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Quantifying PPCP interaction with dissolved organic matterin aqueous solution: Combined use of fluorescence quenchingand tandem mass spectrometry
Selene Hernandez-Ruiz a, Leif Abrell b, Samanthi Wickramasekara b, Benny Chefetz c,Jon Chorover a,b,*aDepartment of Soil, Water and Environmental Science, University of Arizona, 1177 E 4th St, Tucson, AZ 85721, USAbArizona Laboratory for Emerging Contaminants, University of Arizona, 1040 East 4th St, Tucson, AZ 85721, USAcDepartment of Soil and Water Sciences, The Hebrew University of Jerusalem, P.O. Box 12, Rehovot 76100, Israel
a r t i c l e i n f o
Article history:
Received 17 August 2011
Received in revised form
11 November 2011
Accepted 20 November 2011
Available online 26 November 2011
Keywords:
DOM
IHSS
Wastewater
Pharmaceuticals
LC-MS/MS
Fluorescence
Interaction
* Corresponding author. Department of Soil,USA. Tel.: þ1 520 626 5635; fax: þ1 520 621 1
E-mail addresses: selene.hernandez27@arizona.edu (S. Wickramasekara), chefetz0043-1354/$ e see front matter ª 2011 Elsevdoi:10.1016/j.watres.2011.11.061
a b s t r a c t
The documented presence of pharmaceuticals and personal care products (PPCPs) in
water sources has prompted a global interest in understanding their environmental fate.
Dissolved organic matter (DOM) can potentially alter the fate of these contaminants in
aqueous systems by forming contaminant-DOM complexes. In-situ measurements were
made to assess the interactions between three common PPCP contaminants and two
distinct DOM sources: a wastewater treatment plant (WWOM) and the Suwannee River,
GA (SROM). Aqueous DOM solutions (8.0 mg L�1 C, pH 7.4) were spiked with a range of
concentrations of bisphenol-A, carbamazepine and ibuprofen to assess the DOM fluo-
rophores quenched by PPCP interaction in excitationeemission matrices (EEM). Interac-
tion effects on target analyte (PPCP) concentrations were also quantified using direct
aqueous injection ultra high performance liquid chromatography tandem mass spec-
trometry (LC-MS/MS). At low bisphenol-A concentration, WWOM fluorescence was
quenched in an EEM region attributed to microbial byproduct-like and humic acid-like
DOM components, whereas carbamazepine and ibuprofen quenched fulvic acid-like flu-
orophores. Fluorescence quenching of SROM by bisphenol-A and carbamazepine was
centered on humic acid-like components, whereas ibuprofen quenched the fulvic acid-
like fluorophores. Nearly complete LC-MS/MS recovery of all three contaminants was
obtained, irrespective of analyte structure and DOM source, indicating relatively weak
PPCP-DOM bonding interactions. The results suggest that presence of DOM at
environmentally-relevant concentration can give rise to PPCP interactions that could
potentially affect their environmental transport, but these DOM-contaminant interac-
tions do not suppress the accurate assessment of target analyte concentrations by
aqueous injection LC-MS/MSMS.
ª 2011 Elsevier Ltd. All rights reserved.
Water and Environmental Science, University of Arizona, 1177 E 4th St, Tucson, AZ 85721,647.gmail.com (S. Hernandez-Ruiz), abrell@u.arizona.edu (L. Abrell), samanthw@email.@agri.huji.ac.il (B. Chefetz), chorover@cals.arizona.edu (J. Chorover).ier Ltd. All rights reserved.
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4944
1. Introduction concentrations in wastewater effluents are in the mg L�1 range
The chemical quality of treated wastewater is receiving
increased attention, particularly as human population growth
makes it a progressively larger component of the near-surface
hydrologic cycle. Treated wastewater contains partially
degraded or non-degraded endocrine disrupting compounds,
pharmaceuticals and personal care products (PPCPs) that can
make their way into freshwater sources with treatment plant
discharges (Joss et al., 2008; Dickenson et al., 2011). Human
and ecosystemhealth concerns derive from the fact that some
PPCPs are known to cause cancer, mutations, and/or impede
the reproduction and hormone function of living organisms
(Liu et al., 2005; Zhou et al., 2007). Over the past decade,
technical improvements in chemical analytical methods have
enabled the detection and quantification of these contami-
nants in water at sub-parts-per-trillion levels in environ-
mental samples including wastewater effluents and receiving
surface and ground water sources (Capdeville and Budzinski,
2011). Despite the documented adverse effects of PPCPs,
their persistence, transport, and fate in thewater cycle are not
well understood.
Organic contaminants co-occur with dissolved
natural organic matter (DOM), which is ubiquitous and
compositionally diverse in natural waters (Leenheer and
Croue, 2003). It comprises a heterogeneous mixture of vari-
ously aggregated organic molecules (<0.45 mm) deriving from
decaying biomass, including biomolecules and their degra-
dation products. Interactions between DOM and PPCPs can
potentially alter not only contaminant bioavailability and
transport, but also accurate detection and quantification of
these compounds (Lajeunesse and Gagnon, 2007). Aqueous
solubility and hydrophobicity parameters such as the octa-
nolewater partitioning coefficient (Kow) are insufficient
predictors of PPCP-DOM interaction (Kwon and Armbrust,
2008; Yamamoto et al., 2003) because intermolecular mecha-
nisms of association other than the hydrophobic effect e e.g.,
hydrogen bonding, cation bridging e can affect the bonding to
DOM of these relatively polar compounds (Tolls, 2001; Pan
et al., 2009). We hypothesize that physico-chemical proper-
ties of both DOM and PPCPs, including charge and function-
ality, influence the types of bonds formed between them. To
begin testing this hypothesis, the present study employed (i)
fluorescence spectroscopy to probe DOM-PPCP molecular
interaction, and (ii) direct aqueous injection tandem mass
spectrometry to measure DOM impacts on analyte quantifi-
cation. The study sought to probe PPCP-DOM interactions in
the aqueous environment, and at environmentally relevant
pH, ionic strength and dissolved organic carbon (DOC)
concentration.
In the present study, three model compounds e ibuprofen,
bisphenol-A, and carbamazepine e were used to probe PPCP-
DOM interactions in circumneutral pH aqueous systems
(Table 1). Ibuprofen (IBU), which is anionic at circumneutral
pH, is the third most used analgesic worldwide, and is
consumed on average at a level of 1200e1600 mg per person
per day. Of this, ca. 29% is estimated to be metabolized,
whereas the remainder is introduced to the water cycle
(Meulenberg et al., 2005; Kagle et al., 2009). Reported IBU
(Pedrouzo et al., 2007), whereas surface freshwaters typically
show lower values. Bisphenol-A (BPA), which is uncharged at
circumneutral pH, is used worldwide in the production of
plastics (Stavrakakis et al., 2008) and is known to bind to
estrogen receptors and may also be carcinogenic (Vom Saal
et al., 1998; Soto and Sonnenschein, 2010). BPA has been re-
ported to occur at 10e20 mg L�1 in several freshwaters (Kolpin
et al., 2002; Jjemba, 2006) and at several-fold higher values in
wastewaters (Lee et al., 2000e02; Wickramasekara et al., in
press). Carbamazepine (CBZ), which is also uncharged at cir-
cumneutral pH, is a potent pharmaceutical commonly used
for epilepsy and bipolar disorder (Bai et al., 2008). It has been
reported at concentrations ranging from ca. 0.2 mg L�1 in
freshwaters (Focazio et al., 2008) to >5 mg L�1 in wastewater
effluents (Pedrouzo et al., 2007; Jjemba, 2006; Wickramasekara
et al., in press) and can accumulate in plant biomass (Shenker
et al., 2011).
Given the variation in charge and structural chemistry of
these three model contaminants (Table 1), we anticipated
that their impact on DOM fluorescence might also depend on
DOM source and structure. Hence, we employed two distinct
bulk DOM sources, one from a wastewater treatment plant
and the other from a surface freshwater. Two complemen-
tary methods were employed. Fluorescence spectroscopy
was used to elucidate molecular components of DOM
interacting with the target PPCPs, and tandem mass spec-
trometry was used to measure the impact of associated bond
formation on analyte chromatographic separation and
quantification.
2. Materials and methods
2.1. Extraction and characterization of DOM
The wastewater DOM (WWOM) was obtained by pore water
extraction of primary municipal wastewater treatment plant
sludge (Tucson, AZ). Within an hour of collection, wet sludge
was centrifuged at 21,875 g for 20 min and the supernatant
solution filtered through 0.7 mm and 0.45 mm hydrophilic
polypropylene filters to obtain WWOM. The filtrate was
freeze-dried for characterization and experiments. The
freshwater DOM-Suwannee River natural organic matter
(SROM) e was purchased in freeze-dried form from the
International Humic Substances Society (IHSS). The SROMhad
been extracted from the Suwannee River in GA (USA) via
reverse osmosis with a final in-line 0.4 mm filter followed by
a desalting step mediated by cation exchange resin (Hþ form)
(Serkiz and Perdue, 1990).
Prior to characterization, freeze-dried DOM samples were
suspended in nanopure water and incubated at 7 rpm for 24 h
on an orbital shaker for analysis of total organic carbon (TOC)
using high temperature combustion and infrared detection of
CO2 on a Shimadzu TOC-V CSH TOC/TN analyzer (Columbia,
MD). Standard calibrations for organic carbonweremadewith
oven-dried potassium hydrogen phthalate.
Infrared spectra were collected in transmission mode by
drying four 1 mL aliquots of 20 mg L�1 DOM solution (C basis)
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4 945
onto germanium windows followed by collection of Fourier
Transform Infrared (FTIR) spectra with a Nicolet MagnaeIR
560 spectrometer equipped with a CsI beam splitter, DTGS-
detector and OMNIC software (Chorover et al., 2004;
Omoike and Chorover, 2006). A blank germanium window
was subtracted from sample data as background, and
a 20 min purge time was employed between data collection
at 400 scans over the spectral range of 500e3800 cm�1 at
4 cm�1 resolution.
Apparent molar mass distributions for each sample were
determined via high pressure size exclusion liquid chroma-
tography (HPLC-SEC) using a Waters 600 HPLC (Milford, MA)
unit equipped with multisolvent delivery 600 pump, 717 plus
auto sampler, and 996 photodiode array (PDA) detector oper-
ating at 280 nmwavelength. A guard columnand two stainless
steel (8� 300 mm) SEC columns (MCXGPC 1000 & 100,000 �A,
PSS Polymer Standard Service-USA, Inc Warnick RI) were
connected in series to give a linear relationship between log
molar mass and elution time for polystyrene sulfonate (PSS)
standards (Omoike and Chorover, 2006; Navon et al., 2011).
Circumneutral pH values (ca. 7.4) were maintained using
a phosphate buffer solution prepared by bringing 60 mL of
stock 400 mM sodium phosphate buffer solution to 1.0 L with
filtered (0.2 mm) nanopure water. DOM solutions were
prepared in this background electrolyte to give 30 mg L�1
Table 1 e Physico-chemical properties of PPCPs used in this st
PPCP Structure at pH 7.4 pKa Log Kow Log Dow
d
IBU 5.2d 3.72b 1.15c
0.36b
BPA 9.73b 3.32b 3.43b
CBZ 2.3c 2.93e 2.67b
a Collision Energy (CE), Capillary Voltage (CV), Cone Voltage (CNV), and D
b Nghiem and Hawkes (2009).
c Nghiem et al. (2005).
d Jjemba (2006).
e Bai et al. (2008).
f Stavrakakis et al. (2008).
g Gomez et al. (2006).
h Zhang and Zhou (2007).
i Cahill et al. (2004).
j Kuster et al. (2008).
organic C. PSS standards (w2.5 mgmL�1 from 910e48,600 Da)
and 4-ethylbensulfonic acid (186 Da standard) were used for
linear calibration of retention time against log molar mass
(Cabaniss et al., 2000). For each calibration or sample run,
100 mL of solution were injected onto the SEC columns in iso-
cratic mode at a flow rate of 1.0 mLmin�1. Number-average
and weighted-average molar mass (Mn and Mw, respectively),
and polydispersity (r) were calculated in the Empower Pro
software program (Waters 2002, Medford, MA) according to:
Mw ¼XN
i¼1hiðMiÞ=
XN
i¼1hi (1)
Mn ¼XN
i¼1hi=
XN
i¼1
�hi
Mi
�(2)
r ¼ Mw=Mn (3)
where hi andMi are the height andmolarmass, respectively of
the sample SEC-HPLC curve at elution volume i.
2.2. PPCP incubations
All solutions and incubations were prepared on a mass basis
in baked (550 �C, 4 h) amber glass bottles with Teflon caps.
Stock IBU, BPA and CBZ solutions were prepared in 250 mL
udy.
Parent toaughter (m/z)
CEa (eV) CVa (kV) CNVa (V) DTa (�C)
205/ 161g,i,j 8 2.8 20 300
2.8 170
227/ 212f 18 3.3 40 150
3.4 150
237/ 194g,h 16 2.8 50 1502.5 150
esolvation Temperature (DT).
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4946
nanopure water with 1.0 mL of MeOH added to enable
complete dissolution in 48 h. Experimental solutions for IBU
and CBZ were then prepared in an LC-MS/MS-compatible
(volatile) 24 mM NH4HCO3 buffer (1.4 g L�1 of NH4HCO3 in
nanopure water) back-titrated with 0.06 M HCl to pH 7.4.
Freeze-dried WWOM and SROM were likewise dissolved in
24 mM NH4HCO3 buffer to give 16 mg L�1 DOC stock concen-
tration, as verified with Shimadzu TOC-V CSH TOC/TN
analyzer (Columbia, MD). Samples were incubated for 24 h in
shaker at 100 rpm to attain sorption equilibrium (Yamamoto
et al., 2003; Polubesova et al., 2007; Ilani et al., 2005) in dark
amber bottles prior to LC-MS/MS and fluorescence spectros-
copy analyses.
DOM-PPCP incubations were performed in triplicate by
adding 2 mL of each DOC stock (16 mg L�1) to 2 mL of various
concentration PPCP stock solutions for final reactor volume of
4.0 mL, DOC at 8 mg L�1, and BPA at 100, 200, or 1000 mg L�1, or
IBU or CBZ at 10, 20, 200, and 1000 mg L�1. Experiments with
BPA were limited to the higher concentrations because of
a higher limit of LC-MS/MS detection in the aqueous buffer
solution for BPA (50 mg L�1) relative to those for IBU or CBZ
(0.16 and 0.15 mg L�1, respectively). The 4 mL bottles were
incubated (no headspace) for 24 h in an end-over-end rotating
mixer, followed by separation into 1.0 and 3.0 mL aliquots for
LC-MS/MS and fluorescence excitation-emission analyses,
respectively. Controls included DOM alone (no PPCP) to
account for possible analyte recoveries in excess of experi-
mental spikes, PPCP at the respective concentrations alone
(no DOM) to measure recovery from aqueous buffer solutions
and to confirm that analyte fluorescence was insignificant
relative to DOM across the excitation-emission range
employed.
2.3. Fluorescence spectroscopy
To elucidate DOM-PPCP interactions, excitation-emission
matrices (EEMs) of DOM were collected in the absence and
presence of PPCP analytes. EEMs were obtained with a Jobin
Yvon Horiba/Spex Fluoromax-4 fluorometer (Edison, NJ)
equipped with a xenon lamp as the excitation source. The
3.0 mL subsample for each treatment was placed in a square
quartz cuvette cell (light path 10 mm� 10 mm) for EEM
collection. The signal to reference detector ratio was collected
and EEMs were produced using the FluorEssence software
with excitation and emission wavelength ranges of
200e450 nm (5 nm slit) and 250e650 nm (2 nm slit), respec-
tively, both at 5 nm increments. Quenching EEMs were
calculated by subtracting EEMs for the DOM-PPCP treatment
from those of corresponding DOM alone. The contribution of
PPCPs themselves to total fluorescence in DOM-PPCP systems
was found to be negligible relative to that of DOM, such that
quenching EEMs represent accurately the diminished fluo-
rescence of DOM in the presence of PPCPs.
2.4. Fluorescence quantification
Fluorescence intensity was integrated beneath each of five
EEM regions previously characterized as (I) “tyrosine-like”, (II)
“tryptophan-like”, (III) “fulvic acid-like”, (IV) “microbial byproduct-
like”, and (V) “humic acid-like” (Fig. 1C) (Chen et al., 2003). The
volume (Fi) beneath region “i” of the EEM surfacewas obtained
according to
Fi ¼XEx
XEm
IðlExlEmÞDlExDlEm (4)
where Dlex is the excitation wavelength interval, Dlem is the
emission wavelength interval and I (Dlex, Dlem) is the fluo-
rescence intensity >104 at each excitation-emission wave-
length pair. The intensity cut-off of 104 was applied to remove
background noise (fluorescence values at emission wave-
lengths smaller than excitation). The total number of data
points >104 (Ni) for each region were counted to produce the
fractional projected excitation-emission factor (Fi). The
normalized fluorescence intensity volume beneath region ‘i’
of the DOM sources (Fin) was obtained from Fin¼ Fi*Fi, and
from this the fluorescence percentage of reach region was
calculated (Pin¼Fin/FT,n*100%), where FTn is the total
normalized EEM fluorescence deriving from all regions i
(FTn¼ SFin).
2.5. PPCP spike recovery by LC-MS/MS
The recovery of PPCPs following incubation with and without
DOM solutions was accomplished using an Acquity Ultra
Performance Liquid Chromatograph and triple quadruple
Quattro Premier XE mass spectrometer, equipped with
a sample organizer (Waters Corp., Milford, MA). Calibration
standards for IBU, BPA, and CBZ, were evaluated in the same
matrix as experimental treatments (24 mM NH4HCO3 buffer,
pH 7.4). Analytes were separated following 5.0 mL direct
injections through an Acquity UPLC BEH C18 (1.7 um;
2.1 mm� 50 mm) column maintained at 40 �C with a gradient
mobile phase starting at 80:20% water:acetonitrile and ending
at 100% acetonitrile at flow rate 0.3 mlmin�1. Positive mode
electrospray ionization (ESIþ) was used for CBZ, whereas
negative mode (ESI�) was used for IBU and BPA. Recovery was
quantified using multiple reaction monitoring (MRM). Frag-
mentation transitions, collision energies (CE), capillary volt-
ages (CV), cone voltages (CNV), and desolvation temperatures
(DT) are provided in Table 1. Mass lynx software (Waters
Corp.) was used to for identification and quantification of
analytes. An ANOVA/Tukey’s statistical test (95% CI) was used
in Origin 8.5 (Northampton, MA) to assess variance in recov-
eries of all treatments.
3. Results
3.1. Characterization of WWOM and SROM
Chemical analyses indicate that the mass fraction of carbon
is higher for SROM than for WWOM (Table 2). The measured
values for SROM are consistent with those reported previ-
ously (Serkiz and Perdue, 1990). Although the pH and elec-
trical conductivity (EC) of freshly prepared DOM solutions
are shown to vary by source, both samples were normalized
to pH 7.4 and an EC 1200 mS cm�1 for subsequent
experiments.
Table 2 e Baseline characterization for DOM sources.
OC (%) pH EC (mS cm�1) ε (Lmol�1 C cm�1)
SROM 54� 4 5.5� 0.7 31.4 � 2.3 347� 15
WWOM 32� 6.5 7.4� 0.3 1178� 40.1 32� 7
Organic carbon percentage in w/v (OC), Electrical conductivity (EC),
and Molar absorptivity (ε).
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4 947
3.1.1. Fourier transform infrared spectra of WWOM andSROMTransmission FTIR spectra of WWOM and SROM are overlain
for direct comparison in Fig. 2. Both samples display broad
peaks associated with hydroxyl stretching in the
3000e3500 cm�1 range (Leenheer, 1981), and symmetric/
asymmetric CH stretching of eCH3 and eCH2 moieties at
2840e2930 cm�1 (Swift, 1996). SROM displays a major peak at
1731 cm�1 from ester and/or C]O carboxyl stretching
(Swift, 1996), and a peak at 1610 cm�1 attributed to the
asymmetric stretch of eCOO� Swift (1996). Additionally,
SROM shows a peak at 1429 cm�1 ascribed to CeH deforma-
tion of CH2 or CH3 (Chefetz et al., 1998), at 1094 cm�1 due to
CeC stretching of aliphatics (Swift, 1996), and a peak at
829 cm�1 due to CeH out-of-plane bend of condensed
aromatic compounds (Santamaria et al., 2006; Swift, 1996).
Conversely, the WWOM sample displays a peak at 1796 cm�1
that could be attributed to ester C]O stretching (Stuart and
Ando, 1996) and a peak at 1658 cm�1 from C]O vibration of
amides (amide I vibration) associatedwith proteins (Leenheer,
Fig. 1 e Fluorescence EEMs of bulk WWOM (A) and SROM (B). C)
et al., 2003): (I) Tyrosine-like, (II) Tryptophan-like, (III) Fulvic acid
D) Regional normalized fluorescence percentage in accordance
(24 mM NH4HCO3 of pH 7.4 buffer and 8.0 mg LL1 of DOC).
1981; Omoike and Chorover, 2006). A minor amide II peak is
also present (1540 cm�1) confirming a greater relative preva-
lence of protein acious material in WWOM than in SROM
(Chefetz et al., 1998). The WWOM sample also shows peaks at
1356 cm�1from symmetric COO� stretching and CeH bending
of aliphatics (Swift, 1996). The most intense peaks in WWOM
occur around 1000 cm�1 and correspond to the CeO and
CeOeC vibrations of polysaccharides (Chefetz et al., 1998;
Omoike and Chorover, 2006). Overall, the FTIR data suggest
stronger plant derivation of SROM,which is enriched in lignin-
derived products, whereas prevalence of amide and carbo-
hydrate functionalities in WWOM is consistent with domi-
nantly microbial derivation associated with biological waste
water treatment (Hudson et al., 2007).
3.1.2. High performance size exclusion chromatographyThe SROM exhibits a weight average molar mass (Mw) of
2466 Da and polydispersity of three (Fig. 3), consistent with
a prior report of 2190 and 2320 Da for OM in Suwannee River
water (Chin et al., 1994). Similarly, values ranging from 2200 to
2300 Da have been reported for Suwannee River fulvic acid
(Chin et al., 1994; Cabaniss et al., 2000; Zhou et al., 2000). The
Mw and Mn values for WWOM are lower, and the chromato-
gram reveals a subpopulation of small molecules ca. 192 Da
and a large polydispersity value of 12, indicating significant
heterogeneity. Mw values for both DOM types are within the
range reported for DOM from soils, freshwater sources, and
wastewater sources (Leenheer, 1981; Cabaniss et al., 2000;
Chin et al., 1994; Li et al., 2005).
Location of five 1 operationally-defined EEMs regions (Chen
-like, (IV) Microbial byproduct-like, and (V) Humic acid-like.
with regions specified in (C) for bulk SROM and WWOM
Fig. 2 e Transmission FTIR spectra of SROM and WWOM.
Fig. 3 e HPLC-SEC chromatograms for SROM and WWOM
(24 mM sodium phosphate buffer of pH 7.4 and 30 mg LL1
of DOC). Weight-average molar mass (Mw), molar mass of
peak location (MP), number-average molar mass (Mn), and
polydispersity Index (r[Mw/Mn) are indicated.
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4948
3.1.3. Fluorescence spectroscopy of WWOM and SROMMolecular fluorescence is the result of energy emitted during
electron return to the ground state following radiation-
induced excitation (Lakowicz, 1999). Since fluorescence
spectra vary with excitation wavelengths and DOM composi-
tion, excitation-emission matrices (EEMs) are often used to
provide information on the prevalence and structural
composition of DOM fluorophore mixtures (Hudson et al.,
2007; Her et al., 2003).
No correction for inner filtering effects was required as
DOC concentrations (8 mg L�1) were well below threshold
values (ca. 25 mg L�1) for which inner-filtering effects are
observed (Hudson et al., 2007). The selected concentration is
representative of aqueous environmental DOM concentra-
tions in surface water and wastewater solutions (Henderson
et al., 2009; Serkiz and Perdue, 1990), and it enabled collec-
tion of data in the low (200e250 nm) excitation range of EEM
that was often neglected in prior research, but that provides
important information about soluble proteins and lower
molar mass DOM components of potential high reactivity
such as fulvic-acid like molecules.
Fluorescence EEMs for the DOM samples are consistent
with sample origins and literature review (Fig. 1-A and B). The
SROM EEM shows three peaks: a minor intensity peak at Ex/
Em 325/450 ascribed to humic-acid-like molecules in region V,
a secondary peak at Ex/Em 250/450 nm and a high intensity
peak at Ex/Em 220/450 nm attributed to fulvic acid-like
molecules of region III (Chen et al., 2003; Wu et al., 2003; Her
et al., 2003; Henderson et al., 2009). The WWOM EEM shows
a distinct peak at 275/350 nm, which has been reported
previously for wastewater effluents and is attributed to
microbial byproducts (Chen et al., 2003) such as amino acids
including tryptophan and tyrosine (Hudson et al., 2007; Her
et al., 2003; Henderson et al., 2009)and polysaccharides
(Her et al., 2003; Chen et al., 2003).
The SROM sample is rich in aromatic, humic acid-like
components that account for 88% of total EEM fluorescence,
whereas 11% is attributable to fulvic acid-like fluorescence. In
contrast,WWOMdisplayed greater fluorescence in fulvic acid-
like (33%) relative to humic acid-like (49%) regions. Moreover,
the EEM region defined as “microbial byproduct-like”
accounted for up to 10% of the total WWOM fluorescence
(Fig. 1D).
3.2. Fluorescence quenching of WWOM and SROM byPPCPs
DOM fluorescence is diminished (quenched) if static or
dynamic (collisional) interaction occurs with PPCP molecules.
In the case of the dynamic process, the quencher
(PPCP compound) collides with the fluorophore(s) during the
lifetime of the excited state, whereas static quenching is the
result of ground statemolecular interactions (Lakowicz, 1999).
Although fluorescence quenching does not discern dynamic
from static interaction (doing so requires additional fluores-
cence lifetime measurements at different temperatures), the
moieties undergoing the greatest quenching effect are
assumed to be directly involved in contaminant association
(Sun et al., 2007; Yamamoto et al., 2003). Prior studies have
reported differential quenching of DOM fluorophores during
molecular interaction with polar contaminant molecules
(Yamamoto et al., 2003; Bai et al., 2008), indicating the
potential for quenching EEMs to provide a fingerprint of DOM-
PPCP interaction.
Regions of DOM fluorescence that were quenched by IBU,
BPA and CBZ in the present work are depicted as difference
EEMs (DOM alone minus DOM-PPCP) in Figs. 4 (IBU), 5 (BPA)
and 6 (CBZ). These results are quantified in terms of regional
normalized fluorescence quenching percentage (Pin), which is
plotted for all DOM-PPCP combinations (Fig. 7). Fluorescence
quenching of WWOM by addition of IBU was observed in
Fig. 4 e DOM quenching by ibuprofen represented by difference EEMs plotted with increasing PPCP concentration for
WWOM (A-10, B-20, C-200 and D-1000 mg LL1) and for SROM (E-10, F-20, G-200 and H-1000 mg LL1). All treatments conducted
in 24 mM NH4HCO3, pH 7.4, and 8 mg LL1 of DOC.
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4 949
region III, and to a lesser extent in regions I and II at all
concentrations (Figs. 4 and 7A). Similarly, IBU quenching of
SROMfluorescence occurred primarily in region III followed by
region I (Figs. 4 and 7D). Fluorescence quenching of WWOM in
the presence of BPAwasmanifested more broadly throughout
the EEM, and clearly within regions IV and V, attributed to
Fig. 5 e DOM quenching by bisphenol-A represented by differen
WWOM (A-100, B-200 and C-1000 mg LL1) and for SROM (D-100, E
NH4HCO3, pH 7.4, and 8 mg LL1 of DOC.
humic substances (Figs. 5 and 7B). Reaction of BPA with SROM
resulted in quenching of region III fluorophores with an
increasing quenching trend from low to high BPA concentra-
tion (Fig. 5). Although region V showed significant quenching
at low BPA concentration, the relative quenching effect in this
region decreased with increasing BPA concentration as region
ce EEMs plotted with increasing PPCP concentration for
-200 and F-1000 mg LL1). All treatments conducted in 24 mM
Fig. 6 e DOM quenching by carbamazepine represented by difference EEMs plotted with increasing PPCP concentration for
WWOM (A-10, B-20, C-200 and D-1000 mg LL1) and for SROM (E-10, F-20, G-200 and H-1000 mg LL1). All treatments conducted
in 24 mM NH4HCO3, pH 7.4, and 8 mg LL1 of DOC.
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4950
III became increasingly important (Fig. 7E). Quenching in
regions I, II and IV was negligible for all three BPA
concentrations.
Quenching of WWOM fluorescence by CBZ occurred
entirely in regions I, II and III at all tested concentrations (Figs.
6 and 7C). However the quenching of SROM fluorescence by
CBZwas negligible for regions I and II, low for regions III and IV
and consistently high for region V at all concentrations (Fig. 7C
and F). The trends discussed above were highly reproducible
as indicated by replicated experiments.
3.3. LC-MS/MS recovery
The mass balance recoveries of DOM-reacted PPCPs were
evaluated by direct LC-MS/MS injection of DOM-PPCP aqueous
solutions following incubation. Recovery statistics were
analyzed by ANOVA (a¼ 0.05) to determine the significance of
DOM effects relative to DOM-free positive controls (Table 3).
A negative control confirmed that no PPCP signal was detected
from the DOM material alone. The only case where DOM
treatment deviated from the positive control was in the case
of SROM-IBU, where enhancement (109% recovery) was
observed for the 10 mg L�1and for the WWOM-IBU at 1000
concentration. In all other cases, no statistically significant
difference was observed for analyte recovery in the presence
or absence of DOM (Table 3).
4. Discussion
4.1. Comparison of WWOM and SROM
Since traditional wastewater treatments are only partially
effective at removing PPCPs (Snyder et al., 2007), these
contaminants are introduced to environmental water
supplies following discharge of treated wastewater into
surface or ground water systems. Hence, the interactions of
these variably charged and often polar organic compounds
with both wastewater and fresh water DOM constituents are
relevant to their transport and fate (Maoz and Chefetz, 2010;
Chefetz et al., 2008). Wastewater and freshwater DOM sour-
ces are expected to vary significantly in their chemical prop-
erties. Wastewater DOM is dominantly sourced from
residential and industrial waste materials and microbial
biomass associated with biological treatment processes
(Chefetz et al., 2006). Conversely, freshwater DOM is derived
from decay of terrestrial and/or aquatic biomass, with the
relative predominance of plant or algal contributions being
dependent on location and season (McKnight et al., 2001).
These chemical differences between wastewater and
freshwater OM are apparent for the samples used in the
present study. Their distinct EEMs indicate lower contribu-
tions of humic acid-like e and greater contributions of
microbial byproduct-like and proteinaceous constituents e to
fluorescence ofWWOMrelative to SROM (Fig. 1D). TheWWOM
sample also exhibited low apparent molar mass and high
polydispersity (Fig. 3), low molar absorptivity (ε, Table 2) and
hence low aromaticity, and dominantly comprised function-
alities associated with microbial biomolecular fragments,
particularly those deriving from polysaccharides and proteins
observed in FTIR (Fig. 2). Conversely, the SROM exhibited
higher apparent molar mass and lower polydispersity (Fig. 3),
more than 10-fold higher molar absorptivity (ε, Table 2), and
prevalent polyphenolic aromatic moieties characteristic of
partially-degraded lignin originating from vegetation decay
(Fig. 2) (Leenheer and Croue, 2003; Chen et al., 2003). Hence,
WWOM and SROM are expected to exhibit different reactivity
toward PPCPs, with other experimental conditions (DOC
concentration, pH, background electrolyte concentration)
held constant.
10 20 200 10000
25
50
75
100
SROM
Ibuprofen ( g L-1)
Pin
A D
E
F
WWOM
10 20 200 10000
25
50
75
100
Ibuprofen ( g L-1)
Pin
100 200 10000
25
50
75
100
BPA ( g L-1)
Pin
B
100 200 10000
25
50
75
100
BPA ( g L-1)
Pin
10 20 200 10000
25
50
75
100
Carbamazepine ( g L-1)
Pin
Region I Region II Region III
Region IV Region V
C
10 20 200 10000
25
50
75
100
Carbamazepine ( g L-1)
Pin
Fig. 7 e Regional normalized fluorescence quenching
percentage (Pin) of WWOM (AeC) and SROM (DeF) in
24 mM NH4HCO3 of pH 7.4 buffer and 8 mg LL1 of DOC.
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4 951
4.2. DOM-PPCP interactions
The fluorescence quenching of DOM as reflected in an EEM is
evidently sensitive to DOM-PPCP interaction under
environmentally-relevant conditions (Figs. 4e6). Although the
method cannot provide a quantitative assessment of the DOM
bound fraction of PPCP, it does provide “fingerprints” of
reactive DOMfluorophores. A consistent result with respect to
the quenching EEMs is that quenching patterns are qualita-
tively consistent for a given DOM-PPCP pair, whereas signifi-
cant differences are observed between DOM types for a given
PPCP (compare top and bottom rows in Figs. 4e6), and
different PPCPs exhibit distinctly different quenching
patterns. Conversely, within a DOM-PPCP pair, the effects of
PPCP concentration appear to be relatively small.
This qualitative assessment is supported, in part, by
quantitative integration of the EEM “quenching matrix”,
which provides an index for direct comparison among
different PPCPs and DOM types (Fig. 7). These data further
demonstrate the strong dependence of quenched regions of
the EEM on both compound and DOM type, and quenching
trends with PPCP concentration also become evident when
viewed this way.The present experiments probed DOM fluorescence
quenching by three PPCP analytes that were anionic (IBU), and
neutral (BPA and CBZ) at the experimental pH of ca. 7.4.
Ibuprofen, bearing a negatively-charged carboxyl group
attached to a neutral backbone (Table 1), yielded reproducible
quenching of regions III (fulvic acid-like), II and I (aromatic
protein-like) in bothWWOM and SROM.We postulate that the
dominance of quenching of fulvic acid-like moieties may be
the result of cation bridging between deprotonated carboxyls
of fulvate and ibuprofen, irrespective of DOM type. Indeed,
density functional theory calculations indicate that cation
bridging between carboxylated contaminants and ionized,
carboxylated DOM is energetically favorable at circumneutral
pH (Aquino et al., 2008).
In contrast, the neutral BPA molecule showed distinctly
different quenching patterns between WWOM and SROM.
It quenched fluorescence of humicacid-like like constituents
in both cases, but also quenched significantly microbial
byproduct-like fluorophores in WWOM (where such compo-
nents are more prevalent). Also, trends with BPA concentra-
tion were different between the two OM types: increasing BPA
concentration in WWOM solutions gave rise to a relative
increase in quenching of humic acid-like components,
whereas in SROM solutions, quenching increasingly affected
fulvic acid-like fluorophores. We speculate that these trends
signal a change in dominant interaction mechanism as well;
hydrophobic association is expected to be more important for
aromatic moieties characteristic of region V fluorophores,
whereas hydrophilic interaction (i.e., water bridging and/or
hydrogen bonding)would bemore likely to occurwith oxygen-
containing, polar functionalities characteristic of region III
fluorophores. The shift from one fluorophore group to another
not only provides an indication of favorable interactions, but
also offers insight to failed attempts to apply log Kow to
explain DOM-PPCP interactions. Predictions based on log Kow
assume that hydrophobic interactions are paramount,
whereas polar functional groups may in fact mediate
association between ionized chemicals and oxygen-
containing DOM moieties (Sangster, 1997; Tolls, 2001; Pan
et al., 2009).
Carbamazepine quenched distinctly different fluorophores
in the two DOM types: fulvic acid-like constituents in WWOM
and humic acid-like constituents in SROM. In both cases,
hydrogen bonding, pep and van der Waals interactions likely
prevailed (Navon et al., 2011), the difference in quenching
region being largely due to the difference in prevalence of
polar functionalities of fulvate versus humate fluorophores
between the two DOM sources (Fig. 1D). Carbamazepine
interactions with humic acid-like regions of DOM have been
reported in a prior fluorescence quenching experiment that
used a landfill leachate fulvic acid (Bai et al., 2008). However
that study did not measure the low excitation portion
(200e250 nm) of the EEM (regions I, II and III), which precluded
assessment of quenching in these regions.
Much prior research has suggested that the partitioning of
PPCP contaminants to dissolved organic matter may be
based largely on hydrophobic interactions (Cabaniss et al.,
2000; Pan et al., 2009). However, our limited understanding
is based on the prevalent use of terrestrial DOM sources in
such studies. As shown in this study, terrestrial DOM is
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4952
indeed more aromatic as a result of lignin-based precursors
than is DOM formed as a result of microbial processing that
occurs during wastewater treatment (Chen et al., 2003;
Leenheer and Croue, 2003; Leenheer, 2004). However, even
for such terrestrial sources, assessment of DOM-PPCP affinity
on the basis of Kow value does not account for polar inter-
actions that may be important to intermolecular association
(Tolls, 2001).
4.3. LC-MS/MS recovery
Tandemmass spectrometry has been previously used to aid in
understanding of complex formation in biological molecules,
and we assert that such an approach is profitably transferable
to PPCPs as well. Prior work has focused on non-covalent
bonding between known proteins and ligands wherein
complexation reduces the signal intensity detected by the LC-
MS/MS compared to that of a control investigated under the
same solution chemistry (Bolbach, 2005; Loo, 2000; Daniel
et al., 2002). Such a signal decrease can occur due to interac-
tion forces that prevent chromatographic or gas phase sepa-
ration of the complex. The strength and stability of
interactions can vary. For example, as shown by Loo (2000),
protein subunits can be associated via hydrophobic interac-
tion or hydrogen bonding, both of which are more fragile and
thus labile in the gas phase relative to electrostatic attractions
such as those between a cationic and anionic organic species.
Thus a reduction in LC-MS/MS signal compared to that of the
control signals potential strong ionic and/or covalent
interactions.
Table 3 e LC-MSMS Mean percent recovery and standarddeviations of spiked IBU, BPA and CBZ in WWOM andSROM solutions (DOC[ 8 mg LL1) and positive controls(DC) from 10e1000 mg LL1 in 24 mM NH4HCO3 electrolyteat pH 7.4.
PPCP (mg L�1) WWOM mean% recovery
SROM mean% recovery
IBU 10 95� 6 109� 8
þC 10 97� 3 97� 3
20 102� 2 107� 3
þC 20 99� 3 99� 3
200 98� 0 97� 1
þC 200 100� 2 100� 2
1000 93� 2 96� 2
þC 1000 100� 3 104� 6
BPA 200 101� 11 107� 5
þ C 200 96� 14 96� 14
1000 93� 9 86� 11
þC 1000 95� 5 95� 10
CBZ 10 103� 22 103� 3
þC 10 100� 5 100� 5
20 105� 18 100� 2
þC 20 100� 1 100� 1
200 106� 5 110� 2
þC 200 96� 4 96� 4
1000 108� 13 96� 3
þC 1000 101� 3 106� 10
The use of aqueous injection LC-MS/MS to measure PPCP
recovery in the presence of the two DOM sources with
different physico-chemical properties indicates that the
interactions giving rise to fluorescence quenching were not
strong enough to prevent chromatographic separation,
ionization, fragmentation and detection of the target analy-
tes. This suggests that the presence of DOM, at
environmentally-relevant DOC concentrations (8 mg L�1)
should not interfere with the detection and quantification of
these trace contaminant concentrations by direct aqueous
injection.
5. Conclusion
At low PPCP concentrations, as commonly occurs in natural
environments, BPA interacts with soluble wastewater DOM
components including microbially-derived biomolecular
fragments, while IBU and CBZ associate with fulvic acid-like
fluorophores. Interactions with low molar mass, soluble
DOM components likely facilitate the transport of PPCPs
from waste to freshwater environments (Cabaniss et al.,
2000). Convergence of wastewater effluents with freshwater
sources could likely result in transfer of the same contami-
nants to humic acid-like DOM components that are typically
of higher molar mass and greater aromaticity. For both
WWOM and SROM, interactions with the three PPCPs were
sufficiently weak to permit ca. 100% recovery with aqueous
injection LC-MS/MS, irrespective of analyte structure and
DOM source. Additionally, direct injection LC-MS/MS studies
indicate that the presence of DOM at concentrations
employed does not suppress the accurate assessment of
these target analytes.
Acknowledgments
Research support was provided by the Binational Agricultural
Research and Development (BARD) fund, Grant # IS-3822-06,
and Water Research Foundation (Award #4269) and Univer-
sity of Arizona Water Sustainability Program. The comments
and views detailed herein may be necessarily reflecting the
views of theWater Research Foundation, its officers, directors,
affiliates, or agents. Analyses in the Arizona Laboratory for
Emerging Contaminants were supported by NSF CBET
0722579.
Author contributions
Selene Hernandez Ruiz: Experimental design, execution, data
processing and interpretation, manuscript writing.
Leif Abrell: Technical: LC-MS/MS support in method
development.
Samanthi Wickarasemara: Fine tuning of LC-MS/MS
instrument.
Benny Chefetz: Experimental design planning, data interpre-
tation support, manuscript writing.
Jon Chorover: Experimental design planning, data interpre-
tation support, manuscript writing.
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4 953
r e f e r e n c e s
Aquino, A.J.A., Tunega, D., Pa�sali�c, H., Haberhauer, G.,Gerzabek, M.H., Lischka, H., 2008. The thermodynamicstability of hydrogen bonded and cation bridged complexes ofhumic acid modelsda theoretical study. Chem. Phys. 349(1e3), 69e76.
Bai, Y., Wu, F., Liu, C., Guo, J., Fu, P., Li, W., Xing, B., 2008.Interaction between carbamazepine and humic substances:a fluorescence spectroscopy study. Environ. Toxicol. Chem. 27(1), 95e102.
Bolbach, G., 2005. Matrix-assisted laser desorption/ionizationanalysis of non-covalent complexes: fundamentals andapplications. Curr. Pharm. Des. 11 (20), 2535e2557.
Cabaniss, S.E., Zhou, Q.H., Maurice, P.A., Chin, Y.P., Aiken, G.R.,2000. A log-normal distribution model for the molecularweight of aquatic fulvic acids. Environ. Sci. Technol. 34 (6),1103e1109.
Cahill, J.D., Furlong, E.T., Burkhardt, M.R., Kolpin, D.,Anderson, L.G., 2004. Determination of pharmaceuticalcompounds in surface- and ground-water samples by solid-phase extraction and high-performance liquidchromatography-electrospray ionization mass spectrometry.J. Chromatogr. A 1041 (1e2), 171e180.
Capdeville, M.J., Budzinski, H., 2011. Trace-level analysis oforganic contaminants in drinking waters and groundwaters.Trac-Trends Analy. Chem. 30 (4), 586e606.
Chefetz, B., Illani, T., Schulz, E., Chorover, J., 2006. Wastewaterdissolved organic matter: characteristics and sorptivecapabilities. Water Sci. Technol. 53 (7), 51e57.
Chefetz, B., Hadar, Y., Chen, Y., 1998. Dissolved organic carbonfractions formed during composting of municipal solid waste:properties and significance. Acta Hydrochim. Hydrobiol. 26 (3),172e179.
Chefetz, B., Mualem, T., Ben-Ari, J., 2008. Sorption and mobility ofpharmaceutical compounds in soil irrigated with reclaimedwastewater. Chemosphere 73 (8), 1335e1343.
Chen, W., Westerhoff, P., Leenheer, J.A., Booksh, K., 2003.Fluorescence excitation e emission matrix regionalintegration to quantify spectra for dissolved organic matter.Environ. Sci. Technol. 37 (24), 5701e5710.
Chin, Y.P., Aiken, G., Oloughlin, E., 1994. Molecular-weight,polydispersity, and spectroscopic properties of aquatic humicsubstances. Environ. Sci. Technol. 28 (11), 1853e1858.
Chorover, J., Amistadi, M.K., Chadwick, O.A., 2004. Surface chargeevolution of mineral-organic complexes during pedogenesisin Hawaiian basalt. Geochim. Cosmochim. Acta 68 (23),4859e4876.
Daniel, J.M., Friess, S.D., Rajagopalan, S., Wendt, S., Zenobi, R.,2002. Quantitative determination of noncovalent bindinginteractions using soft ionization mass spectrometry. Int. J.Mass Spectrom. 216 (1), 1e27.
Dickenson, E.R.V., Snyder, S.A., Sedlak, D.L., Drewes, J.E., 2011.Indicator compounds for assessment of wastewater effluentcontributions to flow and water quality. Water Res. 45 (3),1199e1212.
Focazio, M.J., Kolpin, D.W., Barnes, K.K., Furlong, E.T.,Meyer, M.T., Zaugg, S.D., Barber, L.B., Thurman, M.E., 2008.A national reconnaissance for pharmaceuticals and otherorganic wastewater contaminants in the United States e II)Untreated drinking water sources. Sci. Total Environ. 402(2e3), 201e216.
Gomez, M.J., Petrovic, M., Fernandez-Alba, A.R., Barcelo, D., 2006.Determination of pharmaceuticals of various therapeuticclasses by solid-phase extraction and liquid chromatography-tandem mass spectrometry analysis in hospital effluentwastewaters. J. Chromatogr. A 1114 (2), 224e233.
Henderson, R.K., Baker, A., Murphy, K.R., Hamblya, A.,Stuetz, R.M., Khan, S.J., 2009. Fluorescence as a potentialmonitoring tool for recycled water systems: a review. WaterRes. 43 (4), 863e881.
Her, N., Amy, G., McKnight, D., Sohn, J., Yoon, Y.M., 2003.Characterization of DOM as a function of MW by fluorescenceEEM and HPLC-SEC using UVA, DOC, and fluorescencedetection. Water Res. 37 (17), 4295e4303.
Hudson, N., Baker, A., Reynolds, D., 2007. Fluorescence analysis ofdissolved organic matter in natural, waste and pollutedwaters e a review. River Res. Appl. 23 (6), 631e649.
Ilani, T., Schulz, E., Chefetz, B., 2005. Interactions of organiccompounds with wastewater dissolved organic matter: role ofhydrophohic fractions. J. Environ. Qual. 34 (2), 552e562.
Jjemba, P.K., 2006. Excretion and ecotoxicity of pharmaceuticaland personal care products in the environment. Ecotoxicol.Environ. Saf. 63 (1), 113e130.
Joss, A., Siegrist, H., Ternes, T.A., 2008. Are we about to upgradewastewater treatment for removing organic micropollutants?Water Sci. Technol. 57 (2), 251e255.
Kagle, J., Porter, A.W., Murdoch, R.W., Rivera-Cancel, G., Hay, A.G.,2009. Biodegradation of pharmaceutical and personal careproducts. Adv. Appl. Microbiol. 67, 65e108.
Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M.,Zaugg, S.D., Barber, L.B., Buxton, H.T., 2002. Pharmaceuticals,hormones, and other organic wastewater contaminants in USstreams, 1999e2000: a national reconnaissance. Environ. Sci.Technol. 36 (6), 1202e1211.
Kuster, M., Jose Lopez de Alda, M., Dolores Hernando, M.,Petrovic, M., Martin-Alonso, J., Barcelo, D., 2008. Analysis andoccurrence of pharmaceuticals, estrogens, progestogens andpolar pesticides in sewage treatment plant effluents, riverwater and drinking water in the Llobregat river basin(Barcelona, Spain). J. Hydrol. 358 (1e2), 112e123.
Kwon, J., Armbrust, K.L., 2008. Aqueous solubility, n-octanol-water partition coefficient, and sorption of five selectiveserotonin reuptake inhibitors to sediments and soils. Bull.Environ. Contam. Toxicol. 81 (2), 128e135.
Lajeunesse, A., Gagnon, C., 2007. Determination of acidicpharmaceutical products and carbamazepine in roughlyprimary-treated wastewater by solid-phase extraction and gaschromatography-tandem mass spectrometry. Int. J. Environ.Anal. Chem. 87 (8), 565e578.
Lakowicz, J.R., 1999. Principles of Fluorescence Spectroscopy.Kluwer, New York.
Lee, K.E., Barber, L.B., Furlong, E.T., Cahill, J.D., Kolpin, D.W.,Meyer, M.T., Zaugg, S.D., 2000e02. Presence and distribution oforganic wastewater compounds in wastewater, surface,ground, and drinking waters, Minnesota, 2000e02. U.S.Geological Survey Scientific Investigations Report 2004-5138.
Leenheer, J.A., 1981. Comprehensive approach to preparativeisolation and fractionation of dissolved organic-carbon fromnatural-waters and wastewaters. Environ. Sci. Technol. 15 (5),578e587.
Leenheer, J.A., 2004. Comprehensive assessment of precursors,diagenesis, and reactivity to water treatment of dissolved andcolloidal organic matter. Water Sci. Technol. Water Supply 4(4), 1e9.
Leenheer, J.A., Croue, J.P., 2003. Characterizing aquatic dissolvedorganic matter. Environ. Sci. Technol. 37 (1), 18Ae26A.
Li, F.S., Yuasa, A., Matsui, Y., Cheong, E.J., 2005. Polydisperseadsorption characteristics of aqueous organic matrices inwater and wastewater sources. Adsorp. J. Int. Adsorp. Soc. 11,691e696.
Liu, R.X., Wilding, A., Hibberd, A., Zhou, J.L., 2005. Partition ofendocrine-disrupting chemicals between colloids anddissolved phase as determined by cross-flow ultrafiltration.Environ. Sci. Technol. 39 (8), 2753e2761.
wat e r r e s e a r c h 4 6 ( 2 0 1 2 ) 9 4 3e9 5 4954
Loo, J.A., 2000. Electrospray ionization mass spectrometry:a technology for studying noncovalent macromolecularcomplexes. Int. J. Mass Spectrom. 200 (1e3), 175e186.
Maoz, A., Chefetz, B., 2010. Sorption of the pharmaceuticalscarbamazepine and naproxen to dissolved organic matter:role of structural fractions. Water Res. 44 (3), 981e989.
McKnight, D., Boyer, E., Westerhoff, P., Doran, P., Kulbe, T.,Andersen, D., 2001. Spectrofluorometric characterization ofdissolved organic matter for indication of precursor organicmaterial and aromaticity. Limnol. Oceanogr. 46 (1), 38e48.
Meulenberg, E.P., Peelen, G.O.H., Lukkien, E., Koopal, K., 2005.Immunochemical detection methods for bioactive pollutants.Int. J. Environ. Anal. Chem. 85 (12e13), 861e870.
Navon, R., Hernandez-Ruiz, S., Chorover, J., Chefetz, B., 2011.Interactions of carbamazepine in soil: effects of dissolvedorganic matter. J. Environ. Qual. 40 (3), 942e948.
Nghiem, L., Schafer, A., Elimelech, M., 2005. Pharmaceuticalretention mechanisms by nanofiltration membranes. Environ.Sci. Technol. 39 (19), 7698e7705.
Nghiem, L.D., Hawkes, S., 2009. Effects of membrane fouling onthe nanofiltration of trace organic contaminants. Desalination236 (1e3), 273e281.
Omoike, A., Chorover, J., 2006. Adsorption to goethite ofextracellular polymeric substances from Bacillus subtilis.Geochim. Cosmochim. Acta 70 (4), 827e838.
Pan, B., Ning, P., Xing, B., 2009. Part V-sorption of pharmaceuticalsand personal care products. Environ. Sci. Pollut. Res. 16 (1),106e116.
Pedrouzo, M., Reverte, S., Borrull, F., Pocurull, E., Marce, R.M.,2007. Pharmaceutical determination in surface andwastewaters using high-performance liquid chromatography-(electrospray)-mass spectrometry. J. Separat. Sci. 30 (3),297e303.
Polubesova, T., Sherman-Nakache, M., Chefetz, B., 2007. Bindingof pyrene to hydrophobic fractions of dissolved organicmatter: effect of polyvalent metal complexation. Environ. Sci.Technol. 41 (15), 5389e5394.
Sangster, J., 1997. OctanoleWater Partition Coefficients:Fundamentals and Physical Chemistry. Wiley, Chichester;New York.
Santamaria, A., Mondragon, F., Molina, A., Marsh, N.D.,Eddings, E.G., Sarofim, A.F., 2006. FT-IR and H-1 NMRcharacterization of the products of an ethylene inversediffusion flame. Combust. Flame 146 (1e2), 52e62.
Serkiz, S.M., Perdue, E.M., 1990. Isolation of dissolved organicmatter from the suwannee river using reverse osmosis. WaterRes. 24 (7), 911e916.
Shenker, M., Harush, D., Ben-Ari, J., Chefetz, B., 2011. Uptake ofcarbamazepine by cucumber plants e A case study related toirrigation with reclaimed wastewater. Chemosphere 82 (6),905e910.
Snyder, S.A., Adham, S., Redding, A.M., Cannon, F.S., DeCarolis, J.,Oppenheimer, J., Wert, E.C., Yoon, Y., 2007. Role of
membranes and activated carbon in the removal of endocrinedisruptors and pharmaceuticals. Desalination 202 (1e3),156e181.
Soto, A.M., Sonnenschein, C., 2010. Environmental causes ofcancer: endocrine disruptors as carcinogens. Nat. Rev.Endocrinol. 6 (7), 364e371.
Stavrakakis, C., Colin, R., Hequet, V., Faur, C., Le Cloirec, P., 2008.Analysis of endocrine disrupting compounds in wastewaterand drinking water treatment plants at the nanogram per litrelevel. Environ. Technol. 29 (3), 279e286.
Stuart, B., Ando, D.J., 1996. Modern Infrared Spectroscopy.Published on behalf of ACOL (University of Greenwich) byWiley, New York,.
Sun, W.L., Ni, J.R., Xu, N., Sun, L.Y., 2007. Fluorescence ofsediment humic substance and its effect on the sorption ofselected endocrine disruptors. Chemosphere 66 (4), 700e707.
Swift, R.S., 1996. Organic matter characterization. In: Sparks, D.L.(Ed.), Methods of Soil Analysis. Part 3, Chemical Methods. SoilScience Society of America: American Society of Agronomy,Madison, Wis, pp. 1011e1070.
Tolls, J., 2001. Sorption of veterinary pharmaceuticals in soils:a review. Environ. Sci. Technol. 35 (17), 3397e3406.
Vom Saal, F.S., Cooke, P.S., Buchanan, D.L., Palanza, P.,Thayer, K.A., Nagel, S.C., Parmigiani, S., Welshons, W.V., 1998.A Physiologically based Approach to the Study of Bisphenol Aand Other Estrogenic Chemicals on the Size of ReproductiveOrgans, Daily Sperm Production, and Behavior. Toxicologyand Industrial Health, pp. 239e260.
Wickramasekara, S., Hernandez Ruiz, S., Abrell, L., Arnold, R.,Chorover, J. Natural dissolved organic matter affectselectrospray ionization during analysis of emergingcontaminants by mass spectrometry. Anal. Chim. Acta., inpress.
Wu, F.C., Evans, R.D., Dillon, P.J., 2003. Separation andcharacterization of NOM by high-performance liquidchromatography and on-line three-dimensional excitationemission matrix fluorescence detection. Environ. Sci. Technol.37 (16), 3687e3693.
Yamamoto, H., Liljestrand, H.M., Shimizu, Y., Morita, M., 2003.Effects of physicalechemical characteristics on the sorption ofselected endocrine disruptors by dissolved organic mattersurrogates. Environ. Sci. Technol. 37 (12), 2646e2657.
Zhang, Z.L., Zhou, J.L., 2007. Simultaneous determination ofvarious pharmaceutical compounds in water by solid-phaseextraction-liquid chromatography-tandem massspectrometry. J. Chromatogr. A 1154 (1e2), 205e213.
Zhou, J.L., Liu, R., Wilding, A., Hibberd, A., 2007. Sorption ofselected endocrine disrupting chemicals to different aquaticcolloids. Environ. Sci. Technol. 41 (1), 206e213.
Zhou, Q.H., Cabaniss, S.E., Maurice, P.A., 2000. Considerations inthe use of high-pressure size exclusion chromatography(HPSEC) for determining molecular weights of aquatic humicsubstances. Water Res. 34 (14), 3505e3514.