. Rhizosphere interactions between copper oxide nanoparticles and wheat root exudates in a sand matrix: Influences on copper bioavailability and uptake. Environ. Toxicol. Chem. 2018, 37 (10), 2619–2632.
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CHAPTER 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil.
Abstract: It has been suggested, but not previously measured, that dissolution kinetics of soluble nanoparticles
such as CuO NPs in soil affect their phytotoxicity. An added complexity is that such dissolution is
also affected by the presence of plant roots. Here, we measured the rate of dissolution of CuO NPs
in bulk soil, and in soil in which wheat plants (Triticum aestivum) were grown under two soil NP
dosing conditions: (a) freshly added CuO NPs (500 mg Cu/kg soil), and (b) CuO NPs aged for 28d
before planting. At the end of the plant growth period (14 days), available Cu was measured in three
different soil compartments: bulk (not associated with roots), loosely attached to roots, and
rhizosphere (soil firmly attached to roots). The labile Cu fraction increased from 17mg/kg to
223mg/kg in fresh treatments and from 283 mg/kg to 305mg/kg in aged treatments over the
growth period due to dissolution. Aging CuO NPs increased the toxicity to Triticum aestivum
(reduction in root maximal length). The presence of roots in the soil had opposite and somewhat
compensatory effects on NP dissolution, as measured in rhizosphere soil. pH increased 0.4 pH units
for fresh NP treatments and 0.6 pH units for aged NPs. This lowered CuO NP dissolution in
rhizosphere soil. Exudates from T. aestivum roots also increased soluble Cu in porewater. CaCl2
extractable Cu concentrations in bulk vs. rhizosphere soil increased from 1.8mg/kg to 6.2mg/kg
(fresh treatment), and from 3.4mg/kg to 5.4mg/kg (aged treatments). Our study correlated CuO NP
dissolution and the resulting Cu ion exposure profile to phytotoxicity, and showed that plant-
induced changes in rhizosphere conditions should be considered when measuring the dissolution of
CuO NP near roots.
This work has been published in Environmental Science & Technology as ‘CuO nanoparticle
dissolution and toxicity to wheat (Triticum aestivum) in rhizosphere soil’ , doi: acs.est.7b05816
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4.1 Introduction
The anticipated benefits of nano-enabled agrochemicals include slow and controlled release of
micronutrients, plant tissue-specific targeted release of micronutrients or pesticides, reduced
amounts of agrochemicals being required, and generally lower toxicity compared to more soluble
products1,2. Copper-based nanoparticles (NPs) are already on the agrochemical market3,4. Copper is
an essential crop micronutrient. Deficiency may lead to reduced disease resistance5 and decreased
crop yields6. However, at high concentrations, Cu can also be toxic to plants,7 the surrounding
microbial communities,8 and soil invertebrates9. Due to its relatively slow dissolution, CuO NPs
have been studied as a potential candidate for agrochemical use. It behaves differently from
dissolved Cu2+ in soil, potentially affecting copper bioavailability, the release of Cu ions over time,
and potential associated risks10–12. However, the connection between NP dissolution, the resulting
dose of Cu ions and its toxicity to terrestrial plants, and the role of root exudates on this process
have not been well elucidated due to a lack of appropriate characterization of the dissolution of the
NPs in soil. Ideally, application rates of these novel materials should be based on their fate and
effects in the terrestrial environment, their bioavailability and potential toxicity to plants. The toxic
effect of Cu species is reflected in physiological changes in plant roots and shoots, such as decreased
root length, increased root compactness, change in root color, shorter leaf length and decreased
shoot biomass13–15. Hyperspectral imaging has been used to visualize NPs in plants and to confirm
macroscopic evidence of NP toxicity16,17.
Previous studies of the toxicity of CuO NPs to terrestrial plants assumed, but did not
measure, dissolution behavior of CuO NP in soil. This has led to conflicting conclusions on the
toxicity of CuO NPs. While some studies attributed the toxic effect of CuO NP to released ionic
Cu15,18,19, others concluded the opposite20. For example, Servin et al. chose a Cu ion control
concentration based on the assumption that only 10% of the CuO NP would dissolve in soil, the
same fraction that dissolved in pure sand, rather than measuring CuO dissolution in soil. They
concluded that dissolution of CuO NPs could not fully explain the plant toxicity because the plant
responses differed from their Cu ion control .20 Much more than 10% CuO NP could have
dissolved in soil because soil organic matter (SOM) acts as a Cu sink, increasing the amount of CuO
NP that can be dissolved11. This weakens their conclusions about a NP-specific effect. Similar
problems occurred in other studies 15,18,21–23. Breaking with this trend, Dimkpa et al. (2013) evaluated
the total CuO NP dissolved in soil using a water-extraction method. 24 Unfortunately, the water-
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extraction does not extract Cu bound to the soil solid matrix which accounts for most of the
dissolved Cu in soils25–27. Thus, their assertion of a CuO NP-specific toxicity in soil is confounded by
the potential that more Cu had dissolved than was assumed or measured. Recently Qiu et al. found
that the toxicity of CuO NP, CuO bulk particles and soluble Cu (Cu(AC)2) depends on their
solubility in soil, and that the distinction in solubility diminished after a 90-day aging period.
However, the actual dissolution during the incubation periods (one day vs. 90 day) was not
quantified. They successfully correlated the toxicity of NP to roots of Hordeum vulgare L. (5-day root
elongation experiment) with ‘free Cu ions’ in soil pore water measured at a single time point before
seeding;19 though convincing, it should be noted that the dissolution during the 5-day toxicity test
was not considered. While the relatively slow dissolution of CuO NP may result in unobservable
impacts on toxicity during a relatively short 5-day toxicity test, dissolution at this rate would
probably affect toxicity of NPs in longer tests.
The dissolution of CuO NPs is a dissolution rate-limited process. Experimental approaches,
such as extraction with CaCl2 or with diethylenetriaminepentaacetic acid (DTPA), have been used to
predict the bioavailability or toxicity of metals in soil.28–31 CaCl2 extracts the Cu ions in soil pore
water that are considered ‘readily available,’ while DTPA extracts the “labile” fraction including
dissolved Cu in soil pore water (free Cu2+and Cu2+ complexed with soluble ligands such as dissolved
organic matter (DOM)), but also the Cu2+ associated with soil solid phases, such as soil organic
matter (SOM), clay particles, and iron oxide minerals.29–31 Whereas CaCl2 extracts metals that are
‘readily available’ to plants29, DTPA extracts this pool as well as the pool that may eventually
become bioavailable in soil, the so called ‘potentially available’ fraction32. One problem with using
these extraction methods to predict the bioavailability of Cu based nanomaterials is that a single time
point extraction does not capture the temporal dynamics of the CuO NP dissolution process. Our
recent study used extraction methods at different times to monitor the kinetics of release of Cu ions
from CuO NP in soil. In that study, the increase in DTPA extractable Cu over 30 days in soils was
used to estimate the dissolved pool of Cu in soil.11 The availability of Cu ions increased with time
over a 30d period, which may explain why previous efforts to correlate the extractable metals in
metal-based NP-amended soils with their bioavailability or toxicity have generally failed33–35.
Plants also may affect the dissolution behavior and availability of CuO NP in soil, especially
in the rhizosphere. Previous studies using extraction methods to predict the bioavailability or toxicity
of metal-based ENMs or the dissolution of ENMs in soil did not typically consider the impact of
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roots on Cu availability.11,12,33–35 Plant roots exude organic acids 36–38 that may affect the pH in
rhizosphere. 39,40 Although soil pH and organic carbon are known to be important factors
influencing the dissolution behavior of CuO NPs in soil11,12, and previous studies have proposed that
exudates from plant roots may affect the dissolution of CuO NP in the rhizosphere41, no studies
have quantified this. Given that the rhizosphere is where plants interact with soil for nutrient
uptake,42,43 a better understanding of how the roots impact NP dissolution and metal availability in
the rhizosphere is needed to design nano-enabled agrichemicals with optimal properties for
delivering nutrients.
The objectives of this study are to quantify the influence of time and near-root chemical
conditions on dissolution and lability of CuO NPs in rhizosphere soil, and to determine the
influence of this dissolution on the toxicity of CuO NPs to Triticum aestivum during a 14-day plant
growth period in soil. Wheat (Triticum aestivum) was used in this study because it is the 2nd most
cultivated plant in the world, and it is sensitive to Cu deficiency44 or excess45. To evaluate the toxicity
of CuO NP to plants, we measured the dissolution behavior of CuO NPs in soil in the presence of
plants with emphasis on the soil-plant interface (rhizosphere) where roots interact with soil. The
toxicity of Cu was evaluated by physiological changes in plant roots and shoots.
4.2 Method and Materials
4.2.1 Chemicals
Calcium chloride (≥99.0%, ACS grade), calcium oxide (CaO), calcium carbonate (CaCO3) (99%+),
and hydrogen peroxide (30%, certified ACS) were purchased from Fisher Scientific. DTPA (>99%)
and triethanolamine (TEA, ≥99.0% (GC)) were purchased from Sigma-Aldrich. Trace metal grade
nitric acid (65%-70%) was purchased from VWR. Copper sulfate (CuSO4) was purchased from
Fisher Scientific . Lufa Standard soils (2.1, 2.2, 2.4 and 2.4) were purchased from Lufa Speyer,
Germany. A calcareous soil (pH 7.6) was collected in Arizona (termed Arizona soil) and used to test
the model’s ability to predict CuO NP dissolution behavior based on soil pH and SOM content.
Another more acidic soil (pH=5.0) was collected from a grassland in northwestern Portugal (termed
Portugal soil). Detailed properties of all the soils used can be found in appendix ( Table A3-1).
Calcium chloride (≥99.0%, ACS grade) and hydrogen peroxide (30%, certified ACS) were
purchased from Fisher Scientific. DTPA (>99%) and triethanolamine (TEA, ≥99.0% (GC)) were
- 56 -
purchased from Sigma-Aldrich. Trace metal grade nitric acid (65%-70%) was purchased from VWR.
Triticum aestivum seeds (Pembroke 2014) were bred by Dr. David Van Sanford (Department of Plant
and Soil Sciences, University of Kentucky).
4.2.2Nanoparticle Characterization
CuO NPs (~40 nm primary particle size), were purchased from Sigma-Aldrich. The primary size of
particles, zeta potential, isoelectric point and hydrodynamic diameter have been characterized and
reported in Chapter 26.
4.2.3 Soils and Characterization of Soil Properties Standard Lufa 2.2 soil (loamy sand) was purchased from Lufa Speyer, Germany. Lufa 2.2 soil
contains 1.6 wt. % soil organic matter, and little total and available Cu (see Appendix 3, Table A3-1
and Table A3-2, control treatment). Using a well-characterized standard soil allows comparisons
between studies. The high carbon organic content (about 1.6%) of Lufa 2.2 makes this soil good for
agricultural studies. Soil was air dried and sieved < 2mm before shipping. The soil was further air-
dried for at least 24 hours before all experiments. Soil pH in different treatments was determined by
the CaCl2 extraction method (see ‘Extraction methods’ section). Soil moisture content (1% for the
air dried soil) was determined gravimetrically after oven-drying the soil at 105 ºC for 24 h46. Soil field
moisture capacity (21%) was determined using a Haines apparatus with 0.1 bar pressure difference
between the wet soil and the atmosphere. Soil was in equilibrium with air and not water saturated,
and presumed to be aerobic for the duration of the experiment (~+400mV)
4.2.4 Soil amendment. The CuO NP suspension (containing Na2SO4), CuSO4 solution, or Na2SO4 solution (control
treatment) were mixed with soil and brought to a moisture content of 21.7% (corresponding to
~50% of the water holding capacity). The soil was mixed with wooden sticks in a beaker for 20min.
The homogeneity was confirmed with the low standard deviation for the total Cu content measured
by soil digestion data (Appendix 3, table A3-1). To test if CuO NP and CuSO4 treatments resulted in
different Cu bioavailability and toxicity, the Cu ion concentration had to be high enough to ensure
some CuO NPs remained in the soil during the study period. We chose 500mg/kg (as Cu) for the
CuO NP treatment, and 300 mg/kg (as Cu) for the CuSO4 treatment based on a preliminary study
to assess the solubility of the CuO NPs in the Lufa 2.2 soil (Appendix 3, Figure A3-1). The results
showed that the solubility of CuO NPs in Lufa 2.2 soil was ~300mg/kg. Therefore, the selected
- 57 -
concentrations provided a similar concentration of added Cu ion in both treatments after one
month.
4.2.5 Germination and plant growth
The seeds of Triticum aestivum were surface sterilized by submerging them in 10% sodium
hypochlorite solution for 10 minutes, and then washed with DI water three times. The seeds were
then kept immersed in DI water overnight on an end-to-end rotator. The following day, the seeds
were transferred to a petri-dish containing moist tissue paper. The petri-dishes were covered with
aluminum foil and incubated in the growth chamber for 7 days, until 90% germination was achieved.
Germinated seeds were transplanted into syringes containing 120g of amended soils either
immediately after adding the Cu (fresh treatment) or 28 days after the Cu was added (aged
treatment).The plants were incubated in a growth chamber with constant moisture content and 16h-
light/8h-dark cycle (25 °C for daytime and 21 °C for night time). A diluted Cu-free Hoagland
solution (quarter strength) was added (1ml/day) to each syringe to maintain the moisture content of
the soil as well as provide nutrients to plants. The concentration of Cu in soil and plant tissue was
determined using a standard digestion method (EPA Method 3050b47) and ICP-MS analysis of Cu in
the digestate. See appendix. Adding moisture content did not induce any vertical transport of Cu, as
suggested by Figure A3-4 in Appendix 3.
4.2.6 Sampling of soil and plant tissue Prior to transplanting the germinated seeds in soils, subsamples of each soil were collected from all
treatments for DTPA extraction (2g of soil per extraction) to measure the labile metal fraction. After
14d of growth, rhizosphere soil, "loosely attached soil," and bulk soil (Appendix 3, Figure A3-2)
were collected for DTPA and CaCl2 extraction to determine the total dissolved metal and readily
available metal, respectively, as described below. After the plants and roots were removed from the
syringe, the soil remaining inside the syringe was defined as bulk soil, presumably minimally affected
by the plant roots. The bottom 5mm of bulk soil was also collected to determine if there was
significant vertical transport of Cu. No vertical transport of Cu was observed (Appendix 3, Figure
A3-4). The roots were separated from shoots. Both roots and shoots were photographed with a
scale bar for determination of length. For each treatment, one plant root replicate was washed with
1mM KCl three times for Cytoviva analysis (described below). The remaining roots were shaken by
hand in a 50 ml centrifuge tube, and the soil that detached during shaking was defined as loosely
attached soil48 (Appendix 3, Figure A3-2). After shaking, the roots were placed on aluminum foil and
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air dried in a fume hood for 24 hours. The roots were then shaken again in a 50 ml centrifuge tube,
and the soil that detached during the air-drying process and the second shaking process was defined
as rhizosphere soil49. Due to the small amount of rhizosphere soil collected per treatment, not all
replicates were suitable for DTPA and CaCl2 extraction. The details of which samples were analyzed
can be found in appendix (Table A3-3 and Table A3-4). For CuSO4 treatments, the roots were
highly compacted, precluding the collection of rhizosphere soil. The shoots and roots were oven-
dried at 105 ºC for 24 h. The mass of the dried roots and shoots was recorded before digestion for
total Cu analysis (details in Appendix C).
4.2.7 Soil extraction
Two standard extraction fluids were used in this study. DTPA extractant was composed of 0.01M
CaCl2, 0.005M DTPA and 0.1M triethanolamine (TEA) (pH=7.6). CaCl2 extractant was 0.01M CaCl2
without pH adjustment. All extractions were done using a reciprocal shaker at 180 rpm for 2 hours.
It is important to note that the centrifuge tubes were laid horizontally in the shaker rather than
vertically to provide the best extraction efficiency. For soil samples collected before the plant growth
experiments, 2g of soil were extracted with 4ml DTPA extractant. For bulk soil samples, loosely
attached soil samples and rhizosphere soil samples, 0.4g of soil was extracted with 0.8ml DTPA
extractant, while 0.35g of soil was extracted with 3.5ml CaCl2 extractant. After extraction, all samples
were centrifuged at 3000 rpm for 10 min, and the supernatants were filtered using a 0.45 µm PTFE
filter. The pH of the CaCl2 extracts for each soil fraction was measured (Figure 5). All samples were
acidified with 20% HNO3 (final HNO3 concentration, 2%) and Milli-Q-water before being analyzed
by ICP-MS. The method for ICP-MS is provided in detail in the appendix C.
4.2.8 Cytoviva analysis
The interaction between roots and NPs were visualized in fresh roots after a rinsing step in 10-3 M
KCl, using a darkfield-based hyperspectral imaging (DF-HSI) system (CytoViva Inc., USA). See
appendix C for additional details.
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4.3 Results and Discussion
4.3.1 Nanoparticle characterization
The properties of the CuO NPs have been previously described11. Briefly, the primary particle size
was determined by TEM to be 38nm ± 1.7nm (278 particles counted, 95% CI). The hydrodynamic
diameter of an 80mg/kg CuO NP in an aqueous 5mM NaHCO3 suspension (pH=7) was measured
to be 560nm±103nm (3 replicates, 95% CI, intensity averaged), and the zeta potential was -
16.1mV±1.7mV (3 replicates, 95%CI) in the same suspension. The pH of the isoelectric point
(pHiep) of the CuO NPs in a 5mM NaNO3 solution was 8.8, while the pHiep was 5.8 in the 5mM
NaHCO3 solution11. The hydrodynamic diameter and zeta potential likely change after they are
added to the soils due to interactions with soil components such as natural organic matter and
calcium50,51.
4.3.2 Change in extractability of Cu in bulk soil during the plant growth experiment.
The DTPA extractable Cu in the bulk soil for CuO NP and CuSO4 amended soils are shown in
Figure 4-1. The DTPA extractable Cu represents the Cu that was released from the CuO NPs
during the treatment. The extractable Cu vs. time is shown for the 14d growth period for both the
freshly added CuO NPs, and for the aged CuO NPs, where plants were added after the CuO NPs
had aged for 28d prior to planting the germinated seeds. The total Cu concentration in the two
treatments and in the control (unamended) soil is provided in the appendix (Table A3-1).
For the CuO NP treatments, the DTPA extractable Cu from bulk soil increased over time
(Figure 4-1a) (ANOVA test, P<0.05), although the increase was much higher for fresh CuO NP
treatment than that for the aged CuO NP treatment. The DTPA extractable Cu in bulk soil collected
form these plants was measured at four points (just after mixing, t=14, 28, and 42d) during the
study. The dissolution rate fit a first-order dissolution model well (R2=0.990). The aged CuSO4
treatments showed the opposite trend, with DTPA extractable Cu slightly decreasing over time
(Figure 4-1b). The DTPA extractable Cu in soils from the control treatment (Na2SO4) was low, and
no change was observed during the 2-week plant growth experiments (Appendix 3, Table A3-2).
At the end of the 14d plant growth period for the fresh CuO NP treatment, the DTPA extractable
Cu was similar to the aged CuSO4 treatments (ANOVA test, P>0.05). The DTPA extractable Cu in
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the aged CuO NP treatment was statistically significantly higher than both fresh and aged CuSO4
treatment (ANOVA test followed by Fisher’s LSD test for multiple comparison, P≤0.05) (Figure 4-
1c). The CaCl2 extractable Cu revealed a different order, with fresh CuSO4 treatment having the
highest CaCl2 extractable Cu, followed by the aged CuO NP treatment and the aged CuSO4
treatment, with the fresh CuO NP treatment having the lowest amount of CaCl2 extractable Cu. The
CaCl2 extractable copper represents the “readily available” Cu in the porewater.
Figure 4-1. Change in DTPA extractable Cu over time for each treatment: a) CuO NP treatment, b)
CuSO4 treatment, and comparison of mean of extractable Cu for each Cu treatments at the end of
the plant growth period: c) DTPA extraction, d) CaCl2 extraction. Error bars show ± 1 standard
deviation. In a and b, capital letters indicate significant differences between DTPA extractable Cu at
four time points for CuO NP treatments (a) and CuSO4 treatments (b). In c and d, capital letters
indicate significant differences in DTPA extractable Cu (c) and CaCl2 extractable Cu (d) among soils
collected after plant harvesting (ANOVA test followed by Fisher’s LSD test for multiple
comparisons, P≤0.05).
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4.3.3 Toxicity of CuSO4 and CuO NP.
Root maximal length, root compactness (root mass/root maximal length), leaf lengths, shoot mass
(Figure 4-2) and root morphology (Appendix 3, Figure A3-5 and Figure A3-6) were used to evaluate
the toxic effect of CuSO4 and CuO NPs.
Root maximal length and root compactness indicated no visual toxic effect from the fresh
CuO NP treatment. For other treatments, significant decreases in root maximal length (a decrease of
6.6cm, 8.2cm, and 6.8cm compared to the control treatment for aged CuO NPs, fresh CuSO4, and
aged CuSO4, respectively) were observed. Increased root compactness was observed for the aged
CuO NP treatment (an increase of 4.0 mg/cm compared to the control) and for the fresh CuSO4
treatment (an increase of 5.1 mg/cm compared to the control). Examples of shortened roots and
compactness of roots are shown in the appendix (Figure A3-5, Appendix 3). Evidence of Cu toxicity
was also observed in Cytoviva images. In comparison to the roots exposed to CuSO4 (Appendix 3,
Figure A3-6), the roots exposed to CuO NP (fresh or aged) did not present the same damaged
physiology. Roots exposed to CuSO4 (both fresh and aged) showed a brown damaged (necrotic)
zone that was not found on any of the CuO NP exposed roots. No effects of Cu on the shoots (leaf
length, biomass) were observed for the CuO NP treatments. Both the freshly amended and aged
CuSO4 treatments resulted in shorter third leaves (shortened by 5.4cm and 4.0 cm compared to the
control for fresh and aged CuSO4 treatments, respectively). The freshly amended CuSO4 treatment
also had less total shoot biomass compared to the control treatment.
Some indication of toxicity was evident in all treatments except for the fresh CuO NP
treatment. Aging of CuSO4 decreased its toxicity to Triticum aestivum, while aging of CuO NP
increased its toxicity. Overall, the two CuSO4 treatments showed more toxic effects to Triticum
aestivum compared with the two CuO NP treatments, even though the CuSO4 was added at a
significantly lower Cu concentration (300mg/kg for CuSO4 treatments vs. 500mg/kg for CuO NP
treatment).
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Figure 4-2. a) Root compactness and b) leaf length (leaf growth stage is noted with number, from 1
being the youngest to 3 the oldest) of wheat seedlings grown in freshly amended and aged CuO NP,
CuSO4-amended soil, and control treatments. Error bars show ±1SD, * indicates P≤0.05; **
indicates P≤0.01. (ANOVA test followed by Fisher’s LSD test for multipal comparisions) compared
to the control treatment.
4.3.4 Cu root association and Cu uptake
The presence of CuO NPs associated with the roots after 2 weeks of plant growth in both fresh and
aged CuO NP amended soils was investigated using enhanced dark-field hyperspectral imaging (DF-
HSI) as shown in Figure 4-3. The pixels containing CuO NP have been highlighted in red. In both
fresh and aged CuO NP amended soils (Figure 4-3), CuO NPs were found associated to specific
locations on the roots, either to the root tip mucilage (Figure3 a, b, f, g), or to soil aggregates
attached to the root hairs or root tips (Figure 4-3 a, c-i). For the concentration of Cu in roots, all Cu
treatments were significantly higher than the control treatment. The Cu concentration in roots
(577mg/kg, s.d.=46mg/kg, 6 replicates) was statistically significantly higher in the freshly amended
CuO NP treatment than in the aged CuO NP treatment (400mg/kg, s.d.=60mg/kg, 6 replicates) or
either ionic treatment (278mg/kg, s.d.=51mg/kg, 6 replicates for fresh CuSO4 and 442mg/kg,
s.d.=67mg/kg, 6 replicates for aged CuSO4) (Appendix 3, Figure A3-7, a). For the shoot
concentrations, no statistically significant differences were found for all Cu treatments (53mg/kg-
88mg/kg) (Appendix 3, Figure A3-7, b).
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Figure 4-3. Hyperspectral imaging of plant roots grown in soil with freshly amended CuO NPs (a-e)
or after aging (f-i). The b, c and g views are magnified views from a and f. Pixels containing the
reflectance spectra specific to CuO NPs are highlighted in red. CuO NPs and their aggregates were
found associated to mucilage, root tissues and root hairs (red arrow), and to soil aggregates attached
to those locations (yellow arrows). Scale bars: 25µm.
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4.3.5 Effect of near-root environment on Cu availability from CuO NP treatment Figure 4-4 shows the differences in extractable Cu in rhizosphere soil, loosely attached soil and bulk
soil for fresh and aged CuO NP treatments. For the CaCl2 extraction in both fresh and aged CuO
NP treatments (Figure 4-4a, b), the extractability of Cu in the rhizosphere soil was significantly
higher than the extractability of Cu in the loosely attached soil or bulk soil (ANOVA test, P≤0.05).
There were no statistically significant differences among DTPA extractable Cu measurements from
rhizosphere soil, loosely attached soil and bulk soil in the freshly amended CuO NP treatment
(Figure 4-4c). However, the DTPA extractable Cu in the rhizosphere soil in the aged CuO treatment
was significantly lower than the DTPA extractable Cu in bulk soil, but similar to that measured in
loosely attached soil (Figure 4d). In control experiments (Na2SO4), the CaCl2 extractable Cu was
below the detection limit (0.08mg/kg in soil, 4ug/L for the diluted samples) in all soil samples.
Aging increased the concentration of CaCl2 extractable Cu and DTPA extractable Cu in loosely
attached soil and bulk soil, and increased the concentration of DTPA extractable Cu in rhizosphere
soil (t test, P<0.05). But aging did not change the CaCl2 extractable Cu in rhizosphere soil (t test,
P>0.05).
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Figure 4-4. CaCl2 and DTPA extractable Cu in fresh (left side) and aged (right side) CuO NP
amended rhizosphere soil, loosely attached soil and bulk soils. Error bars show ± 1 SD. Capital
letters indicate significant differences between groups (One way ANOVA test followed by Fisher’s
LSD test for multiple comparison, P≤0.05).
4.3.6 Soil pH in bulk soil, rhizosphere soil and loosely attached soil
For all CuO NP treatments and the control treatment (no addition), the pH of the rhizosphere soils
was significantly higher than the pH of the loosely attached and bulk soils (Figure 4-5a, b and c). In
freshly amended CuO NP treatments and control treatments, the pH of the loosely attached soils
were not statistically significantly different than the pH in the bulk soils. However, in aged CuO NP
treatments, the pH of the loosely attached soil was statistically significantly higher than the pH in
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bulk soil. In bulk soil, the pH was the highest in freshly amended CuO NP treatments, followed by
aged CuO NP treatment, followed by the control treatment, followed by the aged CuSO4 treatment,
with freshly amended CuSO4 treatment having the lowest soil pH (Figure 4-5d).
Figure 4-5. Mean ± SD of soil pH (measured using CaCl2 extraction) in rhizosphere soil, loosely
attached soil and bulk soil in a) soil freshly amended with CuO NP, b) aged CuO NP treatment c)
control soil, and; d) Comparison of pH of bulk soil among all treatments. Capital letters indicate
significant differences (ANOVA test followed by Fisher’s LSD test for multiple comparison,
P≤0.05).
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4.4 Discussion
4.4.1 CuO NP dissolution is linked to toxicity.
Compared to Cu ions, the dynamic dissolution process of CuO NP in soil led to a very different Cu
exposure profile for plants. At the end of the two growth periods, the DTPA-extracted Cu
concentrations in CuO NP treatments were similar or even higher than in the CuSO4 treatment.
However, a decreasing trend in DTPA extractable Cu on CuSO4 treatments during the two plant
growth periods was observed. This decrease can be attributed to the soil-organic matter interactions,
solid-state diffusion of Cu ions into iron minerals or metal (co)precipitates52–54. Conversely, an
increase in DTPA-extractable Cu over time was shown in CuO NP treatments (fresh treatment and
aged treatment) during the two plant growth periods and the aging period. This can be attributed to
the dissolution of CuO NP11,12. Thus, the plants in the freshly amended CuO NP soil were exposed
to lower amounts of labile Cu compared to either the two CuSO4 treatments or the aged CuO NP
treatment. These findings suggest that when evaluating the chemical availability or toxic effect of
metal-based NPs in soil, single-time point chemical extractions at the end of the plant growth period
cannot capture the dissolution process of NPs in soil, and thus may fail to predict the toxicity or
bioavailability of NPs11,12,55. Considering that it is not feasible to uniformly dose Cu ions into soil
over time to precisely mimic the dosing rate from NP dissolution, toxicity studies with soluble NPs
should measure the dissolution rate in soil and monitor the behavior of soluble ions in soil, and
interpret their results in light of the different dosing conditions that manifest.
A significantly higher toxicity in CuSO4 treatments compared to the fresh CuO NP
treatment is explained by the higher exposure of roots to labile Cu species, even though the CuO
NP treatment had a higher total Cu concentration. Also, dissolution of CuO NPs over time
gradually increased the available Cu in soil, leading to higher toxicity in the aged treatment. The
opposite has been observed with CuSO4, where the available Cu in soil decreased over time, leading
to lower toxicity to the plants in the aged treatment. The effects of time on toxicity of CuO NP and
CuSO4 have already been observed19. The authors attributed this to the dissolution behavior of CuO
NP, although without quantification. Here, we clearly showed that in order to correlate the chemical
availability of CuO NPs with toxicity, the dissolution kinetics, i.e. predicting the total Cu released to
soil during the growth period, should be considered. The dissolution kinetics can be modeled as
first-order dissolution, with the rate constant fit to the extractable Cu over time11, and the total
amount of Cu ion released from CuO NPs can be estimated by integrating the expression relating
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the change of extractable Cu over time. This observation is relevant to NP formulations of
fungicides and micronutrients, so the release rate of the active ingredients can be better timed to the
plant’s needs.
4.4.2 CaCl2 extractable Cu correlates with toxicity of CuO NP to wheat.
Although DTPA extractable Cu gives a better indication of CuO NP dissolution because it extracts
most of the Cu species in soil, CaCl2 extractable Cu is better for correlating toxicity, since it
measures dissolved Cu in pore water that can directly interact with plant roots. DTPA extraction
would predict the toxicity of the aged CuO NP to be higher than the CuSO4 treatment (Figure 4-1 a,
b). However, this was not the case. The aged CuO NP had lower toxicity compared to both the aged
and the fresh CuSO4 treatment, indicating that the DTPA extraction cannot accurately predict
toxicity for the CuO NPs. The CaCl2 extraction ranked them correctly (Figure 1d). Considering that
the extractable Cu in CuO NP amended soil increased over time while the extractable Cu in CuSO4
amended soil decreased over time in Lufa 2.2 soil (Appendix 3, Figure A3-1), the wheat plants were
exposed to a lower overall ‘readily available’ Cu (i.e. Concentration x time) in CuO NP treatments
compared to the CuSO4 treatments. The lower CaCl2 extractable Cu in aged CuO NP treatment is a
result of higher soil pH in aged CuO NP treatment compared to the fresh and aged CuSO4
treatment. Higher soil pH has been previously shown to lower Cu concentration in soil pore water 27,56.
4.4.3 Root-associated CuO NP modulates toxicity. In the freshly amended CuO NP soil, although being exposed to a lower concentration of labile Cu,
the roots of Triticum aestivum were actually exposed to higher total Cu concentration (Appendix3,
Figure A3-7) than the other treatments. This is mainly due to CuO NPs' association with plant roots
(Figure 4-3b). This exposure to CuO NPs did not lead to any detected toxic effects, indicating a low
or de minimis level of toxic effects from the particle itself over the 14d growth period.
4.4.4 Root exudates affect CuO NP dissolution and availability. The increase in the pH of rhizosphere soil compared to bulk soil in our study indicates that the
rhizosphere region was indeed influenced by the plant roots. Excretion of organic acid (dissociated
ions) by plant roots , nitrogen uptake and ionic exchanges by plant roots may explain the higher pH
of the rhizosphere soil compared to the pH of bulk soi39,40,57,58. The observed pH change in
rhizosphere soil was not likely a result from the presence of CuO NP, as similar pH changes
occurred with both the fresh CuO NP treatment and the negative control treatment (0.4pH unit,
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ANOVA test, P>0.05). However, in the aged CuO NP treatment, the pH increase was higher
(0.6pH unit, ANOVA test, P<0.05 ), potentially because the release of Cu2+ from CuO NP attached
to the roots triggered more root responses. A previous study has shown that a high concentration of
ionic Cu can increase root exudation59. The other factor that may affect CuO NP dissolution besides
pH is the organic acids released by plants. Although we did not measure them here, their release is a
well-known mechanism by which Triticum aestivum increases the availability of nutrients and
decreases the toxic effects of metal ions such as Cu and Al36–38.
Influences of rhizosphere soil pH and root exudates on the dissolution of the CuO NPs and on the
lability of Cu derived from CuO NP were revealed by the two extraction methods (Figure 4-4).
Though rhizosphere soil exhibited higher pH (which should reduce the concentration of free Cu
ions), more Cu was extracted by CaCl2 from rhizosphere soil relative to bulk soil, indicating a greater
amount of complexed Cu ions in rhizosphere soil pore water. This complexation is likely a result of
small organic acids released by plants and their microbiomes in the rhizosphere region. The
influence of the roots on the extent of dissolution of CuO NPs was shown by the extraction of
“labile” Cu (DTPA). For the aged CuO NP treatment, the DTPA extractable Cu in rhizosphere soil
was lower than for the bulk soil. DTPA extraction has been shown to extract most of the labile Cu
species in soil (~80%), but it cannot extract CuO NP11(Appendix 3, Figure A3-1). Thus, a reduction
in DTPA extractable Cu suggests diminished CuO NP dissolution in rhizosphere soil. This is
consistent with the slightly higher measured pH in rhizosphere soil compared to bulk soil (Figure 4-
5), which decreases the dissolution rate and solubility of CuO NP. 10,12 This difference is less evident
in the fresh CuO NP treatment than the aged CuO NP treatment, consistent with the slightly lower
rhizosphere soil pH in the fresh CuO NP treatment compared to the aged CuO NP treatment
(Figure 4-5).
For CuO NP treatment, the DTPA extractable Cu in all soil compartments (rhizosphere soil, loosely
attached soil and bulk soil) increased from the fresh treatment to the aged treatment as a result of
CuO NP dissolution. This dissolution also resulted in higher ‘readily available’ CaCl2 extractable in
loosely attached soil and bulk soil. However, in rhizosphere soil, the impact of CuO NP dissolution
on the ‘readily available’ Cu was less significant, possibly because complexation by root exudates
played a more important role in increasing the ‘readily available’ Cu in rhizosphere soil.
The interaction between plant, soil, and Cu was limited to the rhizosphere soil region during
the 14-d plant growth period, as suggested from the measured pH and extractable Cu in different
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soil zones. This finding is consistent with a previous study showing the pH was affected by durum
wheat roots only within a few millimeters60. Even the loosely attached soil collected in this study
remained mostly unaffected by the presence of plants. However, the spatial extent where root
exudates may affect the dissolution of CuO NPs over longer growth periods needs more
investigation. The changes in soil pH and the release of plant exudates changed the chemical
availability and the dissolution rate of CuO NPs in rhizosphere soil. Thus, measuring the changes of
CaCl2 extractable Cu over time in rhizosphere soil may be the best way to correlate CuO NP
dissolution with its toxicity. Further study is needed to quantitatively evaluate the impact of root
exudates and changes in pH on dissolution behavior of CuO NP in rhizosphere soil over time to
decide whether rhizosphere soil should routinely be used in extraction tests to predict the toxicity of
CuO NPs to terrestrial plants. Bacteria in root rhizosphere may also play a role in CuO NP
availability, dissolution and uptake via exuding chelating compounds like pyoverdines (siderophores
produced by certain pseudomonads that can complex with metal ions). Future studies needs to
elucidate the role of these bacteria have in rhizosphere region61,62. However, these experiments
indicate that measurements of Cu availability in bulk soils will not likely provide an accurate
representation of the Cu availability to the plants.
4.4.5 Triticum aestivum regulated Cu uptake. Although we have shown a strong correlation between CuO NP dissolution and the toxicity of
CuO NP to Triticum aestivum, there was no such correlation between CuO NP dissolution and Cu
translocation in Triticum aestivum shoots. Despite being exposed to different concentrations of total
and labile Cu species during the growth periods among all treatments, as shown in Figures 1a and b,
Triticum aestivum tended to take up similar amounts of Cu (Appendix 3, Figure A3-7). This suggests
that the uptake of Cu was regulated by Triticum aestivum, consistent with a previous study that found
low Cu uptake by wheat even when the labile Cu concentration in soil was high63. When Cu uptake
is regulated by a plant, the bioavailable Cu in soil measured by extraction cannot be correlated to Cu
uptake.
4.5 Agricultural implications Our study indicates some potential benefits of using nano-CuO as a micronutrient amendment or
fungicide rather than the soluble Cu salt. Firstly, CuO NPs were less toxic than CuSO4, despite being
applied at a much higher dose to the soils. The slow dissolution of CuO NPs reduced the maximum
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concentration of CaCl2-extractable Cu experienced by plants, providing a continuous release of Cu
over 14 days without showing any visual toxic effect. After aging for 28d, some Cu phytotoxic
effects were observed. However, the dissolution rate of a ‘nano-enabled’ fertilizer or fungicide could
likely be tuned to provide sustained release the Cu ions at a rate where the concentration of Cu ions
would not exceed the phytotoxic concentration22. This tuning could potentially be accomplished
through surface modification, addition of other phases (e.g. via doping), using mixtures of different
sized particles, or adjusting their size or aspect ratio.
Secondly, we have shown that CuO NPs have higher affinity to plant roots than Cu ions. Thus,
‘targeting’ the NPs to plant roots could be another potential benefit for nano-enabled fertilizer. In a
calcareous soil with high pH, the dissolution of CuO NP in bulk soil would be very low, but plant
exudates could still potentially enhance the dissolution behavior of CuO NP and the availability of
Cu in the rhizosphere soil, avoiding the toxicity of excess of Cu. This may also reduce Cu
accumulation in soil, a problem with many Cu-containing fertilizers/pesticides. The improvements
in nutrient uptake efficiency or antifungal properties due to this ‘targeting’ requires more
investigation.
Conceivably, nano-enabled, slow-release fertilizers could thus be designed to be applied at high
concentration but low toxicity, to last for years without re-application. This would save energy and
labor, incentives to growers to adopt those new technologies. Research is still needed to determine
the rate of delivery that can provide its function effectively, but without invoking toxicity.
4.6 References of Chapter 4 (1) Tegenaw, A.; Tolaymat, T.; Al-Abed, S.; El Badawy, A.; Luxton, T.; Sorial, G.; Genaidy, A.
Characterization and potential environmental implications of select Cu-based fungicides and bactericides employed in U.S. markets. Env. Sci Technol 2015, 49 (3), 1294–1302.
(2) Giannousi, K.; Avramidis, I.; Dendrinou-Samara, C. Synthesis, characterization and evaluation of copper based nanoparticles as agrochemicals against Phytophthora infestans. RSC Adv. 2013, 3 (44), 21743–21752.
(3) Liu, R.; Lal, R. Potentials of engineered nanoparticles as fertilizers for increasing agronomic productions. Sci. Total Environ. 2015, 514, 131–139.
(4) Elmer, W. H.; White, J. C. The use of metallic oxide nanoparticles to enhance growth of tomatoes and eggplants in disease infested soil or soilless medium. Environ. Sci. Nano 2016.
(5) Simonin, M.; Colman, B. P.; Tang, W.; Judy, J. D.; Anderson, S. M.; Bergemann, C. M.; Rocca, J. D.; Unrine, J.; Cassar, N.; Bernhardt, E. S. Plant and microbial responses to repeated Cu (OH) 2 nanopesticide exposures under different fertilization levels in an agro-ecosystem. Front. Microbiol. 2018, 9, 1769.
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(6) Gao, X.; Spielman-Sun, E.; Rodrigues, S. M.; Casman, E. A.; Lowry, G. V. Time and nanoparticle concentration affect the extractability of Cu from CuO NP amended soil. Environ. Sci. Technol. 2017, 51 (4), 2226–2234.
(7) McShane, H. V. A.; Sunahara, G. I.; Whalen, J. K.; Hendershot, W. H. Differences in soil solution chemistry between soils amended with nanosized CuO or Cu reference materials: implications for nanotoxicity tests. Env. Sci Technol 2014, 48 (14), 8135–8142.
(8) Sekine, R.; Marzouk, E. R.; Khaksar, M.; Scheckel, K. G.; Stegemeier, J. P.; Lowry, G. V; Donner, E.; Lombi, E. Aging of Dissolved Copper and Copper-based Nanoparticles in Five Different Soils: Short-term Kinetics vs. Long-term Fate. J. Environ. Qual. 2017.
(9) Ivask, A.; Juganson, K.; Bondarenko, O.; Mortimer, M.; Aruoja, V.; Kasemets, K.; Blinova, I.; Heinlaan, M.; Slaveykova, V.; Kahru, A. Mechanisms of toxic action of Ag, ZnO and CuO nanoparticles to selected ecotoxicological test organisms and mammalian cells in vitro: a comparative review. Nanotoxicology 2014, 8 (sup1), 57–71.
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(11) McManus, P.; Hortin, J.; Anderson, A. J.; Jacobson, A. R.; Britt, D. W.; Stewart, J.; McLean, J. E. Rhizosphere interactions between copper oxide nanoparticles and wheat root exudates in a sand matrix: Influences on copper bioavailability and uptake. Environ. Toxicol. Chem. 2018, 37 (10), 2619–2632.
(12) Gao, X.; Lowry, G. V. Progress towards standardized and validated characterizations for measuring physicochemical properties of manufactured nanomaterials relevant to nano health and safety risks. NanoImpact 2017.
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(15) Jiang, C.; Aiken, G. R.; Hsu-Kim, H. Effects of natural organic matter properties on the dissolution kinetics of zinc oxide nanoparticles. Env. Sci Technol 2015, 49 (19), 11476–11484.
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(17) Dimkpa, C. O.; Latta, D. E.; McLean, J. E.; Britt, D. W.; Boyanov, M. I.; Anderson, A. J. Fate of CuO and ZnO nano-and microparticles in the plant environment. Environ. Sci. Technol. 2013, 47 (9), 4734–4742.
(18) Sekine, R.; Brunetti, G.; Donner, E.; Khaksar, M.; Vasilev, K.; Jamting, A. K.; Scheckel, K. G.; Kappen, P.; Zhang, H.; Lombi, E. Speciation and Lability of Ag-, AgCl-, and AgS-Nanoparticles in Soil Determined by X-ray Absorption Spectroscopy and Diffusive Gradients
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CHAPTER 5: Dissolution functional assay improves understanding of metallic nanoparticle toxicity in agricultural soil
Abstract Cu-based ENMs have been used in agriculture. Functional assays are needed to predict their
environmental risks. The objective of this study was to use the dissolution profiles of Cu-based
agrochemicals in soil as a functional assay to predict the toxicity of Cu-based ENMs to Triticum
aestivum. Dissolution kinetics of different Cu-based ENMs, including CuO NP and Cu(OH)2 NP
were measured in standard 2.2 soil with a diethylenetriamine pentaacetic acid (DTPA) extraction
method. DTPA-extractable Cu (i.e. soluble Cu species) was plotted over time to get the dissolution
profile. Biological end points, including the toxicity of Cu-based ENMs to Triticum aestivum (reflected
by reduced maximum root length). The integrated exposure to total soluble Cu (the area under the
dissolution curve corresponding to the exposure interval) was correlated with the biological
endpoint. The integrated exposure to total soluble Cu in different Cu species treatments correlated
well with its toxicity to Triticum aestivum (R2=0.91). In contrast, the standard measure of exposure
(DTPA extractable Cu measured at the end of exposure period) failed to correlate with observed
toxicity (R2=0.09). This study suggested that the integrated exposure of Cu-based ENMs may be a
better measure of exposure than DTPA extractable Cu at the end of an experiment.
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5.1 Introduction Pesticidal, micronutrient, or fungicidal copper-based engineered nanomaterials (ENMs)
could enter agricultural soil though either a direct soil application or indirectly, as overshoot from
foliar application of agrochemicals1–3. It is thus important to understand the fate and bioavailability
of Cu-based ENMs in soil to determine the appropriate application rate of the material, and to
evaluate the risk of the ENMs to non-targeted organisms.
Previous studies have shown that ENMs exhibited a different toxicity than macroparticles or soluble
ions. In aqueous systems, the toxicity of ENMs is usually attributed to the dissolution of ENMs that
releases the soluble ions4–7. Different mechanisms have been proposed to explain the toxicity of
ENMs in soil, either attributing toxicity to a nano-specific effect8 or to ion release9–11. Different
studies have concluded that the same ENMs exerted their toxicity exclusively through one or the
other mechanism. Differences in soil characteristics account for some of the confusion12. This paper
clarifies the seemingly contradictory conclusions of toxicological studies and demonstrates how,
with appropriate functional assays1 to measure the dissolution profile of ENMs in these media,
much of this inconsistency can be explained. Dissolution profile is defined in this paper as the area
under a dissolution curve (dissolved metal concentration over time), as suggested by the shaded
region in Figure 5-1. The objective of this study was to investigate how dissolution profiles
measured directly in soil can explain Cu-based ENMs’ toxicity to wheat plants in soils.
5.2 Methods 5.2.1 Dissolution profile measurement assay
Cu-based ENMs were added to soil and well mixed with wooden sticks. At different aging
periods (T=0h, 1d, 2d, 7d, 14d and 28d, plus T=5h for particles with fast dissolution kinetics), 2.0 g
of air-dried soil were extracted with 4 ml of 0.005M DTPA and 0.1M triethanolamine (TEA)
(pH=7.6) to measure the total soluble Cu released from Cu-based ENMs due to dissolution. Sample
1 Functional assays are semi-empirical methods to measure the processes or functions in a
particular system that can relate to certain endpoints, e.g. toxicity and bioaccumulation19.
Importantly, the outcomes of functional assays are determined by both the properties of the
materials tested and the properties of the environmental systems.
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bottles were horizontally shaken for 2 hours for the extraction. After extraction, all samples were
centrifuged at 3000 rpm for 10 min, and the supernatants were filtered using a 0.2 µm PTFE filter.
All samples were acidified with 20% HNO3 in Milli-Q-water (final HNO3 concentration was 2%)
and analyzed by ICP-MS (Agilent 7700).
DTPA extractable Cu was then plotted over time to get the profile of the dissolution curve,
which was integrated to produce a “integrated exposure to total soluble Cu”.
5.2.2 Plant uptake measurement and toxicity The details of the experimental design is discussed in Chapter 413. Briefly, Triticum aestivum
seedlings were harvested after being exposed to CuO NPs for 14 days in Lufa 2.2 soil. Two
treatments for each copper species were tested, the first when the wheat seedling was transferred
into the copper-amended soil at day 0 and the second when transplantation began on day 28 after
ENM amendment. The latter simulates the aging process in soil. Plant roots and shoots were
photographed with a scale bar for determination of length, as an indication of plant health. In this
study, root length was used as the biological endpoint, as it was the most sensitive toxicity effect
observed for Triticum aestivum.
5.3 Results and discussion:
5.3.1 Differences in dissolution time scale require different assays In soils, different Cu species would have different dissolution profiles (Figure 5-1). For CuO
NP (Figure 5-1 a), the concentration of total soluble Cu in bulk soil increases relatively slowly over
time, before plateauing at this particle’s solubility. In contrast, Cu(OH)2 NPs (Figure 5-1 b)
undergoes an initial fast dissolution followed by sorption to soil organic matter that reduces Cu2+
concentration in soil. For the dissolved species, CuSO4 (Figure 5-1 c), the Cu2+ concentration
remained constant over time after a brief initial equilibration period. These different dissolution
profiles suggest organisms exposed to the same initial doses of different Cu soil amendments will be
exposed to different amounts of Cu2+ initially and over time, and that measuring toxicity without
regard to the dissolution profile will result in noisy at best, and incoherent at worst, results.
Functional assays that measure the actual metal ion exposures therefore should be a part of
nanoparticle toxicity testing protocols in soils.
When the dissolution process is essentially complete before the beginning of the exposure
period, (e.g. CuSO4, Figure 5-1, c ), a single extraction assay measuring the available Cu at the end of
the exposure period will be sufficient.15,16 However, for slowly dissolving NPs (Figure 5-1, a) , or if
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the dissolution profile is complex (Figure 5-1, b,), a single extraction at the end of the experiment is
inadequate because no single time point represents the whole exposure. Instead, integration of the
dissolution curve over the time of exposure to get the integrated exposure is recommended.
0 1 0 2 0 3 0 4 0
0
5 0
1 0 0
1 5 0
2 0 0
T i m e ( d a y s )
Ex
tra
cta
ble
Cu
(mg
/k
g d
rie
d s
oil
)
C u O N P s , L u f a 2 . 2 s o i la )
0 1 0 2 0 3 0 4 0
0
1 0 0
2 0 0
3 0 0
C u ( O H ) 2 , L u f a 2 . 2 s o i l
T i m e ( d a y s )
Ex
tra
cta
ble
Cu
(mg
/k
g d
rie
d s
oil
)
b )
0 1 0 2 0 3 0 4 0
0
1 0 0
2 0 0
3 0 0
T i m e ( d a y s )
Ex
tra
cta
ble
Cu
(mg
/k
g d
rie
d s
oil
)
C u S O 4 , L u f a 2 . 2 s o i lc )
Figure 5-1. The dissolution profile of 250mg/kg of CuO NP, Cu(OH)2 NP and CuSO4 in Lufa 2.2 soil.
5.3.2 Dissolution profile measurement assay predicted toxicity of Cu species to Triticum aestivum Cu-based ENMs in soil usually demonstrated a much slower and more complex Cu release profile,
than soluble Cu species. A solubility measurement at one time point is not sufficient to capture the
different dissolution profiles in soil13,17. To correlate the biological endpoints of Cu-based ENMs
applied to soil (in this study, toxicity), a dissolution profile is needed.
When Triticum aestivum were exposed to fresh and aged CuO NPs and CuSO413, their toxicities
(shortened root length) to Triticum aestivum were correlated with the integrated exposure to total
soluble Cu (arena under the dissolution curve, R2 = 0.91, Figure 5-2 (a)). On the other hand, the
traditional way of measuring bioavailability, which is, using the DTPA extractable Cu at end of the
exposure period, failed to predict the toxicity of Cu species to wheat plants ( R2 = 0.09, Figure 5-2
(b)).
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0 1 0 0 0 2 0 0 0 3 0 0 0 4 0 0 0 5 0 0 0
0
2
4
6
8
1 0
I n t e g r a t e d e x p o s u r e t o t a t a l s o l u b l e C u
( m g * d a y / k g )
Sh
ort
en
ed
ro
ot
ma
xim
al
len
gth
(c
m)
C u O N P ( f r e s h )
C u O N P ( a g e d )
C u S O 4 ( a g e d )
C u S O 4 ( f r e s h )
R2
= 0 . 9 1
( a )
2 0 0 2 5 0 3 0 0 3 5 0
0
2
4
6
8
1 0
C o n c e n t r a t i o n a t t h e e n d o f e x p o s u r e p e r i o d ( m g / K g )Sh
ort
en
ed
ro
ot
ma
xim
al
len
gth
(c
m)
C u O N P ( f r e s h )
C u S O 4 ( a g e d )
C u S O 4 ( f r e s h )
C u O N P ( a g e d )
R2
= 0 . 0 9
( b )
Figure 5-2. Correlations between (a) integrated exposure to total (DTPA-extractable) soluble Cu
and toxicity to Triticum aestivum (shortened root maximal length) and (b) Cu2+ concentration at the
end of exposure period (28 or 42 days of aging) and toxicity to Triticum aestivum (shortened root
maximal length)13. The word “fresh” indicates that plants were exposed to Cu-amended soil right
after the Cu species were applied to soil. “Aged” means the plants were exposed to Cu amended soil
after 28-day aging periods.
The data presented, though suggestive of an important governing principle for soil nanotoxicity,
only cover 2 Cu species (CuO NP and CuSO4) and one plant (Triticum aestivum).In order to be
convincing, this principle would have to be demonstrated for more nanoparticles and more
organisms.
5.3 Environmental Implications
This study suggested that Cu-based ENMs in soil usually require a complete characterization
of the dissolution profile to capture ionic exposure, as a one time measurement of Cu ion
availability is not sufficient to capture the different dissolution profiles 13,17. Previous studies have
developed models that can predict the dissolution profile of CuO NP by knowing the soil pH and
organic carbon content18, which will potentially enable us to predict exposure without measuring
dissolution. In this case, one of the future challenges would be to extend such models and the
functional assays on which they are based, to other types of particles.
This study also raises a concern about the traditional definition of the chronic toxicity test for
nanoparticles in soil. Due to the dissolution behavior of the ENMs, they transformed during a
toxicity test. This suggests that the test result from the toxicity test may not reflect the risk and
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toxicity of the pristine material, but the transformed material (e.g. the effect might be caused by the
released Cu2+ from CuO NP, not the pristine CuO NP). Thus, future studies need to reconsider the
chronic toxicity test for soluble ENMs in soil, with the consideration of their dissolution profile. In
this case, the concept of dissolution profile discussed in this study could be used to better reflect the
exposure of the material instead of concentration. Though not explored here, the variety of shapes
of dissolution profiles may have implications for acute soil nanotoxicity testing in soil as well. Should
the acute exposure coincide with the peak dissolution or be measured always for the same duration
(without consideration of the shape of dissolution profile)?
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(16) Sillanpää, M.; Oikari, A. Assessing the impact of complexation by EDTA and DTPA on heavy metal toxicity using microtox bioassay. Chemosphere 1996, 32 (8), 1485–1497.
(17) Gao, X.; Spielman-Sun, E.; Rodrigues, S. M.; Casman, E. A.; Lowry, G. V. Time and nanoparticle concentration affect the extractability of Cu from CuO NP amended soil. Environ. Sci. Technol. 2017, 51 (4), 2226–2234.
(18) Gao, X.; Rodrigues, S. M.; Spielman-Sun, E.; Lopes, S. P.; Rodrigues, S.; Zhang, Y.; Avellan, A.; Duarte, R. M. B. O.; Duarte, A. C.; Casman, E. A. Effect of soil organic matter, soil pH, and moisture content on solubility and dissolution rate of CuO NPs in soil. Environ. Sci. Technol. 2019.
(19) Hendren, C. O.; Lowry, G. V; Unrine, J. M.; Wiesner, M. R. A functional assay-based strategy for nanomaterial risk forecasting. Sci. Total Environ. 2015, 536, 1029–1037.
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CHAPTER 6: Summary of Major Contributions and Perspective on Future Research
6.1 Summary of Major Contribution This dissertation has made contributions towards the understanding the behavior of engineered
nanomaterials (ENMs) in soil systems, and their impact on soil organisms. The novelty as well as
contributions of this thesis are summarized here.
6.1.1. Major Contribution from Objective 1: A test method to measure dissolution of CuO NP in soil was developed. Two time-based chemical extraction methods were developed to measure the dissolution kinetics of
CuO NPs directly in a soil matrix. The first method is based on a pore-water extraction method1,2
(CaCl2 extraction) that measures the dissolution by monitoring the concentration of released Cu2+
from CuO NP in soil pore water. The second method is based on a labile metal extraction method3
(DTPA extraction) that measures the dissolution by monitoring the reversibly-sorbed fraction
(~80%) of the Cu released from CuO NP4. Both extractions were performed at different aging times
after CuO NPs were dosed to soil. The increase in extractable Cu was fit with a kinetic model to
quantify the dissolution rate constant. This study provided researchers in this field a tool to evaluate
the dissolution of Cu-based ENMs in soil systems.
6.1.2. Major Contribution from Objective 2: A model was developed to evaluate the effect of soil pH and organic carbon content on dissolution kinetics of CuO NP in soil. It is unclear from previous studies how soil properties affect the dissolution kinetics of CuO NP in
soil. In this project, we used the tool developed from Objective 1 to measure the dissolution kinetics
of CuO NP in various soils to answer the fundamental question: How do soil properties affect the
dissolution kinetics of CuO NP in soil? Soil organic matter (SOM) content showed a positive
correlation with the solubility of CuO NP in soil, but did not affect the overall dissolution rate
constant (half-life) of the CuO NPs in soil, whereas soil pH affected the dissolution rate constant of
the dissolution process but showed no effect on the ultimate solubility of CuO NP in soil. Soil
moisture content had minimal effect on dissolution kinetics of CuO NP in soil. Based on these
observations, an empirical model to correlate the CuO NP dissolution kinetics with both SOM
content and soil pH was developed. This dissolution model successfully predicted the dissolution
kinetics of CuO NPs in two unknown soil. This work provides fundamental insights on why SOM
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and soil pH play such an important role on the dissolution kinetics of CuO NP in soil. The
dissolution model could also help to explain the differences in behavior (e.g. toxicity) that has been
observed in different soils. The method that we used to develop the dissolution model for CuO NP
could potentially be extended to other metal/metal oxide particles with low to moderate dissolution
rates (half-lives of days to years).
6.1.3. Major Contributions from Objective 3:Dissolution of CuO NPs in soil was correlated with its toxicity to wheat (Triticum aestivum). Dissolution of CuO NPs under the influence of root activity in rhizosphere soil was quantified. Although previous studies have postulated that dissolution of ENMs would affect its toxicity in
soil5–7, the correlation between ENMs dissolution and their toxicity to soil organisms could not be
made without measurement of their dissolution in soil. Using the dissolution measurement method
developed from Objective 1, this work directly correlated the dissolution of CuO NP with its
toxicity to wheat plants for the first time. The increase in toxicity of CuO NP after aging in soil
relative to unaged CuO NPs is explained by CuO NP’s slow dissolution behavior. Here, the impact
of plant roots on dissolution and lability of CuO NP in rhizosphere soil were also quantified for the
first time. This study emphasized the role of ENMs dissolution in understanding their toxicity,
including properly identifying any “nanoparticle specific” toxicity effects.
6.1.4. Major Contribution from Objective 4: Dissolution kinetics functional assays were used to estimate exposure to ionic Cu from Cu-based ENMs in soil. This exposure correlated to observed toxicity in wheat. With the method developed in Objective 1 to measure the dissolution kinetics of CuO NP in soil,
and the observed correlations between dissolution of CuO NPs and its toxicity to wheat in
Objective 3, this work further extended the understanding of the relationship between ENM
dissolution in soil and their toxicity. Here, a functional assay to measure the dissolution profile, i.e.
available Cu concentration over time, of Cu based ENMs (including CuO NP and Cu(OH)2 NP)
was used to predict exposure to Cu ion over time. This exposure correlated to toxicity to wheat
roots demonstrating the importance of properly quantifying exposures over time when dealing with
ENMs in soil.
Overall, this thesis illustrated a way to study dissolution behavior of ENMs in soil. Aging time was
shown to be one important factor to be considered when investigating dissolution. The dissolution
model for CuO NP in soil provided mechanistic understanding on how soil properties affect the
dissolution of ENMs in aerobic soil. It should be noted that this model is based on lab conditions in
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natural but standard soils, not field conditions, which are more complex. For example, soils may
become anerobic for a period of time which could affect the speciation of Cu and the resulting CuO
NP behavior. However, the behavior of CuO NPs in aerobic field soils can be predicted with the
dissolution model developed in this thesis. Leaching is an important process to consider under field
conditions, it essentially removes mobile Cu2+ from the soil. According to the dissolution model
developed in this work, leaching would likely only remove a minor fraction of the Cu in system since
most Cu is sorbed to organic matter and soil solid surfaces rather than mobile in the porewater.
Thus, it may not significantly change the overall dissolution kinetics of CuO NP after all, depending
on the duration of the leaching.
This thesis will be of interest to nanotoxicologists because the measurement of dissolution of ENMs
in soil could help them understand the mechanism of nanotoxicity and precisely define NP exposure
in standard testing. By knowing the dissolution kinetics in soil, the works in this thesis will also guide
the design of nano-enabled agrochemicals. Environmental regulators will also learn from this thesis
that regulations on ENMs in soil might be soil-property-dependent.
6.2 Perspectives for future research Although already being used in agriculture, the risk and environmental impacts of nano-
enabled agrochemicals still need to be evaluated. While it is yet to be known if nano-enabled
agrochemicals will significantly replace the metallic salts traditionally found in agrochemicals, nano-
enabled agrochemicals could be designed to have high efficiency with low environmental risks.
Dissolution is one of the most important processes for metal and metal oxide ENMs in soil, as it
releases the active ingredients, releasing metal ions into the soil. This thesis provides a tool to
measure the dissolution of metal and metal oxide ENMs in soil, as well as models to predict the
dissolution behavior of CuO NP in soils to better quantify soil organisms’ exposure to ions released
from the ENMs. However, there is still more work needed to ensure the efficiency and safety of
nano-enabled agrochemicals. This section will address relevant key challenges that are needed to
guarantee an efficient and safe application of nano-enabled agrochemicals.
6.2.1. Extension of the model to predict the behavior of other metal/metal oxide ENMs in soil CuO NPs are only one type of nano-enabled agrochemical being used in agriculture. Other
metal/metal oxide ENMs, such as ZnO NPs, Ag NPs, Fe2O3 NPs, Fe (OH)3 NPs, Si NPs all have
the potential to be applied in agriculture, either as micronutrient suppliers, or as
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fungicides/pesicides8,9. Also they can enter agricultural soils via accidental release10. While soil pH
and organic matter were considered as the key properties influencing the dissolution kinetics of CuO
NPs, other processes (e.g. redox reactions) will likely be important, as well, for some other NPs and
these need to be addressed in relevant cases. Thus, it could be important to extend the models for
predicting the dissolution kinetics under different redox scenarios. There are a few major challenges
that need to be accomplished to meet such a goal.
First, the Cu speciation change in soil over time is needed. In Chapter 3, the partitioning
between free Cu2+ and Cu2+ associated with different soil surfaces (SOM, clay, etc.) was simplified
using one partitioning constant, Kd. This method worked for this particular model, as SOM is the
predominant soluble Cu sink in soil11,12. However, this may not be the case for other metals.
Coupling the kinetic dissolution model of CuO NP with the equilibrium Cu speciation model (multi-
surface model) would enable us to get a more precise speciation of Cu in soil over time. Such an
exercise would be important to understand the detailed metal speciation in soil dosed with ENMs as
a function of time.
Second, in this thesis, the effect of particle properties on dissolution of CuO NPs in soil was
not investigated. Particle size and shape have been suggested to be important factors affecting the
dissolution kinetics of ENMs in aqueous systems13. The size and shape of ENMs could also
potentially affect their dissolution kinetics in soil, by affecting the ion-release step of the dissolution,
which was suggested in Chapter 3 to be the rate-limiting step. Future work is needed to incorporate
the size and shape properties of CuO NP into the dissolution model.
Third, the extraction methods may need to be modified to measure the dissolution rate of
other metal/metal oxide ENMs in soil. This is required to collect the data needed to build the
model. As summarized in Section 6.1.1, the two extraction methods worked well on measuring the
dissolution of CuO NPs in soils. The success of applying the two chemical extraction methods is
because 1) For CaCl2 extraction, previous studies have shown that the partitioning of Cu2+ between
the soil solid phases and soil pore water is relatively stable in a given soil media11,14. 2) For DTPA
extraction, Cu2+ has a very high stability constant with DTPA15,16, making DTPA a strong chelating
agent capable of extracting most of Cu2+ released from CuO NP in soil. These two necessary
conditions may not be met for other metal/metal oxide ENMs in soil. For example, DTPA does not
chelate strongly with Silver(Ag)17, making it a low efficiency extractant for Ag+ ions from soil. The
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extraction conditions (e.g. extractant used, extraction time, etc.) should be adjusted for different
species of metal/metal oxide ENMs to achieve the best efficiency.
Fourth, the dissolution model needs to be extended to include other important
transformation processes. For example, redox reactions need to be considered for many metals that
are redox active, including Cu2+. The experimental condition applied in Objective 2 was strongly
oxidizing. In agriculture, agrochemicals are usually applied to, and remain in the top soil18, where soil
conditions are generally aerobic. However, there are cases in agriculture that would create anaerobic
conditions, such as paddies of lowland rice, where anaerobic conditions would result from a flood-
irrigation system19, or if the soils flood temporarily. The chemical transformations of many metals
are sensitive to changes in redox potential of the environment20. In anaerobic conditions, metals
such as Cu2+, could be reduced to Cu1+ species or to Cu0, affecting their dissolution behavior. In
fact, a previous study has already observed that CuO NPs were reduced to Cu(I) species in a rice
growing environment21. In this case, redox reactions must be considered when understanding the
dissolution and toxicity of the CuO NPs to rice. To incorporate redox processes into the dissolution
model is necessary to understand the full picture of transformation and dissolution of metal and
metal oxide ENMs in soil/sediment with anaerobic conditions. One potential method for doing this
could be to measure the dissolution rate constant under different redox conditions, and then
correlate dissolution rate constant with redox potential.
Greater understanding of how the composition of SOM and DOM affect the dissolution of
kinetics of CuO NP in soil is also needed. As discussed in Chapter 3, one of the reasons for the
uncertainty in predicting the dissolution profile is the assumption of similar SOM quality (thus same
complexing capacity with Cu) among soils. With more specific characterization of SOM and DOM
composition, it may indeed be possible to improve the predictive capability of the model.
Finally, root exudates need to be better characterized to refine modeling of dissolution of
metal/metal oxide ENMs in the soil rhizosphere. While dissolution of CuO NP in the rhizosphere
was found to be influenced by root activities (Objective 3), the mechanism of this effect is unclear.
In Objective 3, changes in dissolution and bioavailability in rhizosphere soil was hypothesized to be
a result of plant root exudates, and corresponding changes in rhizosphere soil pH. However, it could
also be a result of bacterial activity in the rhizosphere soil. Previous studies have shown that plant
root exudates affect bacterial activity and relevant enzyme activity22,23. Thus, a mechanistic
understanding of the role of root exudates or change in rhizosphere bacterial activity upon
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dissolution of CuO NP in rhizosphere soil would require evaluating those influences separately.
Characterizing root exudation and bacterial community analysis (or enzyme activity) would be
required to answer this question. Such understanding would be helpful to more accurately model
and predict the dissolution behavior of metal or metal based ENMs in rhizosphere soil.
6.2.2. Optimize the way to measure toxicity of metal/metal oxide ENMs in soil. In Chapter 5, the dissolution profile of various Cu species correlated with their toxicity to
wheat plants. However, such correlation was based only upon four data points. To strengthen such
correlations, more experimental data are needed. Dissolution profiles for other Cu-based ENMs
(e.g. Cu(OH)2) and Cu ENMs containing product (e.g. Kocide® as Cu(OH)2, Nordox ® (Cu2O))
should be measured. Other types of biological endpoints (e.g. toxicity to other soil organisms, e.g.
soil isopods) should be measured.
Chapter 5 also suggested that the dissolution profile of different Cu-based ENMs should be
considered in the design of chronic toxicity tests for soil organisms. Usually, the length of a chronic
toxicity test is decided only by the tested organism. For example, the chronic toxicity test for Eisenia
andrei usually has 28 day period, while the chronic toxicity test for Hyalella Azteca usually has a 42 day
period24–26. There is currently no consideration of the NP properties on test design or duration. This
is a problem for those toxicity tests because the tested nanomaterials are transforming during the
test period (e.g. CuO and Cu(OH)2 NPs dissolve and release Cu2+) and the test could miss the salient
part of the exposure. Thus, future studies need to redefine the chronic toxicity testing protocols for
metal/metal oxide ENMs in soil, with the consideration of their dissolution profile. In this case, the
concept of ‘exposure profile’ explained in Chapter 5 could be used to better reflect the exposure of
the material instead of concentration.
6.2.3. Design ENMs that can solve the micronutrient deficiency problem in calcareous soil. One of the challenges for agriculture is micronutrient deficiency 27,28, especially in calcareous
(high pH) soils. One would predict that applying soluble ENMs to calcareous soil usually would not
solve the problem, due the low availability of metal ion at high pH27. In such cases, supplying
metal/metal oxide ENMs as foliar application and seed coatings might be better approaches to
preventing micronutrient deficiency. Foliar application with ENMs could be a feasible way of
efficiently delivering nutrients. Foliar applied ENMs have been shown to be able to enter and
translocate in plants29,30. Seed coatings could be devised to stick ENMs to the plant roots and slowly
release micronutrients to the rhizosphere during the plant growth period. However, there are still
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challenges remaining: 1) To design ENMs that have the appropriate micronutrient release rate, in
which particle coatings, chemical composition, and size could be important tunable parameters. 2)
To monitor the bioavailability of relevant species in plant rhizosphere and the metal speciation
inside of plants. 3) To measure the availability of other metal/metal ENMs in the rhizosphere soil,
and to track the speciation of metal/metal ENMs inside plants.
6.3 References for Chapter 6 (1) Houba, V. J. G.; Novozamsky, I.; Lexmond, T. M.; Van der Lee, J. J. Applicability of 0.01 M
CaCl2 as a single extraction solution for the assessment of the nutrient status of soils and other diagnostic purposes. Commun. Soil Sci. Plant Anal. 1990, 21 (19–20), 2281–2290.
(2) Houba, V. J. G.; Temminghoff, E. J. M.; Gaikhorst, G. A.; Van Vark, W. Soil analysis procedures using 0.01 M calcium chloride as extraction reagent. Commun. Soil Sci. Plant Anal. 2000, 31 (9–10), 1299–1396.
(3) Feng, M. H.; Shan, X. Q.; Zhang, S.; Wen, B. A comparison of the rhizosphere-based method with DTPA, EDTA, CaCl2, and NaNO3 extraction methods for prediction of bioavailability of metals in soil to barley. Env. Pollut 2005, 137 (2), 231–240.
(4) Gao, X.; Spielman-Sun, E.; Rodrigues, S. M.; Casman, E. A.; Lowry, G. V. Time and nanoparticle concentration affect the extractability of Cu from CuO NP amended soil. Environ. Sci. Technol. 2017, 51 (4), 2226–2234.
(5) Tourinho, P. S.; van Gestel, C. A.; Lofts, S.; Svendsen, C.; Soares, A. M.; Loureiro, S. Metal-based nanoparticles in soil: fate, behavior, and effects on soil invertebrates. Env. Toxicol Chem 2012, 31 (8), 1679–1692.
(6) Dimkpa, C. O.; Latta, D. E.; McLean, J. E.; Britt, D. W.; Boyanov, M. I.; Anderson, A. J. Fate of CuO and ZnO nano- and microparticles in the plant environment. Env. Sci Technol 2013, 47 (9), 4734–4742.
(7) Watson, J.-L.; Fang, T.; Dimkpa, C. O.; Britt, D. W.; McLean, J. E.; Jacobson, A.; Anderson, A. J. The phytotoxicity of ZnO nanoparticles on wheat varies with soil properties. Biometals 2015, 28 (1), 101–112.
(8) Kim, S. W.; Jung, J. H.; Lamsal, K.; Kim, Y. S.; Min, J. S.; Lee, Y. S. Antifungal effects of silver nanoparticles (AgNPs) against various plant pathogenic fungi. Mycobiology 2012, 40 (1), 53–58.
(9) Liu, R.; Lal, R. Potentials of engineered nanoparticles as fertilizers for increasing agronomic productions. Sci Total Env. 2015, 514, 131–139.
(10) Pradas del Real, A. E.; Castillo-Michel, H. A.; Kaegi, R.; Sinnet, B.; Magnin, V.; Findling, N.; Villanova, J.; Carriere, M.; Santaella, C.; Fernandez-Martinez, A. Fate of Ag-NPs in sewage sludge after application on agricultural soils. Env. Sci Technol 2016.
(11) Bonten, L. T. C.; Groenenberg, J. E.; Weng, L.; van Riemsdijk, W. H. Use of speciation and complexation models to estimate heavy metal sorption in soils. Geoderma 2008, 146 (1), 303–310.
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(12) Weng, L.; Temminghoff, E. J. M.; Van Riemsdijk, W. H. Contribution of individual sorbents to the control of heavy metal activity in sandy soil. Env. Sci Technol 2001, 35 (22), 4436–4443.
(13) Peretyazhko, T. S.; Zhang, Q.; Colvin, V. L. Size-controlled dissolution of silver nanoparticles at neutral and acidic pH conditions: kinetics and size changes. Environ. Sci. Technol. 2014, 48 (20), 11954–11961.
(14) Weng, L.; Temminghoff, E. J. M.; Lofts, S.; Tipping, E.; Van Riemsdijk, W. H. Complexation with dissolved organic matter and solubility control of heavy metals in a sandy soil. Env. Sci Technol 2002, 36 (22), 4804–4810.
(15) Baumann, E. W. Investigation of copper (II) chelates of EDTA and DTPA with cupric-selective electrodes. J. Inorg. Nucl. Chem. 1974, 36 (8), 1827–1832.
(16) Sillanpää, M.; Oikari, A. Assessing the impact of complexation by EDTA and DTPA on heavy metal toxicity using microtox bioassay. Chemosphere 1996, 32 (8), 1485–1497.
(17) Kapoor, S.; Mukherjee, T. Growth and reactivity of silver clusters in amino polycarboxylic acid solutions. J. Colloid Interface Sci. 2003, 264 (1), 301–306.
(18) Huang, S. S.; Liao, Q. L.; Hua, M.; Wu, X. M.; Bi, K. S.; Yan, C. Y.; Chen, B.; Zhang, X. Y. Survey of heavy metal pollution and assessment of agricultural soil in Yangzhong district, Jiangsu Province, China. Chemosphere 2007, 67 (11), 2148–2155.
(19) Fageria, N. K.; Slaton, N. A.; Baligar, V. C. Nutrient management for improving lowland rice productivity and sustainability. Adv. Agron. 2003, 80 (1), 63–152.
(20) Gambrell, R. P.; Wiesepape, J. B.; Patrick, W. H.; Duff, M. C. The effects of pH, redox, and salinity on metal release from a contaminated sediment. Water. Air. Soil Pollut. 1991, 57 (1), 359–367.
(21) Peng, C.; Xu, C.; Liu, Q.; Sun, L.; Luo, Y.; Shi, J. Fate and Transformation of CuO Nanoparticles in the Soil–Rice System during the Life Cycle of Rice Plants. Environ. Sci. Technol. 2017, 51 (9), 4907–4917.
(22) Baudoin, E.; Benizri, E.; Guckert, A. Impact of artificial root exudates on the bacterial community structure in bulk soil and maize rhizosphere. Soil Biol. Biochem. 2003, 35 (9), 1183–1192.
(23) Ai, C.; Liang, G.; Sun, J.; Wang, X.; Zhou, W. Responses of extracellular enzyme activities and microbial community in both the rhizosphere and bulk soil to long-term fertilization practices in a fluvo-aquic soil. Geoderma 2012, 173, 330–338.
(24) Ingersoll, C. G.; Brunson, E. L.; Dwyer, F. J.; Hardesty, D. K.; Kemble, N. E. Use of sublethal endpoints in sediment toxicity tests with the amphipod Hyalella azteca. Environ. Toxicol. Chem. An Int. J. 1998, 17 (8), 1508–1523.
(25) Ivey, C. D.; Ingersoll, C. G.; Brumbaugh, W. G.; Hammer, E. J.; Mount, D. R.; Hockett, J. R.; Norberg‐King, T. J.; Soucek, D.; Taylor, L. Using an interlaboratory study to revise methods for conducting 10‐d to 42‐d water or sediment toxicity tests with Hyalella azteca. Environ. Toxicol. Chem. 2016, 35 (10), 2439–2447.
(26) Robidoux, P. Y.; Svendsen, C.; Caumartin, J.; Hawari, J.; Ampleman, G.; Thiboutot, S.;
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Weeks, J. M.; Sunahara, G. I. Chronic toxicity of energetic compounds in soil determined using the earthworm (Eisenia andrei) reproduction test. Environ. Toxicol. Chem. An Int. J. 2000, 19 (7), 1764–1773.
(27) Chen, Y.; Barak, P. Iron nutrition of plants in calcareous soils. In Advances in agronomy; Elsevier, 1982; Vol. 35, pp 217–240.
(28) Rashid, A.; Ryan, J. Micronutrient constraints to crop production in soils with Mediterranean-type characteristics: a review. J. Plant Nutr. 2004, 27 (6), 959–975.
(29) Lowry, G. V; Avellan, A.; Gilbertson, L. M. Opportunities and challenges for nanotechnology in the agri-tech revolution. Nat. Nanotechnol. 2019, 14 (6), 517.
(30) Avellan, A.; Yun, J.; Zhang, Y.; Spielman-Sun, E.; Unrine, J. M.; Thieme, J.; Li, J.; Lombi, E.; Bland, G.; Lowry, G. V. Nanoparticle Size and Coating Chemistry Control Foliar Uptake Pathways, Translocation and Leaf-to-Rhizosphere Transport in Wheat. ACS Nano 2019.
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Appendices Appendix 1- Supporting information for Chapter 2: Develop a functional assay to measure the dissolution kinetics of metal-based nanoparticles in soil. Detailed nanoparticle amendment procedure
Nanoparticle powders were dispersed in 5mM NaHCO3 buffer in Milli-Q water (pH=7) (the final
particle concentration was 62 mg Cu/L for the low dose amendment and 620 mg Cu/L for the high
dose amendment). The suspension was then sonicated using an ultrasonic probe (Sonic
Dismembrators model 550, Fisher Scientific) for 30 s (We used a pulse mode with 5s pulse per 10 s
sonication, energy level 3, ~160W). The dispersed nanoparticles were then immediately added to
soil. Cu(NO3)2•2.5H2O powder was dissolved in Milli-Q water before being amended to the soil
(final concentration was 620 mg Cu/L for the high dose amendment, and 62 mg Cu/L for the low
dose amendment). For extraction experiments, for both Cu(NO3)2 and CuO NP amendments, 40ml
of the NP suspension or solution was added to 250g air-dried soil. For the control soil, we added 40
ml of bicarbonate buffer to 250 g of soil.
The amount of Milli-Q water added to each soil was controlled to ensure that each was at its field
capacity after amendment (16 wt%). For the high dose CuO NP amendment, Cu(NO3)2 amendment
and control amendment, after taking 10g of soil for digestion, the soil was further divided into 21
tubes (3 replicates * 7 different incubation periods). For the low dose CuO NP amendment and
Cu(NO3)2 amendment, after taking 10g of soil for digestion experiment, the soil was further divided
into 24 tubes (3 replicates * 8 different incubation periods).
For XAS study, for both Cu(NO3)2 and CuO NP amendments, 8ml of the NP suspension or
solution was added to 50 g air-dried soil (final moisture content was also 16%). For both high dose
and low dose amendment, after taking 10g of soil for the digestion experiment, soil was further
divided into 7 tubes (7 different incubation periods, no replicates).
All soil samples were incubated in 50ml centrifuge tubes under aerobic conditions for between 0 and
31 days before being extracted and digested. The moisture content of the soil samples was kept at
the soil field capacity by daily additions of Milli Q water.
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Table A1-1. Calibration ranges used for ICP-MS measurement
Sample Calibration Range
DTPA extractable Cu in 10mg/kg
CuO NP amendment 0.01mg/kg-0.1mg/kg
CaCl2 extractable Cu in 10mg/kg
CuO NP amendment 0.002mg/kg – 0.01mg/kg
DTPA extractable Cu in 100mg/kg
CuO NP amendment 0.05mg/kg-1mg/kg
CaCl2 extractable Cu in 100mg/kg
CuO NP amendment 0.01mg/kg-0.1mg/kg
DTPA extractable Cu in 10mg/kg
Cu(NO3)2 amendment 0.01mg/kg-0.1mg/kg
CaCl2 extractable Cu in 10mg/kg
Cu(NO3)2 amendment 0.01mg/kg-0.1mg/kg
DTPA extractable Cu in 100mg/kg
Cu(NO3)2 amendment 0.05mg/kg-1mg/kg
CaCl2 extractable Cu in 100mg/kg
Cu(NO3)2 amendment 0.05mg/kg-1mg/kg
Total Cu in 10mg/kg
CuO NP amendment 0.01mg/kg-0.1mg/kg
Total Cu in 10mg/kg
CuO NP amendment 0.01mg/kg-0.1mg/kg
Total Cu in 100mg/kg
Cu(NO3)2 amendment 0.05mg/kg-1mg/kg
Total Cu in 100mg/kg
Cu(NO3)2 amendment 0.05mg/kg-1mg/kg
All samples collected from
umamended soil 0.001mg/kg – 0.01mg/kg
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Table A1-2. Total Cu measured in amended soils (4 replicates)
High dose CuO NP
amended soil (mg/kg)
(s.d.)
Low dose CuO NP
amended soil (mg/kg)
(s.d.)
High dose Cu(NO3)2
amended soil (mg/kg)
(s.d.)
Low dose Cu(NO3)2
amended soil (mg/kg)
(s.d.)
For extractions 99.6 (6.5) 13.1(0.7) 100.6 (9.4) 13.0 (0.6)
For XANES experiments
108.7 (1.3) 12.9 (0.1) 104.4 (1.7) 13.3 (0.2)
Table A1-3. Extractable Cu in unamended soil (3 replicates)
DTPA extraction CaCl2 extraction
Incubation time Wet soil extraction
(s.d.)
Dry soil extraction
(s.d.)
Wet soil extraction
(s.d.)
Dry soil extraction
(s.d.)
0 0.49 (0.04) 0.49 (0.04) 0.018 (0.010) 0.011 (0.002)
1 0.38 (0.17) 0.39 (0.08) 0.012 (0.001) 0.013 (0.002)
2 0.53 (0.12) 0.49 (0.03) 0.015 (0.006) 0.024 (0.012)
4 0.44 (0.03) 0.50 (0.00) 0.013 (0.009) 0.015 (0.007)
7 0.37 (0.05) 0.45 (0.01) 0.005 (0.002) 0.011 (0.003)
19 0.37 (0.01) 0.44 (0.01) 0.011 (0.002) 0.019 (0.006)
31 0.41 (0.01) 0.60 (0.14) 0.006 (0.001) 0.014 (0.004)
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Figure A1-1. A) Primary particle size distribution determined from counting primary particles from 10 TEM imagines. B-K) Ten TEM images of CuO NP. Red bars indicate the counted particles. Heavily aggregated nanoparticles were not counted.
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Figure A1-2. Size distribution of 80mg/kg CuO NP in pH=7, 5mM NaHCO3 buffer determined by
dynamic light scattering: (a) Number averaged size distribution, (b) intensity averaged size
distribution and (c) and autocorrelation function.
c
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)
b )
p H
Figure A1-3. Zeta potential of 80mg/kg CuO NP as a function of pH measured in (a) 5mM
NaHCO3 buffer and (b) 5mM NaNO3. Error bars indicate ± 1standard error. The shift in the pH
of the isoelectric point from pH=8.8 in NaNO3 to pH=5.5 in bicarbonate indicates a specific
interaction between the CuO NPs and the carbonate species.
- 98 -
Figure A1-4. X-ray diffraction spectrum of CuO NP. The CuO NPs used here are identified as
tenorite.
12 22 32 42 52 62 72 82
Inte
ntis
ity
2 θ(degrees)
CuO NP Tenorite
- 99 -
Figure A1-5. pH of CaCl2 extracts in different amended and blank soils (pH values in CaCl2 extracts
of the high dose CuO NP amended soils at t=7 days after amendment were not measured.) Error
bars indicate ± 1standard error.
- 100 -
- 101 -
Figure A1-6. Extractable Cu and in wet and air dried amended soils as a function of time: (a)
DTPA extraction for 10 mg/kg amendment, (b) CaCl2 extraction for 10 mg/kg amendment, (c)
DTPA extraction for 100 mg/kg amendment and (d) CaCl2 extraction for 100 mg/kg amendment.
Error bars indicate ± 1 standard error. represents extractable Cu in CuO NP amended wet soils;
represents extractable Cu in CuO NP amended soils air dried after incubation; represents
extractable Cu in Cu(NO3)2 amended wet soils, and represents extractable Cu in Cu(NO3)2
amended soils air dried after incubation.
Extractions performed on wet and air-dried soils indicated that the influence of air-drying on DTPA
extractable Cu and CaCl2 extractable Cu in soils is insigficant for all amendments (P>0.05,
- 102 -
Kolmogorov–Smirnov test) (Figure S6). The insensitivity of CuO NP to air drying may result from
because the Cu(II) in the particles being oxygen-insensitive. The extractability of metal from oxygen-
sensitive (redox active) nanoparticles like metallic Cu NP or metallic Ag NP in soil may be more
affected by air drying than the CuO NP used here. This requires further study.
8 9 8 0 9 0 0 0 9 0 2 0 9 0 4 0 9 0 6 0
0
E n e r g y ( e v )
No
rma
liz
ed
ad
so
rpti
on
C u ( 0 )
C u S ( C o v e l l i t e )
C u ( I I ) N i t r a t e
C u ( I I ) P h o s p h a t e
C u ( I I ) S u l f a t e
C u O N P
C u ( I I ) - H A
C u ( I I ) F e r r i h y d r i t e
C u ( I ) - H A
C u 2 S ( C h a l c o c i t e )
Figure A1-7. XANES spectra for model compounds
Cu(II) sorbed to humic acid (Cu(II)-HA) was synthesized using established methods (Fulda et al.
2013)1. 118 mg of a humic acid standard (H1, 675-2, Sigma-Aldrich) was dissolved in 25 mL of DI
water, then 200 mg of CuCl2 was slowly added with stirring. The solution was filtered and dried.
- 103 -
The spectrum for Cu sorbed to ferrihydrite (Cu(II)- ferrihydrite) was obtained from Donner et al.
(2011)2. The Spectrum for Cu(I) sorbed to humic acid (Cu(I)-HA) was obtained from (Fulda et al.
2013)1.
Figure A1-7 indicated that reduction of Cu(II) to Cu(I) and sulfidation of Cu will result in a
significant white line shift, which is not observed in our experimental data. The LCF analysis
indicated CuO NP in soil transformed to Cu(II)-HA slowly over time.
Table A1-4. The results of the LCF analysis of the X-Ray Absorption Near Edge Structure
(XANES) region of the samples
Copper Time Goodness of fit
Percentages of Components
Sum R value Cu(II)- HA (%) CuO NP (%)
CuO NP Day 1 0.00022 14 86 100
CuO NP Day 4 0.00019 35 66 101
CuO NP Day 7 0.00014 41 59 100
CuO NP Day 19 0.00017 44 56 100
Cu(NO3)2 Day 0 0.00026 100 0 100
Cu(NO3)2 Day 1 0.00041 101 0 101
Cu(NO3)2 Day 4 0.00046 102 0 102
Cu(NO3)2 Day 7 0.00020 100 0 100
Cu(NO3)2 Day 19 0.00023 101 0 101
Appendix 2- Supporting information for chapter 3:Quantify the effect of soil properties, including soil moisture content, organic carbon content and pH, on the dissolution kinetics of CuO NP in soil.
Procedure for the extraction of SOM (FA and HA) from Lufa 2.1 soil
The SOM (here considered the fulvic acids (FA) and humic acids (HA) fractions) was extracted from
Lufa 2.1 soil following a procedure described by van Zomeren et al.3 A set of sixteen centrifuge tubes
containing the same amount of Lufa 2.1 soil were subjected to the same extraction procedure. Briefly,
40 mL of 0.1M HCl (pH 1.2) was added to 4 g of Lufa 2.1 soil (L/S=10) in a centrifuge tube, and the
suspension was mixed by continuous tumbling for 1h. Afterwards, the suspension was centrifuged
- 104 -
(3000g, 10 min). The supernatant was decanted and filtered using a 0.45 µm nitrocellulose membrane,
the pH adjusted to pH 2.0 with 1M NaOH and subsampled for dissolved organic carbon (DOC)
analysis. The remaining supernatant was stored frozen for further isolation from the inorganic matrix
(data not shown).
The soil residue was then neutralized with 0.1M NaOH, and thereafter 0.1M NaOH (pH 12) was
added under a N2 atmosphere to a final volume of 40 mL (L/S =10). The suspension was equilibrated
during 20h by intermittent shaking, and then centrifuged (3000g, 10 min), decanted and filtered, using
a 0.45 µm nitrocellulose membrane. The supernatant was acidified to pH 1.0 with 6M HCl and allowed
to stand overnight (≈20h) in a refrigerator (4ºC) to precipitate the HA fraction. The resultant
suspension (containing FA and inorganics) was centrifuged (3000g, 10 min), and the obtained
supernatant was decanted and filtered, the pH adjusted to 2.0 with 1M NaOH and stored frozen for
further extraction/isolation from the inorganic matrix. The precipitate with the HA fraction was then
washed with 10 mL of ultrapure water, the suspension was centrifuged (3000g, 10 min), and the
supernatant was discarded. The HA was then re-dissolved in a 0.1M NaOH solution, containing 0.2M
KCl, under a N2 atmosphere. The KCl was added to NaOH to increase the ionic strength, provoking
the flocculation of the colloidal inorganic material (non-humic material). The suspended solids were
removed by centrifugation (3000g, 10 min) and the HA solution was decanted under N2 atmosphere.
The process was repeated until practically all the HA was separated from the inorganic matrix. The
final solution of HA was acidified with 6M HCl to pH 1.0. The acidified solution was allowed to stand
overnight, and the supernatant was finally separated from the HA precipitate by centrifugation (3000g,
10 min). The HA was then re-dissolved in a minimum volume of 0.1M NaOH under a N2 atmosphere,
and then acidified to pH 2.0 and further desalted onto DAX-8 and cationic exchange resins.
The FA and HA fractions were isolated from the inorganic matrix by adsorption onto a Supelite™
DAX-8, followed by a cationic exchange resin in H+ mode, connected in series. The FA and HA
fractions, previously acidified to pH 2, were pumped through a DAX-8 column, and then the
inorganics were removed from the void volume with one column volume of ultra-pure water. The FA
and HA were then back eluted with 0.1M NaOH (3.5 column volumes) and directly transferred into
the cationic exchange column. Finally, the FA and HA solutions were freeze-dried and kept in a
dessicator over silica gel until further characterization by thermogravimetry and elemental analysis.
The DOC content of the influents and effluents from both resins were also measured for assessing
losses of organic carbon during the isolation/purification procedure.
- 105 -
Recovery of FA and HA from Lufa 2.1 soil and preliminary characterization
For a total LUFA 2.1 soil mass of 256 g, the total organic carbon extracted from the soil in each
fraction was: FA-SOM - 740.5 mg OC kg-1 soil d.w. and HA-SOM - 872.7 mg OC kg-1 soil d.w. The
total OC extracted from LUFA 2.1 soil was 1613.2 mg C kg-1, corresponding to an extraction efficiency
of 23.1 % of the total OC in the original soil sample. The FA-SOM and HA-SOM fractions account
for 46% and 54% of the total SOM extracted from LUFA 2.1 soil, respectively. The elemental analysis
data showed that the FA-SOM and HA-SOM fractions contain c.a. 39% and 32% carbon, respectively,
and that ash and impurities account for 24% and 38% of the sample mass, respectively.
Details on Cu speciation measurements
Cu speciation in Lufa soil and Arizona soil were analyzed by Cu K-edge x-ray absorption spectroscopy
at the Stanford synchrotron Radiation Lightsource (SSRL) on Beamlines 11-2 and 4-1, respectively.
Samples were ground, pressed into pellets, and placed between Kapton tape. Double crystal Si(220)
monochromator was calibrated by setting the first inflection of the K-edge of a metallic Cu foil to
8979 eV. On Beamline 11-2, harmonic rejection was achieved by use of a Rh-coated mirror, and
fluorescence data were recorded at 77 K using a 100-element germanium detector. On Beamline 4-1,
harmonic rejection was achieved by detuning the monochromator crystal by 20%, and fluorescence
data were recorded at 77 K using a 32-element germanium detector. The scans were averaged, energy
corrected using a metallic Cu foil standard and deadtime-corrected using SIXPack data analysis
software (v 1.4).4 Spectra background was subtracted and normalized before linear combination fitting
(LCF) analysis using Athena XAS data processing software (Demeter 0.9.24).5 A variety of spectra of
organic and inorganic Cu reference compounds were considered for LCF.6 Inclusion of a reference
spectrum into the combination fit required at least a 10% decrease in the Rf-value, indicating a
significant change to the quality of the fit.
Preliminary experiments
For all the CuO NP amendments using standard soils, preliminary experiments were conducted to
determine the approximate solubility of CuO NP in each soil. CuO NPs were added to soil to
achieve 50 mg/kg, 100 mg/kg and 500 mg/kg Cu. DTPA extractions were conducted after 30 days
of aging (moisture content was maintained at soil field capacity during aging period) to determine
the solubility of CuO NPs in each standard soil. The initial CuO NP concentration for each
- 106 -
treatment was then selected to ensure that the CuO NPs did not completely dissolve over the 30d
experiments.
Properties of soils (unamended) used in this study
Table A2-1: Properties of sampled soils
Soil Type pH Organic carbon
content (%)
Soil type Clay content
(%)
Background Cu
Lufa 2.1 soil 4.7 0.67 Silty sand 3.9 2.95mg/kg Lufa 2.2 soil 5.6 1.71 Loamy sand 8.3 3.4mg/kg Lufa 2.3 soil 5.9 0.66 Silty sand 7.6 3.3mg/kg Lufa 2.4 soil 7.4 1.99 Clayey loam 26.4 18.3mg/kg Arizona soil 7.6 0.54 Clayey loam 32.5 0.9 mg/kg Portugal soil 5.0 1.2 Sandy loam 3.1 9.1mg/kg
Table A2-2: Mass balance and experimental conditions for each treatment
Treatment
Targeted total Cu concentr
ation (mg/kg)
Total Cu measured from
digestion (mg/kg)
DTPA extractable Cu on
D30 (mg/kg)
CaCl2 extractable Cu on
D30 (mg/kg)
Residue Cu (non-extractable Cu2+) at D30 (mg/kg)
CuO NP remained at D30 (mg/kg)
pH Organic carbon content
(%)
Moisture content
Lufa 2.1 soil
100 114 75.0 4.9 18 21 5.0 0.67 16%
Lufa 2.1 soil, pH adjusted
100 112 1.30 0.050 0.33 110 7.4 0.67 16%
Lufa 2.1 soil, OM adjusted
300 287 190 12 47 49 4.9 1.34 16%
Lufa 2.2 Soil
500 503 249 2.6 62 192 5.9 1.71 21%
Lufa 2.2 Soil
250 265 171 1.3 43 54 5.8 1.71 21%
Lufa 2.2 Soil, pH adjusted
500 501 144 1.4 37 320 6.8 1.71 21%
Lufa 2.2 Soil,
moisture
500 481 234 2.3 58 189 5.9 1.71 10%
- 107 -
content adjusted Lufa 2.3
soil 500 539 68 0.37 17 455 6.5 0.66 17%
Lufa 2.4 soil
500 537 128 0.40 32 377 7.2 1.99 22%
Arizona soil
500 544 51 0.17 13 480 7.6 0.54 12%
Portugal soil
500 500 225 0.64 56 219 1.2 16% 16%
Nanoparticle characterization. The properties of the CuO NPs have been previously published6.
Briefly, the primary particle size was 38nm ± 1.7nm (TEM) and shape is spherical. The
hydrodynamic diameter of CuO NP was 560nm±103nm (pH=7), and the -potential was -
16.1mV±1.7mV. The isoelectric point of CuO NP is 8.8 (in 5mM NaNO3), so all CuO NPs are
positively charged in soil pore water in all the treatments.
Derivation of dissolution model
The transformation of the CuO NPs in the soil is given by equation A2-1:
CuO NP(s) ⇄ 𝐶𝐶𝐶𝐶2+𝑑𝑑𝐸𝐸𝐶𝐶𝑓𝑓𝑓𝑓𝑓𝑓𝐸𝐸𝑟𝑟𝑓𝑓𝐶𝐶𝐸𝐸,𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙���������������𝐶𝐶𝐶𝐶[𝐿𝐿] (𝐴𝐴2 − 1)
Where L represents various ligands that can complex with Cu.
According to the assumptions in our manuscript:
𝑑𝑑[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑇𝑇𝑇𝑇,𝑇𝑇𝑑𝑑𝑑𝑑
= 𝑘𝑘𝑑𝑑,𝐷𝐷𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 − 𝑘𝑘𝑟𝑟,𝐷𝐷[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑1
1+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 (𝐴𝐴2 − 2)
Where the subscript 𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 indicates the total surface area (not mass) of CuO NP. 𝑘𝑘𝑑𝑑,𝐷𝐷 ,𝑘𝑘𝑟𝑟,𝐷𝐷 are the
dissolution rate constant and reverse reaction (precipitation) constant with respect to NP surface
area. 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 is the partitioning constant between Cu associated with natural ligands (including both
DOM and soil surfaces, e.g. SOM, clay, iron oxides) and free Cu2+.
𝐴𝐴𝐶𝐶𝐶𝐶𝐶𝐶 = 𝐶𝐶1 ∗ �[𝐶𝐶𝐶𝐶𝐶𝐶]𝑇𝑇𝜌𝜌𝐶𝐶𝐶𝐶𝐶𝐶
�𝑓𝑓
= 𝐶𝐶2 ∗ [𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 (𝐴𝐴2 − 3)
- 108 -
𝜌𝜌𝐶𝐶𝐶𝐶𝑂𝑂 is the density of CuO NP. For spherical particles (e.g. the CuO NP used in this study),
n=2/3.7 The constant C1 takes the initial size of the particles into account. Assuming density and
shape of the particle would not change during the dissolution process, then C1 and 𝜌𝜌𝐶𝐶𝐶𝐶𝑂𝑂 can be
incorporated into C2, which should be a constant during the dissolution process. Thus, equation S3-
2 can be rewritten as:
𝑑𝑑[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑇𝑇𝑇𝑇,𝑇𝑇𝑑𝑑𝑑𝑑
= 𝑘𝑘𝑑𝑑[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 − 𝑘𝑘𝑟𝑟[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑
11+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙
[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 (𝐴𝐴2 − 4)
Where the exponent 2/3 converts concentration to surface area; 𝑘𝑘𝑑𝑑 𝑎𝑎nd 𝑘𝑘 𝑟𝑟 are the rate constants
and are fitted parameters in this model. 𝐾𝐾𝑓𝑓𝑓𝑓𝑙𝑙𝑓𝑓𝑓𝑓𝑑𝑑 is determined from the CuSO4 control experiment
as described in the “Dissolution models” section in the manuscript. Note, [𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑 = [𝐶𝐶𝐶𝐶𝑂𝑂]𝑜𝑜 −
[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑
The following equation was used to numerically fit the experimental data:
[𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑+1= [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑+t*( 𝑘𝑘𝑑𝑑[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3 − [𝐶𝐶𝐶𝐶2+]𝑇𝑇𝑜𝑜𝑑𝑑,𝑑𝑑
𝑘𝑘𝑟𝑟1+𝐾𝐾𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙
[𝐶𝐶𝐶𝐶𝑂𝑂]𝑑𝑑2/3) (A2-5)
Equation A2-5 was used to fit the experimental data to get kd and kr.
- 109 -
Table A2-3. Comparison between CuO NP dissolution measured by XANES and chemical
extraction
The details on XANES fitting result can be found in table A2-4 and Figure A2-1. XANES could not
detect the minimal (<10%) CuO NP dissolution in the high pH Arizona soil. In Arizona soil, with
500mg/kg CuSO4 dosing, ~20% of Cu precipitated as copper carbonate, with most Cu (~80%)
remaining in its complexed form (not precipitating as a new solid phase). This suggested that in the
CuO NP treatment, with only 64mg/Kg Cu dissolved, the amount of Cu ions released from CuO NP
would not precipitate as other Cu phases as the concentration of Cu ions released from CuO NP did
not exceed the solubility limit of Cu with respect to Copper carbonate in Arizona soil .
Treatment % CuO NP dissolved by
XANES (mg/kg)
% CuO NP dissolved by
extraction(mg/kg)
Lufa 2.2 soil, 500 mg/kg
treatment, 10% moisture,
D7
37 32
Lufa 2.2 soil, 500mg/kg
treatment,10% moisture
D31
52 61
Lufa 2.2 soil, 500mg/kg
treatment, 21% moisture
D31
60 62
Arizona soil, 300mg/kg ,
12% moisture content D0
0 1
Arizona soil, 300mg/kg ,
12% moisture content D7
0 9
- 110 -
Table A2-4. Linear combination fitting results of k3-weighted Cu EXAFS spectra (Figure S3-1) for
Arizona soil exposed to 300mg/kg of CuO NP or CuSO4. Samples were fit over a k range of 3-12 Å.
The percentages have inherent ±15% uncertainties. Data are presented with the R factor (Rf) and the
Reduced χ2 parameters to indicate the quality of the fits.
Reference Compounds Fit Parameters
CuO NP
(%)
Cu-
NOM
(%)
Cu-Ferri
(%)
Cu-Carb
(%) R factor Red. χ2
CuO
NP
Day 0 119 - - - 0.03 0.65
Day 7 109 - - - 0.03 0.56
Day 21 110 - - - 0.03 0.52
CuSO4 Day 0 - 44 41 19 0.08 0.60
Day 7 - 60 23 20 0.10 0.67
- 111 -
Figure A2-1. Cu EXAFS spectra (black) and linear combination fits (red) for CuO NP and CuSO4
exposed soil. (a) Arizona soil, (b) Lufa 2.2 soil. Cu-Ferrihydrite and Cu-NOM models were
synthesized using established methods.1,8
- 112 -
0 1 0 2 0 3 0 4 0
0
2 0
4 0
6 0
8 0
1 0 0
T i m e ( d a y s )E
xtr
ac
tab
le C
u
(mg
/k
g d
rie
d s
oil
)
D T P A e x t r a c t i o n , L u f a 2 . 1 s o i l
p H 5 . 0
p H 7 . 4
Figure A2-2. DTPA extractable Cu in Lufa 2.1 soils dosed with 100mg/kg CuO NPs at pH 5.0
(squares) and pH 7.4 (triangles). Bars are standard deviation of the measurements.
Figure A2-2 shows that the overall dissolution of CuO NPs in Lufa 2.1 soil was greatly reduced by
increasing soil pH from 5.0 to 7.4. However, due to the very slow dissolution, the dissolution
kinetics of CuO NPs in pH 7.4 soil could not be accurately modeled, thus the data was not used to
calculate an dissolution rate constant and solubility.
- 113 -
Cross validation of the pH- kd correlation
0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5
0 . 0
0 . 5
1 . 0
1 . 5
T a k e o u t L u f a 2 . 1
{ H+
} ( m o l / L )
kd
(mg
1/3
·kg
1/3
·s-1
)
0 1 0 2 0 3 0 4 0
0
5 0
1 0 0
1 5 0
T i m e ( d a y s )
Dis
so
lve
d C
u (
mg
/kg
)
E x p e r i m e n t a l d a t a
P r e d i t i o n u p p e r 9 5 % C I
P r e d i t i o n l o w e r 9 5 % C I
0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5
0 . 0
0 . 5
1 . 0
1 . 5
T a k e o u t L u f a 2 . 2 2 1 %
{ H+
} ( m o l / L )
kd
(mg
1/3
·kg
1/3
·s-1
)
0 1 0 2 0 3 0 4 0
0
1 0 0
2 0 0
3 0 0
4 0 0
T i m e ( d a y s )
Dis
so
lve
d C
u (
mg
/kg
)
E x p e r i m e n t a l d a t a
P r e d i t i o n u p p e r 9 5 % C I
P r e d i t i o n l o w e r 9 5 % C I
0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5
0 . 0
0 . 5
1 . 0
1 . 5
T a k e o u t l u f a 2 . 2 2 5 0 p p m
{ H+
} ( m o l / L )
kd
(mg
1/3
·kg
1/3
·s-1
)
0 1 0 2 0 3 0 4 0
0
1 0 0
2 0 0
3 0 0
T i m e ( d a y s )
Dis
so
lve
d C
u (
mg
/kg
)
E x p e r i m e n t a l d a t a
P r e d i t i o n u p p e r 9 5 % C I
P r e d i t i o n l o w e r 9 5 % C I
0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5
0 . 0
0 . 5
1 . 0
1 . 5
T a k e o u t L u f a 2 . 2 1 0 %
{ H+
} ( m o l / L )
kd
(mg
1/3
·kg
1/3
·s-1
)
0 1 0 2 0 3 0 4 0
0
1 0 0
2 0 0
3 0 0
4 0 0
T i m e ( d a y s )
Dis
so
lve
d C
u (
mg
/kg
)
E x p e r i m e n t a l d a t a
P r e d i t i o n u p p e r 9 5 % C I
P r e d i t i o n l o w e r 9 5 % C I
0 . 0 0 0 0 0 0 0 . 0 0 0 0 0 5 0 . 0 0 0 0 1 0 0 . 0 0 0 0 1 5 0 . 0 0 0 0 2 0
0 . 0
0 . 5
1 . 0
1 . 5
2 . 0
T a k e o u t L u f a 2 . 1 w i t h O M a j u d s t e d
{ H+
} ( m o l / L )
kd
(mg
1/3
·kg
1/3
·s-1
)
0 1 0 2 0 3 0 4 0
0
1 0 0
2 0 0
3 0 0
T i m e ( d a y s )
Dis
so
lve
d C
u (
mg
/kg
)
E x p e r i m e n t a l d a t a
P r e d i t i o n l o w e r 9 5 % C I
P r e d i t i o n u p p e r 9 5 % C I
Figure A2-3: Cross validation of the correlation between kd and {H+}. The figures on the left are
the correlations after taking out one soil treatment, the figures on the right are the resulting
predictions on the removed soils.
- 114 -
Multivariate regressions:
Table A2-5.Multivariate regression between dissolution rate constant and soil organic matter content and hydrogen ion activity.
Coefficients Standard
Error P-value Lower 95%
Upper 95%
Organic carbon content 0.119 0.185 0.55 -0.356 0.594 {H+} 5.503*105 1.733*105 0.025 1.05*105 9.96*105
Table A2-6.Multivariate regression between solubility and soil organic matter content and hydrogen ion activity.
Coefficients Standard
Error P-value Lower 95% Upper 95%
Organic carbon content 210.6 35.1 0.0018 120 301 {H+} 8.40*106 3.29*106 0.81 -7.62*106 9.30*106
Table A2-7.Multivariate regression between reverse reaction rate constant and soil organic matter content and hydrogen ion activity.
Coefficients Standard
Error P-value Lower 95% Upper 95%
Organic carbon content -0.108 0.0600 0.13 -0.261 0.0456 {H+} -9.17*103 5.60*103 0.16 -2.36*103 5.25*103
As suggested from Table A2-5 and Table A2-6, solubility is correlated with only organic matter
content (P<0.05), and that dissolution rate is correlated with only {H+}(P<0.05). Table A2-7
indicates that kr shows no correlation with either organic matter content or {H+}.
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Appendix 3- Supporting information for Chapter 4: Elucidate CuO NP dissolution behavior and toxicity to wheat (Triticum aestivum) in rhizosphere and bulk soil.
Table A3-1: Total Cu concentration (mean (SD), mg/kg) in soil for each treatmentga
CuO NP CuSO4 Control
Fresh 534.9 (39.4) 307.1 (3.7) 5.9 (0.3)
Aged 510.5 (6.4) 312.5 (1.3) N.A.
Table A3-2: DTPA extractable Cu (mean (SD), mg/kg) in the control treatment before and after plant growth
Bulk soil Loosely attached
soil
Rhizosphere soil
day 0 1.3 (0.1) N.A. N.A.
day 14 1.5 (0.1) 1.6 (0.4) 1.3 (0.9)
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Figure A3-1. Change in DTPA extractable Cu and CaCl2 extractable Cu for 250mg/kg CuO NP
treatment, 250mg/kg and 500mg/kg CuSO4 treatments (without growing plants) over 30 days. Error
bars are standard deviations.
A decreasing trend in both CaCl2 extractable Cu and DTPA extractable Cu were observed in CuSO4
treatments (both 500mg/kg and 250 mg/kg), while an increasing trend in both CaCl2 extractable Cu
and DTPA extractable Cu were observed in 500mg/kg CuO NP treatment. DTPA can extract
~80% of Cu for both 500mg/kg and 250mg/kg CuSO4. The solubility of CuO NP was estimated to
be 312mg/kg by fitting a first-order dissolution kinetic model to the data (R2=0.990) and assuming
80% of the Cu was extractable.
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Figure A3-2. Different soil regions defined in this study9,10. Bulk soil was presumed to be unaffected
by the roots or root exudates due to lack of proximity to the roots. Loosely attached soil clung to
roots but could be shaken free. Rhizosphere soil was shaken off of the roots after air drying.
Nanoparticle characterization6. Primary particle size distribution of the CuO NP was previously
measured by transmission electron microscopy (TEM, Hitachi H-9000 TEM microscope operating at
300 kV). Particles in 10 TEM imagines were counted for primary particle size distribution 278
nanoparticles were counted). The hydrodynamic diameter and zeta potential of CuO NP (80 mg/kg
as Cu) in 5mM NaHCO3 buffer (pH=7) were measured by dynamic light scattering (Zetasizer Nano,
Malvern) 6. The isoelectric points of CuO NP (80mg/kg) in 5mM NaHCO3 buffer and in 5mM
NaNO3 were calculated from measurements of the zeta potential of the particles in suspension over
a range of pH6.
Soil amendment. A nanoparticle suspension containing 644mg CuO NP, 664mg Na2SO4, and
208ml milli-q water (CuO NP treatment), dissolved Cu2+ solution containing 746mg CuSO4 and
208ml milli-q water (CuSO4 treatment), or Na2SO4 solution (664mg Na2SO4 and 208ml milli-q water
(control treatment) was sonicated (Sonic Dismembrators model 550, Energy level 3) for 30s before
adding to 1000kg air dried Lufa 2.2 soil. The Na2SO4 treatment was used to control for any effect of
added SO4 ions to plants.
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The soil-suspension mixture (for NPs) or soil-solution mixture (for dissolved species) was
hand-mixed using a polycarbonate spoon for 20 min for each treatment. The soil was aged in an
incubator, and the moisture content was maintained at field capacity by daily addition of milli-Q-
water. For the control (no Cu added), only the fresh treatment protocol was followed.
Soil and Plant digestion. The soil samples collected before the plant growth experiment were
analyzed for total Cu content to confirm the amount of Cu in each treatment11. Briefly, 1g of air-dried
soil was digested with concentrated nitric acid and 30% hydrogen peroxide at 95 ºC using a hot plate.
After digestion, the samples were centrifuged at 3000 rpm for 10 min, followed by filtration using a
0.45um filter to remove fine particles in the supernatant. The filtered supernatant was diluted with
Milli-Q water and acidified with 20% HNO3 (final HNO3 concentration was 2%) for ICP-MS (Agilent
7700x) analysis. The plant total Cu concentration was determined using acid digestion according to a
modified version of U.S. EPA Method 3050B (1996).11 The dried root and shoot tissues were digested
with concentrated nitric acid and 30% hydrogen peroxide at 95°C using a hot plate. Due to the small
sample size, the chemical additions were 1/10th the volumes listed in EPA Method 3050B11. After
digestion, the samples were centrifuged at 3000 rpm for 15 min and the supernatant was diluted with
Milli-Q water to a final HNO3 concentration of ~5% for analysis by ICP- MS (Agilent 7700x).
ICP-MS measurement. Germanium was used as an internal standard for quality control. The range
of the calibration curve used in each treatment was between 0.005mg Cu/kg to 0.5mgCu/kg. All the
extraction samples, digestion samples were diluted within the range of the calibration curve. All the
samples were diluted and acidified right after sample collection, and measured within two days after
preparation. The calibration samples were measured each time when measuring extraction/digestion
samples (with all the calibration curves measured, R2 values were > 0.998 at all times).
Cytoviva analysis. The interaction between roots and NPs were visualized in fresh roots after a
rinsing step in 10-3 M KCl, using a DF-HSI system (CytoViva Inc., USA). This enhanced resolution
dark–field microscope system (BX51, Olympus, USA) was equipped with a 150 W halogen light source
for the dark–field sample illumination (Fiber-Lite®, Dolan-Jenner, USA), and a hyperspectral camera
(CytoViva Hyperspectral Imaging System 1.4). The roots were observed with 600× and 1000×
magnification. Hyperspectral images were acquired using 60% light source intensity and 0.1 s
acquisition time per line. Each pixel of the hyperspectral image contains its light reflectance spectrum
ranging from 400 to 1000 nm with a step of 1.5 nm.
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Hyperspectral images (datacubes) were acquired on samples of CuO NPs mixed with Lufa 2.1
soil for 15min using ENVI 5.2 software (Exelis Visual Information Solutions, CO, United States).
After normalizing the datacubes for the lamp signal, spectral pre-libraries of CuO NPs in gels were
collected from these reconstructed RGB images based on pixel purity index as described in an earlier
article12. The spectra contained in the pre-library that were not specific to the CuO NP were filtered
by matching the datacubes against negative controls (soil without CuO NP, and control roots) using
a Spectral Angular Mapping algorithm (SAM, ENVI 5.2), an algorithm comparing angles between
vectors. Spectra in the pre-library that matched spectra of pixels in the control hyperspectral images
were considered as unspecific false positives and removed from the pre-library, whereby two vectors
(i.e. spectra) with angles ≤0.085 rad were considered as similar. The remaining spectra constitute the
final CuO NP library (i.e. exclusively containing specific hyperspectral CuO NP signature). The spectral
library of soil only (background), CuO NP spectral libraries, and the SAM results to test their
specificity on negative controls are shown in Figure S3.
The CuO NP library was used to perform SAM on hyperspectral images of exposed and
control roots, using the threshold angle of 0.085 rad, with two replicate pictures per condition on
several root areas. Each pixel in the images matching the hyperspectral signature of CuO NPs was
highlighted in red.
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Figure A3-3. (A) Spectral library of the CuO-NPs. The spectral library has been built using
datacubes of CuO mixed with hydrated soil. (B) Example of SAM (Spectral Angle Mapping) results
to test for the specificity of the spectral library using positive controls (soil containing CuO NPs) or
negative controls (soil without CuO NPs or control root) images. The pixels containing the spectral
signal of CuO NP are highlighted in red (bottom line). Note that only the positive control contained
the signal of CuO-NP.
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Figure A3-4. DTPA extractable Cu on bulk soil and bulk bottom soil in different Cu treatments
In all treatments, no significant differences (P<0.05, unpaired t-test) were found between DTPA
extractable Cu in bulk soil and bulk bottom soil, suggesting no vertical transport of Cu in all
treatments. Error bars are one standard deviation.
0
100
200
300
CuO Fresh CuO Aged CuSO4 Fresh CuSO4 Aged
Ext
ract
able
Cu
conc
entr
atio
n (m
g/kg
)DTPA extraction
bulk soil
bulkbottom soil
- 122 -
Table A3-3: Samples that provided sufficient soil for DTPA extraction for rhizosphere soil and loosely attached soil
Rhizosphere soil Loosely
attached soil
Plant number
CuO NP Fresh
CuO NP aged
Control CuO NP Fresh
CuO NP aged
Control
1
X X
X X 2 X X X X 3 X
X X
X
4
X X X 5 X
X X X X
6
X X X X
Table A3-4: Samples that provided sufficient soils for CaCl2 extraction for rhizosphere soil and loosely attached soil
Rhizosphere soil Loosely attached soil
Plant number
CuO NP Fresh
CuO NP aged
Control CuO NP Fresh
CuO NP aged
Control
1 X
X
X X 2
X X X X
3 X
X X
X 4 X
X X
5
X X X X 6
X X X X X
- 123 -
Figure A3-5. Representative photos showing Cu toxicity led to shortened root and/or root
compactness in fresh CuSO4 treatment (a) aged CuSO4 treatment (b) and aged CuO NP treatment
(d), whereas these toxic effects were not observed in fresh CuO NP treatments (c) and in control
treatments (e).
a b
c d e
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Figure A3-6. Hyperspectral imaging of plant roots grown in soil with CuO-NP, CuSO4 or Na2SO4
(control) freshly amended or after aging. Roots exposed to CuSO4 (both after soil aging or not)
showed a brown-damaged (necrotic) zone, that was not found on any of the CuO NP exposed
roots.
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Figure A3-7. Mean concentration of Cu (mg/kg) in wheat tissue (dry weight): a) Cu concentration
in shoots, b) Cu concentration in roots. Error bars show ± 1 SD. Capital letters indicate significant
different differences between groups (one way ANOVA test, P≤0.05).
Appendix 4. Explanation on the solubility of CuO NP in Chapter 3
When reviewing Chapter 3 of this thesis, we recognized an inconsistency in the data set: In section
3.2.6, 500mg/kg CuSO4 was shown to be completely soluble in Lufa 2.2 soil, i.e. having no
observable Cu precipitation (<~5 wt%) as suggested from the XANES spectra. However, the
modeled solubility of CuO NP in Lufa 2.2 soil was ~300mg/kg. For the equilibrium condition to
exist, these two scenarios (adding the same mass of either CuO NPs of CuSO4 salt) should have led
to the same final condition, but they did not. One reasonable explanation is presented schematically
in Figure A4-1. The CuSO4 treatment involved missing of Cu ions thoroughly through the soil,
providing a high degree of mixing between Cu and the soil NOM (Figure A4-1, b). In contrast, the
CuO NPs are not mixed as homogeneously in soil compared to CuSO4 treatment (Figure A4-1,a).
Instead, there is a localized dissolution-sorption equilibrium near the CuO NPs, and therefore not all
of the soil NOM is contacted with Cu. Thus, in CuO NP treatment, there is a fraction of SOM that
is not occupied by the Cu ions released by CuO NPs when the localized solubility of CuO NP has
been reached, resulting a less solubility of CuO NP compared to the sorption capacity of Cu in soil.
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Figure A4-1. Conceptual model of Cu speciation in Lufa 2.2 soil a) CuO NP treatment b) CuSO4
treatment. The black dots in a) represents CuO NPs, and the shadowed brown area suggest Cu-OM
complex. The Cu-OM complex on b) is evenly distributed in soil, whereas in a) it is only evenly
distributed around particles. Figure A4-1. Conceptual model of Cu speciation in Lufa 2.2 soil a)
CuO NP treatment b) CuSO4 treatment. The black dots in a) represent CuO NPs, and the
shadowed brown area represents the localize area around the CuO NPs where the released Cu has
formed Cu-OM complexes. The Cu-OM complex in b) is evenly distributed in soil because the
added Cu ion can be uniformly mixed into the soil, contacting all of the soil NOM. While the CuO
NPs are uniformly distributed in soil, the released ions do to contact all of the soil NOM. This
lowers the overall "apparent" solubility in soil relative to the case for Cu ion addition.
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(3) van Zomeren, A.; Comans, R. N. J. Measurement of humic and fulvic acid concentrations and dissolution properties by a rapid batch procedure. Environ. Sci. Technol. 2007, 41 (19), 6755–6761.
(4) Webb, S. M. SIXpack: a graphical user interface for XAS analysis using IFEFFIT. Phys. Scr. 2005, 2005 (T115), 1011.
(5) Ravel, B.; Newville, M. ATHENA, ARTEMIS, HEPHAESTUS: Data analysis for X-ray absorption spectroscopy using IFEFFIT. J. Synchrotron Radiat. 2005, 12 (4), 537–541.
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(6) Gao, X.; Spielman-Sun, E.; Rodrigues, S. M.; Casman, E. A.; Lowry, G. V. Time and nanoparticle concentration affect the extractability of Cu from CuO NP amended soil. Environ. Sci. Technol. 2017, 51 (4), 2226–2234.
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(10) Turpault, M. P. Sampling of rhizosphere soil for physico-chemical and mineralogical analyses by physical separation based on drying and shaking. Handb. methods used Rhizosph. Res. Swiss Fed. Res. Inst. WSL, Birmensd. 2006, 196–197.
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