ENVIRONMENTAL SURVEILLANCE OF PHARMACEUTICALS AND SORPTION TO
CLAY MINERALS
By
YUNJIE DING

A DISSERTATION
Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of
DOCTOR OF PHILSOPHY
Crop and Soil Sciences
2011

ABSTRACT
ENVIRONMENTAL SURVEILLANCE OF PHARMACEUTICALS AND SORPTION TO
CLAY MINERALS
By
YUNJIE DING
Pharmaceuticals are commonly used to animals and humans for disease control and
treatment. Large fractions of administered pharmaceuticals are released into the environment via
land application of animal manures and biosolids from sewage sludge of wastewater treatment
plants (WWTPs). In this study, we investigated the occurrence and fate of amprolium, carbadox,
monensin and tylosin in an agricultural farm. Amprolium and monensin were frequently detected
in surface runoff with concentrations at nanograms per liter of water, whereas tylosin or
carbadox was rarely found. Soil is a reservoir for these veterinary pharmaceuticals. Accompanied
with this study, we developed analytical methods to determine pharmaceuticals in water and
biosolids. Among commonly used fifteen pharmaceuticals, fourteen pharmaceuticals were
frequently detected in the collected biosolids from WWTPs. This result suggests that
pharmaceuticals survive wastewater treatment process and are disseminated into ecosystems via
land application. Soil is a major media for retention of pharmaceuticals in the environment, and
soil mineral is a major component for sorption of pharmaceuticals. Tetracycline sorption
mechanism to clays from aqueous solution was investigated. Cation exchange reaction and
metal-tetracycline complexation were identified to be the major processes responsible for
tetracycline sorption on minerals. Overall, this study provides useful information to evaluate
occurrence and fate of pharmaceuticals in the environment.

ACKNOWLEDGEMENTS

I would like to present great gratitude to my committee members: Dr. Hui Li, Dr. Brian
Teppen, Dr. Stephen Boyd and Dr. Irene Xagoraraki for their insightful, helpful comments and
guidance. I especially want to thank my major advisor Dr. Li for his patience, encouragement
and professional advice throughout the pursuit of my Ph.D degree. This dissertation could not
have been completed without his support and help.
I also want to thank my colleagues: Dr. Wenlu Song, Cun Liu, Dr. Cheng Gu, Yingjie
Zhang, Dr. Wan-Ru Chen, Dr. Weihao Zhang, Dr. Cuiping Wang and Dr. Kelvin Wong for their
help and friendship.
It is gratefully acknowledged that United States Department of Agriculture provided
financial support by research competitive grant 2007-35107-18353 and 2009-65102-05847.
Finally, the most profound gratitude to my wife, Leilei Qian, for her love, support and
sacrifices during the past five years of study, and to my coming baby. I also acknowledge my
parents: Jinhua Ding and Rongying Mao, and parents-in-law: Zhong Qian and Sucheng Ge for
their great support and sacrifices.

iii 
 

TABLE OF CONTENTS

 
LIST OF TABLES ............................................................................................................. vi
LIST OF FIGURES .......................................................................................................... vii
CHAPTER I: SELECTED VETERINARY PHARMACEUTICALS IN AGRICULTURAL
WATER AND SOIL FROM LAND APPLICATION OF ANIMAL MANURE.............. 1
Abstract ........................................................................................................................... 2
Introduction ..................................................................................................................... 3
Materials and Methods .................................................................................................... 7
Results and Discussion .................................................................................................. 11
Conclusions ................................................................................................................... 25
Reference ....................................................................................................................... 27

CHPATER II: DETERMINATION OF PHARMACEUTICALS IN BIOSOLIDS USING
ACCERLATED SOLVEN EXTRACTION AND LIQUID CHROMATOGRAPHY/TANDEM
MASS SPECTROMETRY ............................................................................................... 30
Abstract ......................................................................................................................... 31
Introduction ................................................................................................................... 32
Experimental ................................................................................................................. 35
Results and discussion ................................................................................................... 41
Conclusions ................................................................................................................... 57
Reference ....................................................................................................................... 59

CHAPTER III: TETRACYCLINE SORPTION BY K- AND Ca-SMECTITE CLAYS FROM
WATER ............................................................................................................................ 64
Abstract ......................................................................................................................... 65
Introduction ................................................................................................................... 66
Experiment Section ....................................................................................................... 71
Results and Discussion .................................................................................................. 75
Reference ....................................................................................................................... 90

Chapter IV: TETRACYCLINE SORPTION TO DIFFERENT CLAY MINERALS FROM
WATER ............................................................................................................................ 95
Abstract ......................................................................................................................... 96
Introduction ................................................................................................................... 98
Materials and Methods ................................................................................................ 102
iv 
 

Results and Discussion ................................................................................................ 104
Conclusion ................................................................................................................... 114
Reference ..................................................................................................................... 117
 

v 
 

LIST OF TABLES

Table I-1. Selected properties of the soils used in measurement of sorption isotherms ... 11 
 

Table I-2. Sampling time, time since last manure application, precipitation during sampling
-1
intervals, and amprolium and monensin concentrations (ng L ) in surface water .......... 13
Table II-1. Precursor, product ions and mass spectrometer parameters used to identify and
quantify pharmaceuticals. ................................................................................................. 38
 

Table II-2. Extracted concentrations of pharmaceuticals vs. spiked concentrations from biosolid
by water-organic solvent mixtures using accelerating solvent extraction ........................ 43
Table II-3. Extraction efficiency of pharmaceuticals under varying accelerating solvent
extraction parameters ........................................................................................................ 45
2

Table II-4. Linear range and correlation coefficient (r ) of standard curves, method limit of
detection (LOD), limit of quantification (LOQ), and recovery. ....................................... 50
 

Table II-5. Recoveries of solid-phase extraction for pharmaceuticals spiked into biosolid extracts
after accelerating solvent extraction. ................................................................................ 53
Table II-6. Measured pharmaceutical concentrations and standard deviations (µg/kg dry matter)
in the collected biosolid samples. ..................................................................................... 56
 

Table IV-1. Selected properties of clay mineral used in this study. ............................... 103
 

Table IV-2. Langmuir equation fittings for tetracycline sorption onto clay minerals .... 108
 

Table IV-3. Cation exchange site and surface area of clay mineral occupied by tetracycline
......................................................................................................................................... 114

vi 
 

LIST OF FIGURES

Figure I-1. Molecular structures and selected properties of amprolium, carbadox, monensin, and
tylosin.................................................................................................................................. 6
Figure I-2. Distribution of measured concentration and detection frequencies of amprolium and
monensin in stagnant surface water and drainage water sampled from September 2006 to
September 2007 ................................................................................................................ 15
Figure I-3. Distribution of concentration of amprolium and monensin in soils sampled in May
and October 2006 .............................................................................................................. 20
Figure I-4. Sorption isotherms of (A) monensin and (B) amprolium obtained from laboratory
studies using soils adjacent to the study site which were devoid of target pharmaceuticals22
Figure III-1. Tetracycline structure and speciation distribution as a function of solution pH
........................................................................................................................................... 68
Figure III-2. Relationship between tetracycline sorption and exchanged cation released from clay
........................................................................................................................................... 76
Figure III-3. Tetracycline sorption isotherms for (A) K-smectite from KCl solution and (B) Casmectite from CaCl2 solution............................................................................................ 79
Figure III-4. Relationship between reduction of tetracycline concentration in aqueous phase due
to sorption by smectite and the consumption of proton in solution .................................. 81
Figure III-5. Molecular dynamic simulations of tetracycline sorption by Ca-smectite .... 85
Figure III-6. (A) X-ray diffraction patterns of K-smectite with increasing loading of tetracycline,
(B) increase of basal spacing as a function of tetracycline sorbed concentration ............ 87

vii 
 

Figure IV-1. (A) Tetracycline structure, and (B) speciation distribution of tetracycline as a
function of aqueous solution pH ..................................................................................... 101
+

Figure IV-2. Tetracycline sorption to K -saturated clay minerals ................................. 106
Figure IV-3. Tetracycline sorption to K-Panther Creek from aqueous solution containing 0.0,
0.001, 0.01 and 0.10 M KCl............................................................................................ 109
Figure IV-4. Tetracycline sorption to (A) K- and Ca-SHCa-1 and (B) K- and Ca-Panther Creek
clays ................................................................................................................................ 111

viii 
 

CHAPTER I

SELECTED VETERINARY PHARMACEUTICALS IN AGRICULTURAL WATER
AND SOIL FROM LAND APPLICATION OF ANIMAL MANURE

1 
 

Abstract
Veterinary pharmaceuticals are commonly administered to animals for disease control,
and added into feeds at subtherapeutic levels to improve feeding efficiency. As a result of these
practices, a certain fraction of the pharmaceuticals are excreted into animal manures. Land
application of these manures contaminates soils with the veterinary pharmaceuticals which can
subsequently lead to contamination of surface and ground waters. Information on the occurrence
and fate of pharmaceuticals in soil and water is needed to assess the potential for exposure of atrisk populations and the impacts on agricultural ecosystems. In this study, we investigated the
occurrence and fate of four commonly used veterinary pharmaceuticals (amprolium, carbadox,
monensin and tylosin) in a farm in Michigan. Amprolium and monensin were frequently detected
-1

in nearby surface water with concentrations ranging from several to hundreds of ngL , whereas
tylosin or carbadox was rarely found. These pharmaceuticals were more frequently detected in
surface runoff during non-growing season (October to April) than during growing season (May
to September). Pharmaceuticals resulting from post-harvest manure application appeared to be
more persistent than those from spring application. High concentrations of pharmaceuticals in
soils were generally observed at the sites where the respective concentrations in surface water
were also high. For monensin, the ratios of soil-sorbed to aqueous concentrations obtained from
field samples were within the order of the distribution coefficients obtained from laboratory
studies. These results suggest that soil is a reservoir for veterinary pharmaceuticals that can be
disseminated to nearby surface water via desorption from soil, surface runoff, and soil erosion.

2 
 

Introduction
Veterinary pharmaceuticals are commonly administered to animals for disease control,
and added into feeds at subtherapeutic levels to improve livestock growth via enhanced feeding
efficiency. Such pharmaceutical use facilitates the success of large-scale concentrated animal
feeding operations (CAFOs). It was estimated that in the United States, approximately 11.2
million kilograms of antibiotics were used as nontherapeutic feed additives, corresponding to
about 70% of the annual antibiotic production (Mellon et al., 2001). Large fractions (e.g., up to
95%) of animal feeding pharmaceuticals are excreted into manure either as parent compounds or
as bioactive metabolites (Aga et al., 2005; Jacobsen et al., 2004; Kay et al., 2004). After a period
of storage, the manure is usually land-applied for its fertilizer value. This practice introduces
manure-borne pharmaceuticals into agricultural ecosystems where their fate is largely unknown
(Halling-Sorensen et al., 2005; Hamscher et al., 2005; Kolpin et al., 2002; Snow et al., 2008).
There is growing concern regarding the environmental fate and potential impacts of these
veterinary pharmaceuticals on human and ecosystem health.
Compared to point sources such as wastewater treatment plants, non-point sources of
contaminant release such as those related to agricultural practice, e.g., surface runoff from
manure-treated fields, has received much less attention. Land application of animal manures can
lead to veterinary pharmaceutical contamination of surface water and groundwater via surface
runoff and leaching (Dolliver and Gupta, 2008; Kay et al., 2004; Kay et al., 2005; Kuchta et al.,
2009; Larsbo et al., 2009). Reliable and robust analytical methods have been developed to
measure the levels and frequencies of pharmaceuticals present in water and soil. This
information is needed to assess adverse effects of pharmaceuticals on environmental quality and
public health, and for the development of regulations to limit the release of these unregulated

3 
 

bioactive compounds. Solid-phase extraction (SPE) coupled with liquid chromatography/tandem
mass spectrometry (LC-MS/MS) is currently the most effective approach to identify and quantify
trace levels of pharmaceuticals in environmental samples. Using this capability, several classes
of veterinary pharmaceuticals, including tetracyclines, sulfonamides, macrolides, lincosamides,
and ionophores, have been detected in surface water adjacent to concentrated animal feeding
facilities, with the concentration ranging from low to hundreds of nanograms per liter (Brown et
al., 2006; Hamscher et al., 2005; Hao et al., 2006; Kolpin et al., 2002; Song et al., 2007). The
pharmaceuticals manifesting relatively weak sorptive interactions with soils, such as some of
sulfonamides, have been shown to leach through soil profile and into groundwater. For example,
-1

sulfonamides have been detected in groundwater at concentrations up to 215 ng L (Batt et al.,
2006).
Soil is the primary reservoir for many veterinary pharmaceuticals released from landapplied livestock manures (Jacobsen et al., 2004; Kim and Carlson, 2007; Stoob et al., 2007).
Pharmaceuticals commonly contain several polar and/or ionic functional groups that may engage
in multiple interactions with soil components, such as by ion exchange, metal-organic
complexation, and hydrogen bonding. Partitioning into soil organic matter (SOM) is, by itself,
insufficient to account for pharmaceutical retention by soils, as evidenced by the wide range (up
to 4 orders of magnitude) of SOM-normalized sorption data (e.g., Kom values) for individual
pharmaceutical among soils (Tolls, 2001). In comparison, other nonionic compounds with
similar lipophilic properties but lacking polar/ionic functionalities display a much narrower
range of Kom values (Tolls, 2001). The relatively strong affinity of certain pharmaceuticals with
soil reduces their availability for biotic or abiotic transformations, leading to enhanced
4 
 

-1

persistence in the environment. For example, Hamscher et al. (2002) reported up to 200 µg kg

of tetracycline in the soils fertilized with livestock liquid manures. In a two-year field study, they
found that the rate of tetracycline accumulation in soils exceeded the rate of degradation
suggesting that antibiotics might persist in soils over years with continuous applications of
animal manures (Hamscher et al., 2002). As a result, precipitation and irrigation can disseminate
manure-borne veterinary pharmaceuticals and their metabolites from soil to surface and ground
waters via surface runoff or leaching.
The objective of this study was to document the concentrations of selected veterinary
pharmaceuticals in agricultural soil and surface runoff in an agricultural farm near Lansing,
Michigan. Four veterinary pharmaceuticals, amprolium, carbadox, monensin, and tylosin were
selected as target compounds, and their levels in soil and water were quantified at various time
points for a period of one year. The selected pharmaceuticals had been used as livestock feed
additives at the study site (farm), and encompass a range of distinct structures and
physicochemical characteristics (Figure I-1). Amprolium is a cationic compound added to
livestock feeds to inhibit the growth of protozoan coccidian. Carbadox is commonly used to treat
dysenteries for swine and promote their growth. Monensin is a polyether ionophore widely used
in dairy cows and beef cattle to protect against certain types of gram-positive bacteria. The
macrolide tylosin is a broad-spectrum antibiotic commonly used to treat livestock infection and
promote livestock growth. The measured concentrations of these pharmaceuticals in soil and
water over the period of one year were used to evaluate their environmental fate in an
agricultural setting.

5 
 

              

 
 

monensin
molecular weight = 693 g/mol
pKa = 4.26
log Kow = -2.5

amprolium
molecular weight = 278 g/mol
log Kow = -2.5

  

carbadox
molecular weight = 262 g/mol
pKa = 1.8, 10.5; log Kow = -1.37

 

tylosin
molecular weight = 916 g/mol
pKa = 7.39, 13.06; log Kow = 3.5

Figure I-1. Molecular structures and selected properties of amprolium, carbadox, monensin, and tylosin
6 
 

                    

Materials and Methods
Chemicals. Amprolium (5-[(2-methylpyridin-1-ium-1-yl)methyl]-2-propyl-pyrimidin-4amine chloride), carbadox (methyl(2E)-2-[(1,4-dioxidoquinoxalin-2yl)methylene]hydrazinecarboxylate), monensin (4-[2-[5-ethyl-5-[5-[6-hydroxy-6(hydroxymethyl)-3,5-dimethyl-oxan-2-yl]-3-methyl-oxolan-2-yl]oxolan-2-yl]-9-hydroxy-2,8dimethyl-1,6-dioxaspiro[4.5]dec-7-yl]-3-methoxy-2-methyl-pentanoic acid), and tylosin
([(2R,3R,4E,6E,9R,11R,12S,13S,14R)-12-{[3,6-dideoxy-4-O-(2,6-dideoxy-3-C-methyl-α-Lribo-hexopyranosyl)-3-(dimethylamino)-β-D-glucopyranosyl]oxy}-2-ethyl-14-hydroxy-5,9,13trimethyl-8,16-dioxo-11-(2-oxoethyl)oxacyclohexadeca-4,6-dien-3-yl]methyl 6-deoxy-2,3-di-Omethyl-β-D-allopyranoside) were purchased from Sigma-Aldrich Chemical Company (St. Louis,
MO). Formic acid, ammonium acetate, and heptafluorobutyric acid (HFBA) were also obtained
from Sigma-Aldrich Chemical Company. Methanol was of HPLC-purity purchased from J. T.
Baker Chemical Company (Phillipsburg, NJ). Water was obtained from a Millipore Mill-Q
(Billerica, MA) water system then filtered through a 0.22 µm membrane.
Site Description. This study was conducted in a CAFO located near Lansing, Michigan.
The facility consists of swine, beef cattle, horse, sheep, dairy, and poultry farms. The farms
7

6

produce approximately 1.2 x 10 L of liquid and 9.5 x 10 kg of solid manures each year. The
animal manures produced are stored in the farm for a few months or composted, and then spread
over 650-acre agricultural lands adjoining these farms. Surface runoff from the agricultural farm
is released into the surrounding surface water. The animal manures are usually applied to the
lands twice (spring and fall seasons) per year. During our study period, animal manures were
applied in April, May, August, October and November 2006, and again in April and May 2007.
The more intensive application of manure to the fields occurred in spring season in which 4.2 x
7 
 

6

5

6

6

10 L of liquid and 7.5 x 10 kg of solid manures were applied in April and May 2006, and 4.1 x
10 L of liquid and 1.4 x 10 kg of solid manures were applied in April and May 2007. In
5

August 2006 a relatively small amount of liquid animal manure (1.6 x 10 L) was applied to the
cropland. The application during post-harvest season (October and November 2006) was 1.9 x
6

10 L of liquid manure.
Water Sampling and Preparation. Surface water samples were collected from the
agricultural farm once per month from September 2006 to September 2007. The samples were
taken at six locations from drainage tile channels exiting from the farm, and five additional
locations in ditches surrounding the cropland. These samples were collected within two days
after a moderate to heavy precipitation event (i.e., > 10 mm). Water samples were placed in 1-L
amber glass bottles, transferred to laboratory, passed through 0.4-μm glass fiber filter (Millipore
Corp., Billerica, MA) to remove suspended solid particles, and stored in a refrigerator at 4 ºC. A
PrepSep 12-port vacuum manifold (Fisher Scientific, Pittsburgh, PA) connected to a filtration
pump was used for solid-phase extraction. A hydrophilic-lipophilic balance (HLB) cartridge
(Waters Oasis, Milford, MA) (30μm, 6cc/200mg) was used for sample extraction, cleanup, and
preconcentration. The cartridge was conditioned with 6 mL of methanol and 6 mL of water prior
to use. Water sample (50 mL) was mixed with 0.2 mL of 1.0 M ammonium acetate, and passed
-1

through the preconditioned cartridge at a rate of 1.5 mL min . The cartridge was washed with 5
mL of water, vacuum-dried for 3 min, and eluted with 5.0 mL of methanol containing 0.5%
-1

formic acid at a rate of ~1 mL min . Pharmaceuticals in the eluate were analyzed using a LCMS/MS (described below). For samples containing low levels of the target compounds, the

8 
 

volume of eluate (5.0 mL) was reduced to 1.0 mL or less using a gentle N2 flow to concentrate
-1

the samples. The quantification limits of this method were 5, 15, 0.6, and 20 ng L for
amprolium, carbadox, monensin, and tylosin, respectively. The average extraction recoveries (%)
were 116 ± 20 for amprolium, 74 ± 2 for carbadox, 87 ± 12 for monensin, and 71 ± 7 for tylosin.
A detailed description of the method development is presented in the supporting information.
Soil Sampling and Preparation. Soil samples were collected from the top 10 cm of soil
profile in the farm, air dried in the dark, and manually homogenized using stainless steel spatulas.
Five to ten grams of soil samples were weighed into 25-mL Corex glass centrifuge tubes
(Thermo Scientific, Waltham, MA), mixed with 3 g of Ottawa sand (20-30 mesh) (Fisher
Scientific, Pittsburgh, PA), and extracted with 15 mL of 4/1 (v/v) methanol/0.3 M ammonium
acetate solution by vigorously shaking for 2 h at 200 rpm on a platform shaker. The tubes were
then centrifuged at 4300 g for 20 min, and the supernatants were transferred to 15-mL Pyrex
graduated glass vials. The volume of extractant was reduced to 3 mL using a nitrogen-stream
evaporator apparatus. Approximately 12 mL of deionized water was added to the graduated vial
to reach 15 mL, and the entire solution was then extracted via the SPE procedure described
above. The average overall recoveries (%) were 53 ± 3 for amprolium, 67 ± 2 for carbadox, 116
± 10 for monensin, and 80 ± 3 for tylosin. The quantification limits for the soil extraction were
-1

0.035, 0.06, 0.004, and 0.09 μg kg for amprolium, carbadox, monensin, and tylosin,
respectively.
LC-MS/MS Analysis. The LC-MS/MS system used in this study was a Shimadzu HPLC
(Columbia, MD) consisting of two LC-20AD pumps and an autosampler (SIL-20A) coupled
with an Applied Biosystems (Foster City, CA) Sciex 3200 triple quadrupole mass spectrometer.
9 
 

High-purity gases required for mass spectrometer were supplied by a gas generator (NM20ZA,
PEAK Scientific, Billerica, MA). The Turbo IonSpray source of mass spectrometer was operated
with electrospray ionization source in positive mode. The target pharmaceuticals were quantified
using multiple reaction monitoring mode. Three pairs of precursor/product ion transitions were
simultaneously monitored in order to unambiguously identify the target analytes in the
environmental samples (Song et al., 2007). In this study the precursor/product ion pairs (m/z)
selected were: 243/150, 243/122, 243/94 for amprolium, 263/231, 263/145, 263/130 for carbadox,
693/675, 693/479, 693/461 for monensin, and 916/174, 916/772, 916/145 for tylosin. The first
pair of transition in each set was used to quantify the analyte.
A binary gradient mobile phase for the liquid chromatography system was programmed
with a flow rate of 300 µL min-1. Phase A (95% water and 5% methanol) and phase B (95%
methanol and 5% water) both contained 2 mM ammonium acetate and 20 mM HFBA.
Separation of the four analytes was achieved using a Thermo Hypersil Gold column (Thermo
Scientific, Waltham, MA, 50 x 2.1 mm, particle size, 5 µm) with the following mobile phase
gradient program: 0 to 2.0 min, A:B = 95:5; 2.0 to 7.0 min, phase B linearly increased to 70%;
7.0 to 7.5 min, phase B ramped to 100% and held until 10.5 min; 10.5 to 11.0 min, phase B
reduced to 5% and equilibrated in the flow system until 12.0 min. The injection volume was 10
μL. After each sample injection, the first five-minute elution volume was diverted to waste
through a built-in valve, then switched to the inlet of the mass spectrometer.
Sorption Isotherm Measurement. Sorption isotherms of monensin and amprolium were
measured using a batch equilibration method for four soils collected from the sites nearby the
farm which contained no detectable target pharmaceuticals. Selected soil properties are
-1

summarized in Table I-1. A series of initial monensin (50 to 1000 μg L ) and amprolium (100 to
10 
 

-1

2000 μg L ) solutions were prepared, then mixed with a known weight of soil placed in glass
centrifuge tubes. The mixtures were equilibrated in an end-over-end shaker at 40 rpm for 24 h at
room temperature (~23 ºC). Preliminary studies indicated that sorption approached equilibrium
within 24 h. The tubes were then centrifuged at 4000 g for 30 min. An aliquot of supernatant was
sampled and analyzed using LC-MS/MS. Experimental controls consisted of each
pharmaceutical initial solution without soil. No changes of pharmaceutical concentrations were
detected in the controls. Therefore, sorbed concentrations were calculated from the difference
between the initial and equilibrium concentrations of pharmaceuticals in the aqueous phase.
Table I-1. Selected properties of the soils used in measurement of sorption isotherms

Soil

pH

Sand (%)

Silt (%)

Clay (%)

SOM (%)

CEC
(cmol(+)/kg)

Capac A

7.3

55

24

21

5.5

15.8

Capac B

7.6

60

20

20

1.1

10.0

Colwood

6.9

64

21

15

6.3

15.3

Oshtemo

4.7

70

12

18

1.4

6.6

Results and Discussion
Water samples were collected monthly at six sites from drainage tile channels (drainage
water) exiting the farm, and at five sites from the ditches (stagnant water) in close proximity to
the farmland site during a one-year period (September 2006 to September 2007). Due to the
snow in January and February 2007, we could not collect surface water samples. In September
and December 2006 drainage water samples were not collected because of insufficient flow in
the drainage channels. The total number of samples collected was 109 including 55 stagnant
water samples and 54 drainage water samples (Table I-2). Among the four target
11 
 

pharmaceuticals, amprolium and monensin were most frequently detected; the concentrations
(without correction by extraction recoveries) are reported in Table I-2, along with the days since
last manure application and precipitation during sampling intervals. The concentration ranges
and the detection frequencies are presented in Figure I-2. For the stagnant water samples
collected 25 out of 55 samples contained amprolium above the detectable level, with
-1

concentrations ranging from 12 to 288 ng L , and 30 samples contained monensin at
-1

concentrations ranging from 1 to 189 ng L . For the 54 drainage water samples, amprolium was
-1

detected in 27 samples with concentrations between 6 and 53 ng L , and monensin was detected
-1

in 15 samples at concentrations ranging from 2 to 75 ng L . A similar concentration range for
-1

monensin (e.g., 20-220 ng L ) was found in agricultural runoff in the Grand River watershed,
Ontario, Canada (Lissemore et al., 2006).

12 
 

 

Table I-2. Sampling time, time since last manure application, precipitation during sampling intervals, and amprolium and monensin
-1
concentrations (ng L ) in surface water
Sampling time

Days since
last manure
application
15

Precipitation
(mm)

a

Analytes
Site 1

Drainage water
Site 2 Site 3 Site 4

Site 5

Site 6

c
------Monensin
------October 14, 2006
44
61.2
Amprolium
24
21
13
10
39
10
Monensin
nd
2
nd
nd
nd
3
November 12, 2006
11
60.2
Amprolium
22
16
53
nd
25
11
Monensin
nd
12
nd
nd
nd
nd
-----December 15, 2006
44
85.8
Amprolium
-Monensin
------March 24, 2007
143
111.0
Amprolium
nd
nd
nd
nd
6
nd
Monensin
nd
8
nd
nd
nd
nd
April 27, 2007
3
79.0
Amprolium
53
37
17
34
21
33
Monensin
15
11
6
4
5
8
May 17, 2007
9
67.1
Amprolium
nd
nd
nd
nd
nd
nd
Monensin
nd
5
nd
nd
nd
nd
June 20, 2007
37
122.9
Amprolium
nd
nd
nd
nd
nd
nd
Monensin
nd
4
nd
nd
nd
nd
July 24, 2007
63
9.9
Amprolium
34
18
nd
nd
nd
27
Monensin
nd
75
nd
nd
nd
nd
August 26, 2007
96
107.7
Amprolium
36
28
21
nd
nd
12
Monensin
nd
nd
18
nd
nd
nd
September 27, 2007
128
87.9
Amprolium
41
nd
38
nd
nd
nd
Monensin
nd
17
nd
nd
nd
nd
a
Drainage water samples were collected from the drainage channels exiting the farm; b Stagnant surface water samples were collected
from the ditches surrounding the farmlands; c not sampled due to insufficient flow in drainage channels; d not detected

September 15, 2006

48.0

Amprolium

13 
 

 

Table I-2 (cont’d)
Sampling time
September 15, 2006

Days since
last manure
application
15

Precipitation
(mm)
48.0

October 14, 2006

44

61.2

November 12, 2006

11

60.2

December 15, 2006

44

85.8

March 24, 2007

143

111.0

April 27, 2007

3

79.0

May 17, 2007

9

67.1

June 20, 2007

37

122.9

July 24, 2007

63

9.9

August 26, 2007

96

107.7

September 27, 2007

128

87.9

Site 7
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin
Amprolium
Monensin

14 
 

b

Analytes
d

nd
11
41
2
17
1
19
2
nd
nd
64
29
nd
nd
nd
nd
49
nd
nd
nd
nd
nd

Stagnant water
Site 8 Site 9 Site 10

Site 11

96

288

26

nd

10
82
4
30
nd
34
3
nd
nd
57
18
nd
nd
nd
nd
12
nd
nd
nd
nd
5

29
184
9
19
15
25
33
nd
94
260
189
nd
20
nd
13
282
18
nd
13
nd
8

nd
95
3
nd
nd
93
1
nd
1
33
10
nd
nd
nd
nd
71
nd
nd
nd
nd
8

nd
nd
5
18
2
23
2
nd
nd
17
5
nd
nd
nd
nd
nd
nd
nd
nd
nd
nd

 

350

300

S tagnant W a ter (55)

90th

Concentration (ng/L)

250

S tagn ant W ater (55 )
200

150

100

D ra in age W ater (5 4)
75th

D rainage W a ter (54)

50
M ed ia n
25th
1 0th
0
45%

50%

5 5%

28%

M o nensin

A m prolium

D ete ctio n F re q ue n cy o f T o ta l S a m p le s (% )

Figure I-2. Distribution of measured concentration and detection frequencies of amprolium and
monensin in stagnant surface water and drainage water sampled from September 2006 to
th

th

th

September 2007. The box-and-whisker plot represents the range, and the 10 , 25 , median, 75 ,
th

and 90 percentile of data distribution

15 
 

 

Concentrations of both monensin and amprolium in surface water fell within a relatively
narrow range for drainage water samples as compared to those in stagnant water (Figure I-2).
This could plausibly be attributed to mixing during the process of surface runoff. Surface runoff
from different locations of the farm converged and flowed into drainage channels. It is proposed
that this mixing process reduced the variation of pharmaceutical concentrations in drainage water.
The veterinary pharmaceuticals present in the stagnant water could result from surface runoff
that bypassed the drainage channels, overflow from the nearby livestock waste storage lagoons,
and/or desorption from soils contaminated by the pharmaceuticals.
Carbadox and tylosin were not found above detection limits in any of the drainage water
samples, but were occasionally detected in stagnant water. Carbadox was detected in three
-1

samples of stagnant water at concentrations of 20, 30 and 250 ng L , and tylosin was detected in
-1

two samples at 100 and 190 ng L , respectively. The relatively low occurrence of tylosin and
carbadox in this study is consistent with previous laboratory and field studies (Christian et al.,
2003; Kay et al., 2004; Kolpin et al., 2002). Kay et al. (2004) reported that in a farm where
animal manures were applied, no detectable tylosin was found in tile drainage water whereas
sulfachloropyridazine and oxytetracycline were frequently detected. The relatively low
occurrence of tylosin in surface water at that farm is possibly due to relatively quick abiotic
and/or biotic transformations in soils. Hu and Coats (2007) reported that the dissipation half-life
of tylosin in soils was approximately 7 days.
The level and frequency of pharmaceuticals present in surface runoff at the farm are
influenced by the timing of animal manure application, plant growth in manure-applied soils, and
weather conditions. During our sampling period, a relatively small amount of liquid animal
16 
 

 
5

manure (1.6 x 10 L) was applied to the field in August 2006. The application during post6

harvest season (October and November 2006) was 1.9 × 10 L of liquid manure. The more
6

intensive application of manure to the fields occurred in spring in which 4.1 × 10 L of liquid
6

and 1.4 x 10 kg of solid manures were applied in April and May 2007. This application
schedule resulted in the highest detection frequency, and the highest concentrations of
pharmaceuticals, in water samples collected in April 2007 (3 days after manure application); all
these samples were found to contain monensin and amprolium. For the stagnant water samples,
-1

-1

concentrations ranged from 17 to 260 ng L for amprolium, and 5 to 189 ng L for monensin.
In comparison, concentrations of these pharmaceuticals in drainage water samples were lower.
Presumably, convergence of surface runoff into the drainage tile system diluted the
pharmaceutical concentrations in water thereby narrowing the concentrations to between 17 and
-1

-1

53 ng L for amprolium, and 4 and 15 ng L for monensin (Table I-2). The high number of
detections of veterinary pharmaceuticals in the April samples is consistent with the study by
Dolliver and Gupta (2008) in which the highest detection frequency of antibiotics (i.e., monensin
and tylosin) in leachates from a farm in Wisconsin was also in April. A large amount of animal
manure was applied in April 2007. The relatively cold weather in Michigan at this time of year
likely inhibited pharmaceutical transformations by reducing microbial activity and abiotic
reaction rates. Snowmelt commencing during April 2007 in Michigan increased the soil-water
contact time thereby enhancing the release of soil-sorbed pharmaceuticals into water. These
results suggest that manure application after harvest (i.e., in the late fall and early spring seasons)

17 
 

 

in cold regions could lead to the dissemination of pharmaceuticals to the surrounding water
during spring snowmelt.
Relatively cold temperatures could be responsible for the enhanced detection frequency
of pharmaceuticals in the samples collected in October, November, and December 2006 (Table I2). Amprolium and monensin were frequently detected in surface runoff samples taken during
October to December 2007, even though a comparatively small amount of manure was applied to
the field during this period. Although a much larger volume of manure was applied in April and
May 2007, pharmaceuticals were only occasionally detected in surface water during the growing
season (i.e. from May to September 2007). This is likely due to the enhanced biotic and abiotic
pharmaceutical transformations at comparatively high temperatures during the growing season,
as well as more biological activities in the plant root zones. It was noted that high concentration
and frequency of amprolium were found in the water samples collected in July 2007 due possibly
to the small precipitation (9.9 mm) compared to the results in other months during that period.
Dolliver and Gupta (2008) reported a significant number of detections of monensin and tylosin in
farm runoff and leachate in the non-growing season vs. the growing season. Taken together,
these results indicate that land application of manure shortly before the growing season (e.g.,
early May) could result in less dissemination of veterinary pharmaceuticals to farm runoff water
as compared with the application during the non-growing season.
Among all the sampling sites, water samples collected from site 9 had the highest
frequency of detection, and contained relatively high levels of amprolium and monensin (Table
I-2). This site had a small ditch connecting it to a manure pool. The overflow from the livestock
manure pool concurrent with each rainfall event most assuredly resulted in localized high
concentrations of pharmaceuticals in the stagnant surface water. In this instance, direct discharge
18 
 

 

of animal waste to surface runoff served as a significant point source responsible for
contamination of agricultural environment with veterinary pharmaceuticals.
To examine soil contamination by veterinary pharmaceuticals, 13 surface soil grab
samples were collected from the farm in May and October 2006. Again, similar to the water
samples, amprolium and monensin were the two pharmaceuticals most frequently detected in
soils. Amprolium and monensin were found above quantification limits in 21 and 24 of the 26
soil samples, respectively. The concentration ranges were determined between 0.03 to 0.26 μg
-1

kg-1 for amprolium, and 0.004 to 0.50 µg kg for monensin (Figure I-3). The average
-1

concentration of samples collected in October 2006 was 0.01 µg kg for monensin, and 0.08 µg
-1

kg for amprolium, which were less than the average concentrations in soils sampled in May
-1

-1

2006 (0.095 µg kg for monensin and 0.11 µg kg for amprolium). The frequent detection of
veterinary pharmaceuticals in soil suggests that soil could be a large reservoir for these bioactive
compounds. These contaminants could subsequently be released into water through desorption
from soil into water and/or movement of soil particles containing sorbed pharmaceuticals with
surface runoff. These processes lead to the dissemination of manure-borne pharmaceuticals in
agricultural environment.

19 
 

 

Concentrations (μg/kg)

0.6

Monensin

Amprolium

0.5
0.4
0.3
0.2

90th

0.1

75th
median
25th
10th

0.0

May 2006 Oct 2006 May 2006 Oct 2006

 

Figure I-3. Distribution of concentration of amprolium and monensin in soils sampled in May
th

th

and October 2006. The box-and-whisker plot represents the range, and the 10 , 25 , median,
th

th

75 , and 90 percentile of concentration distribution

During sampling in October 2006 five soil samples were collected beneath stagnant
surface water at site 7 to 11 (Table I-2). At site 11, amprolium was not detected in surface water
or in the soil beneath. Monensin, however, was present in both phases. In all other samples
amprolium and monensin were detected in both soil and water. The concentrations of amprolium
and monensin determined in soils were plotted against their corresponding aqueous
concentrations (Figures I-4A and I-4B), and it was observed that the concentrations in soil
increased with increasing aqueous concentrations. The distribution coefficients, defined here as
20 
 

 

the soil-sorbed concentration divided by the corresponding concentration in water, were ca. 1.4,
-1

2.0, 2.8, 3.2, and 11.0 L kg for monensin, and 0.4, 0.7, 0.8, and 2.3 L kg-1 for amprolium, at
sites 7-11, respectively.

21 
 

 

4000

A: monensin

0.04
0.03
0.02
0.01
0.00
0.00

Field measurement
Colwood
Capac A
Capac B
Oshtemo

3000

Concentration in Soil (mg/kg)

2000

0.01

0.02

1000
0
70000

0
100
200
B: amprolium

300

60000
50000
40000

400

500

600

700

1.0
0.8
0.6
0.4
0.2
0.0
0.00 0.05 0.10 0.15 0.20

30000
20000
10000
0
0

500

1000

1500

2000

Concentration in Water (mg/L)

 

Figure I-4. Sorption isotherms of (A) monensin and (B) amprolium obtained from laboratory
studies using soils adjacent to the study site which were devoid of target pharmaceuticals (open
symbols). The filled symbols in the inserts represent soil-sorbed concentrations/aqueous
concentrations measured for field samples. These data allow comparison of soil-water
distribution of monensin and amprolium in actual field samples to that expected based on
sorption measurement using soils without manure amendment
22 
 

 

Sorption isotherms of monensin and amprolium were measured for four soils obtained
from sites nearby the farm which contained no detectable levels of the target pharmaceuticals.
Sorption isotherms were nearly linear for the relatively low aqueous concentrations measured,
-1

-1

i.e., up to 600 µg L for monensin, and 1600 µg L for amprolium (Figure I-4). The distribution
-1

coefficients (taken from the slopes of the linear isotherms) were from 1.4 to 8.3 L kg for
-1

monensin, and 8.4 to 43.6 L kg for amprolium. Soil organic matter contents of the four soils
ranged from 1.1 to 6.3 %. Normalization of the distribution coefficients to SOM (Kom) manifests
greater variation in the sorption coefficients, implying that soil components other than SOM, e.g.,
soil minerals, also contribute to sorption. Sassman and Lee (2007) measured the sorption
isotherms of monensin by several soils; the isotherms were linear with the distribution
-1

coefficients varying from 0.9 to 33.7 L kg . Assuming the sorption isotherms remain linear
-1

down to ng L levels, the estimated distribution coefficients for monensin from field sites are
within the same order of magnitude as those obtained from the laboratory measurements (insert
in Figure I-4A). This consistency in soil-water distribution of monensin among different sites
and studies suggests that monensin sorbed by soils is predictably released into the surface water
during rainfall events. In contrast to monensin, the distribution coefficients estimated from field
data for amprolium are 4 to 100 times less than the sorption measured in laboratory (insert in
Figure I-4B). This distinction can be partially attributed to the relatively low extraction recovery
of amprolium from soil (i.e., 53%). However, the recovery-corrected distribution coefficients
only increase by approximately two times, which is insufficient to account for the discrepancy in
sorption magnitude obtained from laboratory studies vs field samples. Unlike monensin (an

23 
 

 

organic acid with pKa = 4.2), amprolium has a permanently positive charge that can engage in
strong electrostatic interactions with the negatively charged sites associated with DOM. Several
recent studies (Gu et al., 2007; MacKay and Canterbury, 2005; Sibley and Pedersen, 2008) have
demonstrated that cationic/zwitterionic antibiotics, such as tetracyclines and clarithromycin, can
be strongly associated with dissolved humic acid in aqueous solution. Soil amended with animal
manure can result in a significant amount of DOM present in surface runoff (Clark et al., 1998).
Hence, it is plausible to suggest that amprolium forms associations with DOM thereby reducing
its sorption by soil.

24 
 

 

Conclusions
This study indicates that land application of animal manures could lead to the
accumulation of veterinary pharmaceuticals in agricultural lands. The pharmaceuticals are more
frequently detected in surface runoff from the farm during non-growing season (October to April)
than during growing season (May to September). Pharmaceuticals resulting from post-harvest
manure application appear to be more persistent than those from spring application. Soil is a
reservoir for the veterinary pharmaceuticals that can be disseminated to nearby surface water by
rainfall events. Several factors, such as manure application timing and rates, weather conditions,
soil types, landscape, and crop growth, are confluent to affect the fate and transport of veterinary
pharmaceuticals in the agricultural environment. A better understanding of the major factor(s)
controlling veterinary pharmaceuticals in surface runoff is useful for guiding the land application
of livestock manures while minimizing the release of veterinary pharmaceuticals to the
environment from agricultural practice. More systemic evaluations of the input/output and
dissemination of veterinary pharmaceuticals, along with detailed information of agricultural
production, are also needed to assess the potential risks posing to human and ecosystem health.

25 
 

 

REFERENCE

26 
 

 

Reference

Aga, D.S., S. O'Connor, S. Ensley, J.O. Payero, D. Snow, and D. Tarkalson. 2005.
Determination of the persistence of tetracycline antibiotics and their degradates in
manure-amended soil using enzyme-linked immunosorbent assay and liquid
chromatography-mass spectrometry. Journal of Agricultural and Food Chemistry
53:7165-7171.
Batt, A.L., D.D. Snow, and D.S. Aga. 2006. Occurrence of sulfonamide antimicrobials in private
water wells in Washington County, Idaho, USA. Chemosphere 64:1963-1971.
Brown, K.D., J. Kulis, B. Thomson, T.H. Chapman, and D.B. Mawhinney. 2006. Occurrence of
antibiotics in hospital, residential, and dairy, effluent, municipal wastewater, and the Rio
Grande in New Mexico. Science of the Total Environment 366:772-783.
Christian, T., R.J. Schneider, H.A. Farber, D. Skutlarek, M.T. Meyer, and H.E. Goldbach. 2003.
Determination of antibiotic residues in manure, soil, and surface waters. Acta
Hydrochimica Et Hydrobiologica 31:36-44.
Clark, M.S., W.R. Horwath, C. Shennan, and K.M. Scow. 1998. Changes in soil chemical
properties resulting from organic and low-input farming practices. Agronomy Journal
90:662-671.
Dolliver, H.A.S., and S.C. Gupta. 2008. Antibiotic losses from unprotected manure stockpiles.
Journal of Environmental Quality 37:1238-1244.
Gu, C., K.G. Karthikeyan, S.D. Sibley, and J.A. Pedersen. 2007. Complexation of the antibiotic
tetracycline with humic acid. Chemosphere 66:1494-1501.
Halling-Sorensen, B., A.M. Jacobsen, J. Jensen, G. Sengelov, E. Vaclavik, and F. Ingerslev.
2005. Dissipation and effects of chlortetracycline and tylosin in two agricultural soils: A
field-scale study in southern Denmark. Environmental Toxicology and Chemistry
24:802-810.
Hamscher, G., S. Sczesny, H. Hoper, and H. Nau. 2002. Determination of persistent tetracycline
residues in soil fertilized with liquid manure by high-performance liquid chromatography
27 
 

 

with electrospray ionization tandem mass spectrometry. Analytical Chemistry 74:15091518.
Hamscher, G., H.T. Pawelzick, H. Hoper, and H. Nau. 2005. Different behavior of tetracyclines
and sulfonamides in sandy soils after repeated fertilization with liquid manure.
Environmental Toxicology and Chemistry 24:861-868.
Hao, C.Y., L. Lissemore, B. Nguyen, S. Kleywegt, P. Yang, and K. Solomon. 2006.
Determination of pharmaceuticals in environmental waters by liquid
chromatography/electrospray ionization/tandem mass spectrometry. Analytical and
Bioanalytical Chemistry 384:505-513.
Hu, D., and J.R. Coats. 2007. Aerobic degradation and photolysis of tylosin in water and soil.
Environmental Toxicology and Chemistry 26:884-889.
Jacobsen, A.M., B. Halling-Sorensen, F. Ingerslev, and S.H. Hansen. 2004. Simultaneous
extraction of tetracycline, macrolide and sulfonamide antibiotics from agricultural soils
using pressurised liquid extraction, followed by solid-phase extraction and liquid
chromatography-tandem mass spectrometry. Journal of Chromatography A 1038:157-170.
Kay, P., P.A. Blackwell, and A.B.A. Boxall. 2004. Fate of veterinary antibiotics in a
macroporous tile drained clay soil. Environmental Toxicology and Chemistry 23:11361144.
Kay, P., P.A. Blackwell, and A.B.A. Boxall. 2005. Transport of veterinary antibiotics in overland
flow following the application of slurry to arable land. Chemosphere 59:951-959.
Kim, S.C., and K. Carlson. 2007. Temporal and spatial trends in the occurrence of human and
veterinary antibiotics in aqueous and river sediment matrices. Environmental Science &
Technology 41:50-57.
Kolpin, D.W., E.T. Furlong, M.T. Meyer, E.M. Thurman, S.D. Zaugg, L.B. Barber, and H.T.
Buxton. 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in
US streams, 1999-2000: A national reconnaissance. Environmental Science &
Technology 36:1202-1211.
Kuchta, S.L., A.J. Cessna, J.A. Elliott, K.M. Peru, and J.V. Headley. 2009. Transport of
Lincomycin to Surface and Ground Water from Manure-amended Cropland. Journal of
Environmental Quality 38:1719-1727.
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Larsbo, M., D.R. Lapen, E. Topp, C. Metcalfe, K.C. Abbaspour, and K. Fenner. 2009.
Simulation of Pharmaceutical and Personal Care Product Transport to Tile Drains after
Biosolids Application. Journal of Environmental Quality 38:1274-1285.
Lissemore, L., C.Y. Hao, P. Yang, P.K. Sibley, S. Mabury, and K.R. Solomon. 2006. An
exposure assessment for selected pharmaceuticals within a watershed in southern Ontario.
Chemosphere 64:717-729.
MacKay, A.A., and B. Canterbury. 2005. Oxytetracycline sorption to organic matter by metalbridging. Journal of Environmental Quality 34:1964-1971.
Mellon, M., C. Benbrook, and K.L. Benbrook. 2001. Hogging it: estimates of antimicrobial
abuse in livestock Union of Concerned Scientists Publications, Cambridge, MA.
Sassman, S.A., and L.S. Lee. 2007. Sorption and degradation in soils of veterinary ionophore
antibiotics: monensin and lasalocid. Environmental Toxicology and Chemistry 26:16141621.
Sibley, S.D., and J.A. Pedersen. 2008. Interaction of the macrolide antimicrobial clarithromycin
with dissolved humic acid. Environmental Science & Technology 42:422-428.
Snow, D.D., S.L. Bartelt-Hunt, S.E. Saunders, S.L. Devivo, and D.A. Cassada. 2008. Detection,
Occurrence and Fate of Emerging Contaminants in Agricultural Environments. Water
Environment Research 80:868-897.
Song, W.L., M. Huang, W. Rumbeiha, and H. Li. 2007. Determination of amprolium, carbadox,
monensin, and tylosin in surface water by liquid chromatography/tandem mass
spectrometry. Rapid Communications in Mass Spectrometry 21:1944-1950.
Stoob, K., H.P. Singer, S.R. Mueller, R.P. Schwarzenbach, and C.H. Stamm. 2007. Dissipation
and transport of veterinary sulfonamide antibiotics after manure application to grassland
in a small catchment. Environmental Science & Technology 41:7349-7355.
Tolls, J. 2001. Sorption of veterinary pharmaceuticals in soils: A review. Environmental Science
& Technology 35:3397-3406.

29 
 

 

CHPATER II
DETERMINATION OF PHARMACEUTICALS IN BIOSOLIDS USING ACCERLATED
SOLVEN EXTRACTION AND LIQUID CHROMATOGRAPHY/TANDEM MASS
SPECTROMETRY

30 
 

 

Abstract
An analytical method was developed to quantitatively determine pharmaceuticals in
biosolid (treated sewage sludge) from wastewater treatment plants (WWTPs). The collected
biosolid samples were initially freeze dried, and grounded to obtain relatively homogenized
powders. Pharmaceuticals were extracted using accelerated solvent extraction (ASE) under the
optimized conditions. The optimal operation parameters, including extraction solvent,
temperature, pressure, extraction time and cycles, were identified to be acetonitrile/water mixture
(v/v 7:3) as extraction solvent with 3 extraction cycles (15 minutes for each cycle) at 100 °C and
100 bars. The extracts were cleaned up using solid-phase extraction followed by determination
by liquid chromatography coupled with tandem mass spectrometry. For the fifteen target
pharmaceuticals commonly found in the environment, the overall method recoveries ranged from
49% to 68% for tetracyclines, 64% to 95% for sulfonamides, and 77% to 88% for other
pharmaceuticals (i.e. acetaminophen, caffeine, carbamazepine, erythromycin, lincomycin and
tylosin). The developed method was successfully validated and applied to the biosolid samples
collected from WWTPs located in six cities in Michigan. Among the fifteen target
pharmaceuticals, fourteen pharmaceuticals were detected in the collected biosolid samples. The
average concentrations ranged from 2.6 μg/kg for lincomycin to 743.6 μg/kg for oxytetracycline.
These results indicated that pharmaceuticals could survive wastewater treatment processes, and
accumulate in sewage sludge and biosolids. Subsequent land application of the contaminated
biosolids could lead to the dissemination of pharmaceuticals in soil and water environment,
which poses potential threats to at-risk populations in the receiving ecosystems.

31 
 

 

Introduction
Administration of pharmaceuticals is the most common practice for human disease
control and treatments. Large fractions of administrated dosage are, in fact, not assimilated or
metabolized in human body, but released into influents to municipal wastewater treatment plants
(WWTPs) as either parent compounds or bioactive metabolites (Halling-Sorensen et al., 1998;
Leclercq et al., 2009; Plosz et al., 2010). Unfortunately, currently operating WWTPs are not
designed for effective removal of pharmaceuticals; as a result, a portion of pharmaceuticals could
survive wastewater treatment processes, and disseminate to the environments via effluents from
WWTPs in a dissolved form (Daughton and Ternes, 1999; Ternes, 1998) or sorbed
pharmaceuticals by sewage sludge (Carballa et al., 2004). Most sewage sludge is treated to
produce biosolid and is subsequently applied to agricultural lands for its fertilizer value. This
practice undoubtedly introduces sludge-born pharmaceuticals into soil and water environment
where their fate and risks remain largely unknown (Kolpin et al., 2002; Lapen et al., 2008;
Lindberg et al., 2005; Lozano et al., 2010; Song et al., 2007; Ternes, 2001). There is growing
concern regarding the potential impacts of these pharmaceuticals on at-risk populations in the
ecosystems.
The information on pharmaceutical compositions and concentrations in land-applied
biosolids is critical to assessing the exposure and potential risks. Analytical methods have been
developed to analyze pharmaceuticals in water (Lin et al., 2007; Lindberg et al., 2004; Lindberg
et al., 2005; Song et al., 2007), animal manure (Blackwell et al., 2004; Haller et al., 2002;
Jacobsen and Halling-Sorensen, 2006; Prado et al., 2009) and soil (Blackwell et al., 2004;
Morales-Munoz et al., 2004; Turiel et al., 2006; Vazquez-Roig et al., 2010). In several studies
analytical methods have been also reported to identify and quantify pharmaceuticals in sewage
32 
 

 

sludge from WWTPs (Aparicio et al., 2007; Fernandez-Sanjuan et al., 2009; Garcia-Valcarcel
and Tadeo, 2009; Golet et al., 2002; Lillenberg et al., 2009; Lindberg et al., 2005). These
methods are worthy of further examination for their applicability to multiple classes of
pharmaceuticals in sewage sludge or biosolids (Diaz-Cruz et al., 2009). The procedure of the
analytical methodology usually consists of extraction of pharmaceuticals from sludge/biosolid,
cleanup of extracts, and analysis using liquid or gas chromatography. Liquid-solid partitioning is
the most common approach utilized in combination of ultrasonic-assisted extraction (Blackwell
et al., 2004), microwave-assisted extraction (Turiel et al., 2006) and accelerated solvent
extraction (ASE). The extraction efficiency is dependent of sample matrices and characteristics
of target analytes (Blasco et al., 2009; Navarro et al., 2009; Sanchez-Prado et al., 2010). Among
these three extraction techniques, ASE becomes more preferred because it provides apparent
advantages such as less solvent assumption, automatic procedure for simultaneous extraction of
multiple samples, short sample preparation time, and higher extraction recoveries. Solid-phase
extraction (SPE) is usually selected as the cleanup tool for the sludge/biosolid extracts, and the
corresponding cartridges commonly used included Waters Oasis hydrophilic-lipophilic balance
(HLB), strong-anion exchange (SAX), strong cation-exchange (SCX) or their combinations
(Lillenberg et al., 2009; Nie et al., 2009; Vazquez-Roig et al., 2010). The pharmaceuticals are
then analyzed using liquid chromatography with fluorescence or ultraviolet detectors (Golet et al.,
2002; Raich-Montiu et al., 2007; Rebiere et al., 2007) or gas chromatography coupled to mass
spectrometer (Van Hoeck et al., 2009). Recently, liquid chromatography/tandem mass
spectrometry (LC-MS/MS) became the most efficient tool for identifying and quantifying trace
levels of pharmaceuticals in environmental samples (Diaz-Cruz et al., 2006; Hermo et al., 2005;
Nakata et al., 2005; Song et al., 2007; Ternes et al., 2005).

33 
 

 

The objective of this study was to develop a robust, sensitive and practical analytical
method to simultaneously identify and quantify multiple classes of pharmaceuticals in biosolids.
The ASE experimental parameters (i.e., extraction solvent, pressure, temperature, extraction time
and cycles) were optimized to achieve the maximum extraction efficiencies for pharmaceuticals.
SPE was used to pre-concentrate and clean up the ASE extracts. The amounts of pharmaceuticals
were quantitated by a LC-MS/MS equipped with electrospray ionization (ESI) source. The
method was validated, and applied to determine multiple classes of pharmaceuticals in biosolid
samples collected from several WWTPs of different locations in Michigan, USA.

34 
 

 

Experimental
Chemicals and materials
Tetracycline, demeclocycline, chlortetracycline, oxytetracycline, meclocycline,
sulfadiazine, sulfamerazine, sulfamethazine, sulfamethoxazole, carbamazepine, acetaminophen,
caffeine, tylosin and lincomycin were purchased from Sigma-Aldrich Chemical Company (St.
Louis, MO, USA). Doxycycline was obtained from Fisher Bioreagents (Pittsburgh, PA, USA).
13

C6-sulfamethazine was obtained from Cambridge Isotope Laboratories (Andover, MA, USA).

Simeton was purchased from AccuStandard (New Haven, CT, USA). Methanol, sodium
ethylenediaminetetraacetate (EDTA), formic acid, and sulfuric acid were purchased from J.T.
Baker (Phillipsburg, NJ, USA). Acetonitrile was purchased from EMD Chemicals (Gibbstown,
NJ, USA). Waters Oasis hydrophilic-lipophilic balance (HLB) cartridge was purchased from
Waters Corporation (Milford, MA, USA).
Biosolid samples
During the winter season of 2009, six grab biosolid samples were collected from WWTPs
of Lansing, East Lansing, St. Clair, Plainwell, Traverse City, and Imlay City, Michigan, USA.
The biosolids collected from WWTPs of Traverse City, Plainwell, St. Clair, Imlay City, and
Lansing were classified as class B biosolids which meet the criteria for land application. The
biosolids collected from East Lansing WWTP was sent off to landfills. At each sampling site,
one liter of biosolid was collected, and stored in a polypropylene narrow-necked bottle. The
bottles were capped, placed in an iced cooler, and transported immediately to laboratory. The
samples were then freeze-dried to obtain dry solids, ground using a mortar and pestle, passed
through a 0.5 mm sieve, and stored in a refrigerator (-20 ºC) prior to use.

35 
 

 

Sample extraction and cleanup
The dried biosolid sample was weighed (500.0 mg) into cellulose extraction thimble, and
mixed with diatomaceous earth (5.0 g) to prevent aggregation during extraction process.

13

C6-

sulfamethazine and meclocycline were used as surrogates for sulfonamide and tetracycline
antibiotics, respectively. The collected biosolid samples had been checked not to contain
meclocycline. 100 µL of

13

C6-sulfamethazine (1.0 mg/L) and 200 µL of meclocycline (1.0 mg/L)

were spiked into the biosolid samples, which were then placed on a rotator and mixed overnight.
The thimbles were then placed into 22 mL stainless steel extraction cells, and extracted using a
Dionex ASE 200 accelerated solvent extractor (Sunnyvale, CA, USA). The extraction solvent
and operating conditions had been optimized in which acetonitrile and water mixture (v/v = 7:3)
was used. The samples were preheated for 5 min, and then extracted at 100 °C under 100 bars.
The static extraction period was set 15 min with three extraction cycles. The flushing solvent was
100% of cell volume, and extraction cells were purged with N2 for 120 s. The extracts were
collected and transferred to volumetric flasks, and diluted to 100 mL with deionized water. The
pH value of the solution was adjusted to ~3.0 using H2SO4. An aliquot of 10.0 mL of the diluted
solution passed through a preconditioned HLB solid-phase cartridge (Waters Corporation,
Millford, MA, USA) at a rate of 1 mL/min. The analyte-loaded HLB cartridge was rinsed with 3
mL of deionized water, and eluted by 5 mL of methanol-water mixture (v/v = 1:1) containing
150 mg/L EDTA. The elute volume was reduced to 1.0 mL using a gentle N2 flow to concentrate
the samples, which took ~45 min. The prepared samples were transferred to amber LC

36 
 

 

autosampler vials, and 10 ng of simeton (internal standard) was added prior to LC-MS/MS
analysis.
Liquid chromatography and tandem mass spectrometry
The prepared samples were analyzed using a LC-MS/MS system consisting of a
Shimadzu high-performance liquid chromatography (HPLC, Columbia, MD, USA) fully
integrated with an Applied Biosystems Sciex 3200 triple quadrupole mass spectrometer (Foster
City, CA, USA). A PEAK Scientific gas generator (NM20ZA, Billerica, MA) was used to supply
gases required for the mass spectrometer. Turbo IonSpray source of the mass spectrometer was
operated with electrospray ionization source in positive mode, ionspray voltage at 5500 V and
temperature at 600 °C. Curtain gas pressure was 10 psi, collision gas pressure was 6 psi, and ion
source gas pressure was 10 psi. The compound-dependent mass spectrometer parameters were
optimized for each analyte (Table II-1). Pharmaceutical concentrations were quantified using
multiple-reaction monitoring mode. Two pairs of precursor/product ion transitions were
simultaneously monitored during analysis in order to unambiguously identify the target
pharmaceuticals in environmental samples (Ghanem et al., 2007; Song et al., 2007). The pair of
precursor and product 1 ions were selected to quantify the analytes because the product ion 1
manifested a relatively greater response abundance than product ion 2.

37 
 

 

Table II-1. Precursor, product ions and mass spectrometer parameters used to identify and quantify pharmaceuticals

Pharmaceuticals
Sulfadiazine
Sulfamerazine
Sulfamethazine
Sulfamethoxazole
Chlortetracycline
Demeclocycline
Doxycycline
Oxytetracycline
Tetracycline
Acetaminophen
Caffeine
Carbamazepine
Erythromycin
Lincomycin
Tylosin
a

Precursor ion
(m/z)
250.9
264.9
279.0
254.1
479.2
465.0
445.4
461.3
445.4
152.0
195.0
237.1
734.6
407.3
916.5
b

Product ion 1
(m/z)
108.2
155.7
186.1
156.2
444.2
448.1
428.2
426.4
428.2
110.0
138.0
194.3
576.4
126.3
174.0

Product ion 2
(m/z)
156.1
172.4
162.0
107.8
462.2
430.3
339.3
444.2
339.3
93.0
110.0
165.4
558.4
359.2
771.9

c

d

b

EP
(Volts)
9
4
5
8
4
9
9
7
9
10
5
3
11
5
9
e

c

CEP
(Volts)
30
20
20
15
27
30
50
22
50
150
20
14
30
22
36

d

CE
(Volts)
27
22
23
22
27
27
31
21
31
23
29
23
32
33
47

declustering potential. entrance potential. cell entrance potential. collision energy. collision cell exit potential

38 
 

a

DP
(Volts)
37
45
37
38
40
36
33
28
33
40
49
50
17
56
63

e

CXP
(Volts)
36
10
18
14
40
36
45
41
45
31
16
18
48
11
14

 

The Shimadzu HPLC system consisted of two LC-20AD pumps, a SIL-20A autosampler
and a DGU-20A degasser operated by a CBM-20A controller. The analytes were separated using
a Phenomenex Luna C18 column (150 mm × 4.6 mm, particle size: 3 µm, Torrance, CA, USA).
A binary gradient mobile phase was applied with a flow rate of 350 µL/min in which phase A
(water) and phase B (acetonitrile) both contained 0.1 % (v/v) formic acid. The mobile phase
gradient was programmed as: at 0 min 100% phase A; 0 to 2.0 min phase B linearly increased to
12%; 2 to 20 min phase A and phase B held at the ratio of 88:12; 20 to 22 min phase B linearly
increased to 25%; 22 to 35 min phase B increased to 40%; 35 to 40 min phase B increased to
80%, and held at this ratio until 42 min. From 42 to 43 min phase B reduced to 0%, and
equilibrated in the flow system for 5 min before next injection. The injection volume was 10 µL.
After each sample injection, the first three-minute elution was diverted to waste through a builtin valve, and then switched to the inlet of tandem mass spectrometer.
Method validation
External standard solutions were prepared in methanol-water mixture (v/v = 1/1)
containing 150 mg/L EDTA. The same background solution was also used to elute analytes from
the HLB cartridge during the cleanup step of sample extracts. Standard curves were established
with the concentrations spanning 2 orders of magnitude. The method sensitivity was determined
by analyzing pharmaceuticals at ~20 μg/L for 10 times using the ASE extracts fortified with
target compounds, and standard deviation of these measurements was calculated. Limit of
detection (LOD) was estimated as three times the standard deviation, and limit of quantification
(LOQ) was defined as ten times the standard deviation. To evaluate extraction efficiency and
repeatability of the overall method, the selected target pharmaceuticals and two surrogates (i.e.,

39 
 

 
13

C6-sulfamethazine and meclocycline) were spiked into biosolid samples, and mixed together

overnight. Eight replicates of spiked biosolid samples were prepared followed by accelerating
solvent extraction, SPE cleanup and LC-MS/MS analysis (described above). Biosolid control
samples without spiked pharmaceuticals were also prepared; the amounts of pharmaceuticals
measured in the control were subtracted from the amounts measured in the spiked samples. The
sample matrix effects on the responses of tandem mass spectrometer was evaluated via spiking
known amounts of target pharmaceuticals into the extract of biosolid which was obtained by
ASE followed by the SPE cleanup. The measured pharmaceutical concentrations in the spiked
samples were subtracted by the background pharmaceutical amounts originally present in the
biosolids, and compared with the amounts of the spiked analytes.

40 
 

 

Results and discussion
Selection of extraction solvent
The biosolid sample collected from Lansing WWTP was fortified with carbamazepine,
sulfamethazine, sulfamethoxazole, tetracycline, chlortetracycline and oxytetracycline with
concentrations ranging from 780 to 1862 μg/kg on dry matter basis, and subjected to ASE using
water and organic solvent mixtures. The extracted concentrations of pharmaceuticals were
subtracted by the measured concentrations of the corresponding compounds in the unfortified
biosolid sample, and reported in Table II-2. Extraction solvents used here were aqueous
methanol or acetonitrile mixtures with volume ratios of organic solvent to water 7:3, 5:5 and 3:7.
Several previous studies indicated that polar organic solvents (e.g. acetonitrile, methanol) and
water mixtures manifested the superior capability to extract pharmaceuticals from sewage sludge
and soil (Hermo et al., 2005; Jacobsen and Halling-Sorensen, 2006; Lillenberg et al., 2009).
Among the extraction solvents tested, it appeared that higher extraction efficiencies were
achieved for the mixtures containing greater contents of organic solvent (Table II-2).
Acetonitrile/water mixture (v/v = 7:3) was more effective to remove pharmaceuticals from
biosolids than methanol/water mixture. Therefore, acetonitrile/water mixture (v/v = 7:3) was
chosen as the extraction solvent in this study. To evaluate the influence of pH on extraction
efficiency, pH of acetonitrile/water mixture (v/v = 7:3) was adjusted to ~2.0 using hydrochloric
acid, and then extracted the selected pharmaceuticals from the biosolid. The use of acidified
extraction solvent, in fact, lowered extraction efficiency by 36% for tetracycline, 42% for
chlortetracycline, and 75% for oxytetracycline compared to unacidified acetonitrile/water
mixture; no apparent discrepancy was observed for extraction of carbamazepine and
sulfonamides. Several previous studies indicated that acidified water-organic solvent mixture
41 
 

 

(e.g., pH < 2.5) enhanced extraction recoveries of pharmaceuticals, but the target
pharmaceuticals were not the same as those in this study (Golet et al., 2002; Lillenberg et al.,
2009). The acidification of extraction solvent protonates the acidic functional groups (e.g.,
carboxylic acids and phenols) in the organic fractions of sewage sludge/biosolids, and thereby
reduces electrostatic interactions between sewage sludge/biosolid and cationic moiety of
tetracyclines. In these previous studies, sand was used to mix with sewage sludge or soil samples
prior to extraction (Crescenzi et al., 2000; Golet et al., 2002; Lillenberg et al., 2009).
Diatomaceous earth, used as the dispersant of biosolid in our study, could release a certain
+

amount of inorganic cations (e.g., Na , Mg

2+

etc.) which could effectively suppress the

interactions of biosolid with cationic speciation of tetracyclines hence resulting in higher
extraction efficiency.

42 
 

 

Table II-2. Extracted concentrations of pharmaceuticals vs. spiked concentrations from biosolid by water-organic solvent mixtures
using accelerating solvent extraction

Pharmaceuticals
Sulfamethazine
Sulfamethoxazole
Chlortetracycline
Oxytetracycline
Tetracycline
Carbamazepine
a

Spiked
conc.
(g/kg)
982
780
1512
1862
1518
840

a

Extraction solvents
acetonitrile
/water
(7:3)
b

947 ± 21
564 ± 29
786 ± 18
1008 ± 25
1065 ± 28
700 ± 32

acetonitrile/water
(7:3), pH 2.0

acetonitrile
/water (5:5)

acetonitrile
/water (3:7)

920 ± 31
710 ± 13
456 ± 5
261 ± 21
681 ± 76
705 ± 53

551 ± 25
494 ± 17
540 ± 17
627 ± 24
538 ± 4
480 ± 34

393 ± 11
266 ± 28
306 ± 23
308 ± 12
282 ± 39
272 ± 19

b

methanol
/ water
(5:5)
475 ± 23
459 ± 20
525 ± 24
655 ± 10
359 ± 20
260 ± 18

Operation conditions: 100 bars, 100 °C, 3 × 15 min. measured concentration: mean ± standard deviation (n = 3)

43 
 

methanol
/ water
(7:3)
758 ± 24
541 ± 27
708 ± 20
776 ± 14
628 ± 36
437 ± 27

methanol/
water
(3:7)
368 ± 28
339 ± 17
373 ± 22
360 ± 10
302 ± 35
214 ± 15

 

Optimization of ASE operating parameters
High pressure is applied during ASE operation in order to keep extraction solvent in
liquid state while temperature is held at or above the boiling point. The application of high
pressure also facilitates the contact of extraction solvent with analytes via penetrating into the
domains in biosolids where the extraction solvent could not enter under normal atmospheric
pressure. In this study four levels of pressure (55, 80, 100, and 130 bars) were applied to the
extraction, and the corresponding extracted concentrations are reported in Table II-3. The
pharmaceutical concentrations at different extraction pressures were compared using a linearmixed model (SAS 9.1 software, SAS Institute Inc., Cary, NC, USA). Significant level was set at
p < 0.05 with 95% confidence. The results indicate that the extraction efficiencies at pressure of
100 bars were significantly higher than those conducted at other pressures (p = 0.0049). This
extraction pressure falls within the similar range reported in several previous studies (Golet et al.,
2002; Lillenberg et al., 2009).

44 
 

 

Table II-3. Extraction efficiency of pharmaceuticals under varying accelerating solvent extraction parameters

Pharmaceuticals

Sulfadiazine
Sulfamerazine
Sulfamethazine
Sulfamethoxazole
Chlortetracycline
Demeclocycline
Doxycycline
Oxytetracycline
Tetracycline
Acetaminophen
Caffeine
Carbamazepine
Erythromycin
Lincomycin
Tylosin
 

Spiked
Conc.
(μg/kg)
466
414
491
390
756
507
509
931
759
788
391
420
636
538
484
 

Pressurized liquid extraction conditions
Pressure (bar)
Extraction time (min)
55

80

100

130

2 × 15

3 × 15

2 × 25

3 × 25

270 ± 6
249 ± 10
437 ± 12
274 ± 17
413 ± 22
316 ± 10
335 ± 7
496 ± 43
398 ± 52
617 ± 21
406 ± 8
377 ± 29
231 ± 22
424 ± 11
363 ± 14

246 ± 6
222 ± 17
385 ± 11
256 ± 33
488 ± 14
151 ± 8
266 ± 5
483 ± 54
315 ± 37
579 ± 12
397 ± 12
365 ± 38
266 ± 28
480 ± 21
359 ± 26

293 ± 13
289 ± 7
456 ± 11
279 ± 11
434 ± 29
304 ± 11
355 ± 12
523 ± 60
475 ± 50
761 ± 32
309 ± 10
389 ± 45
408 ± 28
449 ± 22
365 ± 25

243 ± 16
242 ± 15
396 ± 26
264 ± 26
365 ± 20
300 ± 12
328 ± 6
580 ± 20
340 ± 49
724 ± 10
308 ± 10
347 ± 28
358 ± 26
385 ± 9
277 ± 14

192 ± 12
200 ± 28
439 ± 24
203 ± 34
393 ± 44
193 ± 29
335 ± 26
457 ± 41
377 ± 30
432 ± 30
259 ± 24
281 ± 25
432 ± 28
356 ± 15
366 ± 22

264 ± 28
269 ± 22
453 ± 25
285 ± 27
397 ± 19
311 ± 43
342 ± 36
491 ± 52
394 ± 24
592 ± 35
315 ± 12
365 ± 26
541 ± 35
382 ± 10
407 ± 54

193 ± 18
188 ± 23
420 ± 10
174 ± 12
401 ± 52
193 ± 23
367 ± 26
438 ± 30
466 ± 45
464 ± 34
248 ± 24
321 ± 14
450 ± 20
347 ± 15
371 ± 33

250 ± 27
240 ± 15
444 ± 35
312 ± 22
408 ± 60
272 ± 51
298 ± 42
505 ± 21
432 ± 38
583 ± 69
318 ± 11
349 ± 28
509 ± 26
385 ± 23
403 ± 19

45 
 

 

Table II-3 (cont’d)

Pharmaceuticals

Sulfadiazine
Sulfamerazine
Sulfamethazine
Sulfamethoxazole
Chlortetracycline
Demeclocycline
Doxycycline
Oxytetracycline
Tetracycline
Acetaminophen
Caffeine
Carbamazepine
Erythromycin
Lincomycin
Tylosin

Spiked
Conc.
(μg/kg)
466
414
491
390
756
507
509
931
759
788
391
420
636
538
484

Pressurized liquid extraction conditions
Temperature (ºC)
50

75

100

300 ± 23
277 ± 14
380 ± 19
243 ± 11
579 ± 42
323 ± 19
266 ± 24
503 ± 52
431 ± 42
648 ± 43
289 ± 31
371 ± 36
394 ± 38
377 ± 38
380 ± 36

342 ± 18
326 ± 33
422 ± 26
248 ± 15
556 ± 20
313 ± 40
328 ± 32
614 ± 51
456 ± 46
571 ± 26
299 ± 17
384 ± 14
474 ± 48
388 ± 35
396 ± 27

338 ± 24
308 ± 29
446 ± 32
266 ± 24
572 ± 48
404 ± 32
348 ± 45
604 ± 85
481 ± 35
653 ± 39
333 ± 26
378 ± 13
498 ± 53
406 ± 39
428 ± 21

46 
 

 

Static extraction with 2 or 3 consecutive cycles was conducted at 100 °C and under 100
bars. Two periods of the static extraction time (15 vs. 25 min) were tested for extraction
efficiency (Table II-3). In general, prolonging extraction time and more extraction cycles achieve
greater extraction efficiency (Bjorklund et al., 1999; Popp et al., 1997). For the static time of 15
min vs. 25 min with the same numbers of extraction cycles, the analysis of the linear-fixed model
results indicated no significant difference (p = 0.897) of extraction efficiencies. This suggests
that 15 min of static period is sufficient for the analytes to approach partition equilibrium
between biosolid and solvent. However, increase of consecutive extraction cycles from 2 to 3
significantly enhanced the extraction efficiencies for all pharmaceuticals tested (p = 0.0002). In
each extraction cycle the induction of fresh solvent into the sample cell reestablishes new
partition equilibrium for the analytes between the added solvent and the biosolid residues. This
partitioning process drives the release of biosolid-sorbed pharmaceuticals to extraction solvent
phase. Taken together, the observation of enhanced extraction efficiencies of 3 consecutive
cycles (versus 2 cycles), and no apparent difference with prolonged static time (25 vs. 15 min),
warrant that the optimal operative condition consists of 3 extraction cycles with 15 min static
time in each cycle.
Increasing temperature in ASE operation resulted in enhanced extraction of the
pharmaceuticals from biosolid (Table II-3). The extraction temperature was tested at 50, 75, and
100 °C with the pressure of 100 bars, 3 extraction cycles and static period of 15 min in each
cycle. The extraction efficiencies were observed to significantly increase with increasing
temperature (p < 0.0001). Therefore, higher temperature is preferred in the ASE operation.
However, considering thermal stability of analytes, the extraction was not conducted at
temperature >100 °C. In this study, the temperature of 100 °C was selected as optimal operating
47 
 

 

condition. The increase of temperature enhances dissolution of pharmaceuticals in extraction
solvent and hence increasing extraction efficiencies. At the same time, the viscosity of extraction
solvent and surface tension on biosolid surfaces decrease as temperature increases, which also
facilitates the solvent to access the domains where the analytes are present. In addition, the
enhanced extraction efficiency could be due to accelerating mass transfer rate from sample
matrices to solvent at increased temperature.
Method validation
The linear ranges of standard curves and correlation coefficients, along with method
LOD and LOQ values, are reported in Table II-4. The standard curves were prepared with two
order of magnitude, and demonstrated an excellent linearity with the correlation coefficients
2

(r ) >0.999. The described method was tested for simultaneous extraction and determination of
15 pharmaceuticals in biosolid, which manifested varying levels of LOD and LOQ. For the class
of sulfonamides the LOD ranged from 0.6 μg/kg (for sulfamethazine) to 15.0 μg/kg (for
sulfadiazine) on dry weight basis of biosolid. This method was less sensitive for tetracycline
antibiotics as shown with the LOD ranging from 4.6 μg/kg (for doxycycline) to 146 μg/kg (for
tetracycline). The overall method recoveries reported were the average of eight replicates of
spiked samples ranging from 49.3% to 94.6%. The standard deviations of the recoveries (n = 8)
were < 10% indicating the reported method achieved satisfactory repeatability. It was noted that
tetracycline antibiotics manifested the relatively lower recoveries ranging from 49% to 68%
(Table II-4). This range of recoveries fell within the low end of the recovery range (i.e., 47.2% to
125 %) reported by Jacobsen et al. (2006) in which tetracyclines were analyzed in swine manure
using the similar analytical approach. However, the extraction recovery of tetracycline (53.8%)

48 
 

 

measured in this study is significantly greater that (~27%) reported by Lillenberg et al. (2009).
The overall recoveries for sulfonamides (from 63.6 % to 94.6%) are comparable with the results
in several previous studies (Diaz-Cruz et al., 2006; Jacobsen and Halling-Sorensen, 2006;
Lillenberg et al., 2009). For instance, sulfamethazine manifested the highest recovery (94.6%) in
this study; the similar results were also reported for extraction of sulfamethazine from swine
manure (84.1% to 97.8%) (Jacobsen and Halling-Sorensen, 2006), and from sewage sludge
(104%) (Diaz-Cruz et al., 2006).

49 
 

 
2

Table II-4. Linear range and correlation coefficient (r ) of standard curves, method limit of detection (LOD), limit of quantification
(LOQ), and recovery

Pharmaceuticals
Sulfadiazine
Sulfamerazine
Sulfamethazine
Sulfamethoxazole
Chlortetracycline
Demeclocycline
Doxycycline
Oxytetracycline
Tetracycline
Acetaminophen
Caffeine
Carbamazepine
Erythromycin
Lincomycin
Tylosin

2

Linear range (μg/L)

r

5.8 - 580
5.2 - 518
6.1 - 614
4.9 - 488
9.4 - 945
6.3 - 634
6.4 - 636
11.6 - 1160
9.5 - 1960
9.8 - 986
4.9 - 488
5.2 - 525
8.0 - 795
6.7 - 673
6.0 - 605

LOD (μg/kg) LOQ (μg/kg) Recovery (%) (n = 8)

0.9992
0.9998
0.9999
0.9997
0.9995
0.9996
0.9993
0.9997
0.9997
0.9994
0.9999
0.9997
0.9993
0.9999
0.9992

15.0
5.5
0.6
1.0
37.2
25.9
4.6
13.7
146
30.7
8.4
2.9
6.6
0.8
5.4

50 
 

50.6
18.2
1.9
3.3
124
86.3
15.5
45.8
488
102
28.3
9.7
21.8
2.7
18.1

63.6 ± 5.2
71.4 ± 5.3
94.6 ± 2.7
77.6 ± 8.0
49.3 ± 5.4
55.0 ± 5.0
68.3 ± 1.1
52.4 ± 5.3
53.8 ± 8.7
84.7 ± 10
79.8 ± 3.8
88.1 ± 3.1
79.0 ± 9.8
83.6 ± 9.9
77.3 ± 5.4

 

The manifestation of relatively low recoveries (e.g. ~50%) of the overall method for a
certain class of pharmaceuticals such as tetracyclines could be attributed to inefficient extraction
of ASE operation, analyte loss during SPE cleanup, and/or matrix effects on mass spectrometer
responses from extracts. The loss of target analytes during SPE cleanup step was < 11 %
(describe below). To examine sample matrix effects on mass spectrometer responses, standard
addition method was used in which the analytes were spiked into biosolid extracts after ASE
treatment and SPE cleanup. The spiked pharmaceutical concentrations were tested at two levels:
10-23 μg/L and 100-232 μg/L. The ratios of measured concentrations to the corresponding
spiked concentrations ranged from 0.90 to 1.15, indicating matrix effects caused a minimal
impact to the overall method recovery. Therefore, the low recoveries measured in this study
result primarily from inefficient extraction of pharmaceuticals from ASE procedure, particularly
for the class of tetracycline antibiotics which usually demonstrate a great affinity with sorbents.
SPE cleanup
In order to evaluate the extent of pharmaceutical loss during SPE cleanup step, the target
compounds were spiked into biosolid extracts after ASE treatment. The solution passed through
the preconditioned HLB cartridge, followed by an elution with methanol-water mixture (v/v =
1:1) containing 150 mg/L EDTA and analyzed by LC-MS/MS. Two concentration levels (10-23
μg/L and 98-232 μg/L) of pharmaceuticals were spiked into the biosolid extracts (i.e., after ASE
treatment). The measured pharmaceutical concentrations were subtracted by the amount
pharmaceuticals originally present in the samples, and compared to the spiked concentrations
(Table II-5). For most pharmaceuticals, the recoveries of spiked amount ranged from 86.8 to
120%, indicating that the loss of analytes during SPE cleanup step was minimal. Demeclocycline
and lincomycin manifested relatively higher recovery (i.e., 155 % for demeclocycline at 12.7
51 
 

 

μg/L, and 176% for lincomycin at 134.6 μg/L), which was attributed to the experimental
variations since reasonable recoveries were obtained for these two compounds at another
concentration. Therefore, the relatively low recoveries of the overall method for some analytes
(e.g., tetracyclines) could be attributed to the inefficient extraction from biosolids during the
ASE step.

52 
 

 

Table II-5. Recoveries of solid-phase extraction for pharmaceuticals spiked into biosolid extracts after accelerating solvent extraction

Pharmaceuticals
Sulfadiazine
Sulfamerazine
Sulfamethazine
Sulfamethoxazole
Chlortetracycline
Demeclocycline
Doxycycline
Oxytetracycline
Tetracycline
Acetaminophen
Caffeine
Carbamazepine
Erythromycin
Lincomycin
Tylosin

Spiked concentration
(µg/L)
11.6
10.4
12.3
9.8
18.9
12.7
18.9
23.3
12.7
19.7
9.8
10.5
15.9
13.5
12.1

Recovery (%)
(n = 3)
110.9 ± 9.8
86.8 ± 6.5
91.0 ± 8.5
115.2 ± 8.6
95.6 ± 7.5
155.5 ± 16.5
116.9 ± 9.1
98.4 ± 6.1
108.5 ± 12.8
102.1 ± 8.2
103.7 ± 4.6
100.5 ± 8.5
112.4 ± 11.1
129.2 ± 13.4
98.2 ± 7.5

53 
 

Spiked concentration
(µg/L)
116.4
103.5
122.8
97.5
189.0
126.8
189.8
232.7
127.2
197.1
97.7
105.0
159.0
134.6
120.9

Recovery (%)
(n = 3)
102.6 ± 8.0
102.5 ± 8.3
100.9 ± 6.4
97.7 ± 7.1
89.1 ± 9.4
118.1 ± 14.2
111.4 ± 11.3
121.1 ± 10.9
119.4 ± 9.0
113.5 ± 8.9
110.0 ± 5.2
101.4 ± 8.0
107.3 ± 8.1
176.5 ± 18.9
96.8 ± 8.8

 

Application to biosolid samples
Biosolid samples were collected at WWTPs of six cities (Lansing, East Lansing, St Clair,
Plainwell, Traverse City, and Imlay City) in Michigan, USA. These samples were extracted
using the optimized method described above. The samples were conducted in triplicates. The
measured concentrations and the corresponding standard deviations (n = 3) are presented in
Table II-6. Among these samples, 14 out of the 15 target pharmaceuticals were found except
tylosin. The average concentrations of the pharmaceuticals manifested a relatively wide range
from 2.6 to 743.6 μg/kg (Table II-6). The calculated standard deviations for all the measurements
were < 17%, indicating a reasonable repeatability of the analysis. Tetracycline antibiotics were
frequently detected with the concentrations ranging from 36.6 to 743.6 μg/kg. Doxycycline was
detected in all 6 biosolid samples with concentrations ranging from 159.9 to 292.4 μg/kg.
Relatively higher concentrations of tetracycline (281.9 μg/kg) and chlortetracycline (346.6 μg/kg)
were found in the sample of Lansing WWTP compared to the samples collected from other five
WWTPs. Oxytetracycline was detected in four biosolid samples from 51.9 to 743.6 μg/kg. Trace
amounts of demeclocycline were found in three biosolid samples (i.e., 36.6, 98.2 and 131.2
μg/kg) from WWTPs in Plainwell, East Lansing, and St. Clair. Sulfonamides were also
frequently detected, and the corresponding concentrations ranged from 4.8 (sulfamethoxazole) to
668.9 μg/kg (sulfamerazine). For other six pharmaceuticals (acetaminophen, carbamazepine,
caffeine, erythromycin, lincomycin and tylosin), the measured concentrations spanned from 2.6
to 370.4 μg/kg. Caffeine and carbamazepine, both of which are solely used by humans, were
detected in the biosolids from five cities except Imlay City. Interestingly, Imlay City has the least
populations among the six cities; this might be the reason responsible for the least occurrence of
human medicine in the biosolid of Imlay City WWPT (Peeler et al., 2006). Tylosin is commonly
54 
 

 

used as veterinary pharmaceutical to control livestock infection and promote livestock growth. It
is reasonable that no samples were found to contain tylosin in municipal WWTP biosolids.
Acetaminophen was detected in three biosolid samples from Lansing, East Lansing and St. Clair
with concentrations of 88.6 to 370.4 µg/kg. In fact, these three cities have more populations than
Plainwell, Traverse City, and Imlay City.

55 
 

 

Table II-6. Measured pharmaceutical concentrations and standard deviations (µg/kg dry matter) in the collected biosolid samples

Sampled wastewater treatment plants
Pharmaceuticals
Lansing
a

East Lansing

St. Clair

Plainwell

Traverse City

Imlay City
nd
nd
131.8 ± 13.0
nd
b
69.6 ± 7.6
nd
159.9 ± 19.1
174.2 ± 4.6
nd
nd
nd
nd
nd
2.6 ± 0.2
nd

Sulfadiazine
Sulfamerazine
Sulfamethazine
Sulfamethoxazole
Chlortetracycline

nd
112.0 ± 6.3
nd
4.8 ± 0.5
346.6 ± 22.1

nd
668.9 ± 45.5
127.8 ± 15.8
26.1 ± 3.0
nd

nd
nd
124.8 ± 11.8
8.8 ± 1.2
nd

nd
nd
nd
35.9 ± 4.4
nd

Demeclocycline
Doxycycline
Oxytetracycline
Tetracycline

nd
291.2 ± 22.0
nd
b
281.9 ± 42.3
370.4 ± 12.9
47.9 ± 5.2
b
6.2 ± 1.2
62.8 ± 5.8
nd
nd

98.2 ± 6.2
224.3 ± 30.8
51.9 ± 4.4
nd

131.2 ± 11.3
149.6 ± 11.0
nd
nd

36.6 ± 4.0
292.4 ± 48.5
743.6 ± 21.7
nd

562.2 ± 61.4
nd
nd
nd
b
90.2 ± 10.0
nd
234.0 ± 26.8
201.4 ± 17.8
nd

101.1 ± 4.7
47.1 ± 6.1
b
4.6 ± 0.5
nd
6.2 ± 1.1
nd

88.6 ± 7.6
46.9 ± 3.7
b
5.1 ± 0.3
10.4 ± 0.8
8.7 ± 1.4
nd

nd
75.5 ± 4.1
16.4 ± 0.9
16.8 ± 1.3
nd
nd

nd
33.9 ± 3.7
22.3 ± 5.9
nd
6.5 ± 2.0
nd

Acetaminophen
Caffeine
Carbamazepine
Erythromycin
Lincomycin
Tylosin
a

b

b

Not detected. Detected but less than limit of quantification.

56 
 

b

 

Conclusions
This study describes an efficient and reproducible analytical approach for simultaneous
determination of multiple classes of pharmaceuticals in biosolids. The experimental steps
consisted of accelerated solvent extraction of freeze-dry biosolid sample, and cleanup by solidphase extraction followed by simultaneous analysis by liquid chromatography/tandem mass
spectrometry. The extraction parameters of ASE including extraction solvent, temperature,
pressure, static period and extraction cycle were optimized to achieve satisfactory extraction
efficiency. Overall, this method offers less laborious work, automatic extraction procedure, less
consumption of organic solvent and relatively high extraction efficiency. The method has been
tested successfully for simultaneously determining multiple classes of pharmaceuticals in
WWTP biosolids collected from different locations. The results also suggest that many
pharmaceuticals could survive wastewater treatment processes and accumulate in solid phase
(e.g., biosolids). Land application of the contaminated biosolids could result in the dissemination
of pharmaceuticals in soil and water environment. Chronic exposure to low levels of antibiotics
might exert selection pressure on development of antibiotic resistant bacterial strains in the
environment.

57 
 

 

REFERENCE
 

 

58 
 

 

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63 
 

 

CHAPTER III
TETRACYCLINE SORPTION BY K- AND Ca-SMECTITE CLAYS FROM WATER

64 
 

 

Abstract
Tetracyclines are a class of antimicrobials that have been extensively used to treat
diseases and improve health for humans and animals. A large portion of tetracyclines used on
this purpose are released into the environment, raising concerns on the potential risks to
ecosystems and environmental quality. Sorption by soils is a determinant controlling transport,
fate and hence exposure in the environment. Soil mineral fractions are a dominant sorptive
+

2+

domain to retain tetracyclines. In this study, sorption of tetracycline by K - and Ca -saturated
smectite clays was measured using a batch equilibration method. Sorption manifested nonlinear
+

2+

isotherms, and was strongly suppressed by high levels of inorganic cations (e.g., K and Ca ) in
solution. Sorption of tetracycline by K-, and Ca-smectite was concomitant with the release of K
and Ca

2+

from the clays, indicating that cation exchange process is the primary driving force for

sorption of tetracycline. In addition, tetracycline sorption to smectite concurred with
consumption of protons from solution. X-ray diffraction patterns revealed that the interlayer
distances of smectite clays increased as tetracycline sorption increased, and gradually
approached a constant basal spacing of 17.5 to 18 Å. These results indicate that the clay-sorbed
tetracyclines are intercalated in interlayer regions.

65 
 

+

 

Introduction
Tetracyclines are a class of antibiotics that have been widely administered to human and
veterinary treatments to maintain health and prevent diseases since they were first discovered
from 1940s. In 2005, it was estimated approximately 3.8 million kilograms of tetracyclines were
used to animal treatments and livestock production in the United States; the amount increased
about 30% compared to that used in 2004 (Phillips, 2006). Halling-Sorensen et al. (1998)
reported that 60 to 90% of tetracyclines administered to animal feeding operations could pass
through the digestion system, and excreted into liquid and solid manures. Tetracyclines released
from human treatments could survive residential wastewater treatment processes, and
disseminate in the environment (Karthikeyan and Meyer, 2006; McArdell et al., 2003; Miao et
al., 2004). The short-term lagoon storage (e.g., ~3 months) could remove certain fractions of
tetracyclines in animal manures; however, the subsequent land application of animal manures for
their fertilizer values results in the contamination of agricultural land and water with the residual
tetracyclines (Eichhorn and Aga, 2004; Smith et al., 2004). Kolpin et al. (2002) reported that
tetracycline, chlortetracycline and oxytetracycline were found in stream waters collected from 30
states in the U.S. with the detection frequency of 1.2, 2.4 and 1.2%, respectively, and the
corresponding maximum measured concentrations were 0.11, 0.69 and 0.34 μg/L. More than
fifty-year utilization of tetracyclines and the resultant widespread distributions in environmental
media have exerted chronic and selective pressures on the development of antibiotic resistance
(Davies, 1994; Witte, 1998; Zhang et al., 2009). Sorption and fate of tetracyclines in the
environment is a crucial factor in assessing their exposure and risks in the ecosystems.
Tetracycline contains four fused six-carbon cyclic rings to which amide,
tricarbonylmethane, diketone and dimethylammonium functional groups are attached (Figure III-

66 
 

 

1). These functional groups are ionized in aqueous solution resulting in cation (+ 0 0), zwitterion
(+ - 0), and anion (+ - - , 0 - -) species with the fraction distributions depending on solution pH
values (Figure III-1). These species and formed complexes with commonly water-soluble cations
2+

(e.g., Ca

2+

and Mg ) demonstrated varying steric configuration in aqueous solution (Othersen

et al., 2006; Wessels et al., 1998). For instance, in acidic and neutral aqueous solutions
dimethylamino functional group is protonated locating above the BCD ring system to reduce the
steric crowding with the hydroxyl group OH associated with C12a. In alkaline aqueous phase or
nonaqueous solution the dimethylamino functional group is located below the BCD ring planar
structure (Gulbis and Everett, 1976; Mitscher et al., 1968; Mitscher et al., 1972). The structure of
tetracycline and dissociation to form multiple species could lead to multiple interaction
mechanisms with soil or soil components such as replacement with cations affiliating with
surfaces, complexation with multivalent cations on surfaces, and possible partitioning into soil
organic matter.

67 
 

 

A
HO

CH3

7
D

C
11

9
OH

O
pKa2=7.7
diketon

B

H(TC)-

H2(TC)

H3(TC)+

100
Speciation fraction (%)

H

HN(CH 3)2
4
OH
5
3
B
A
2
12
1
C NH2
pKa1=3.3
OH
OH
O
O
tricarbonyl methane

6

10

8

pKa3=9.7
dimethylammonium

TC2-

80
60
40
20
0
0

2

4

6

pH

8

10

12

Figure III-1. (A) Tetracycline structure, and (B) speciation distribution as a function of solution
+
2pH. H3(TC) , H2(TC), H(TC) , and TC represent the cation, zwitterion, anion, and divalent
anion of tetracycline species.

68 
 

 

Sorption by soils/sediments is a major process influencing transport, fate and
bioavailability of tetracycline in the environment. In soils both soil organic matter and clay
mineral fractions are the active domains for sorption of tetracyclines from water solution
(Figueroa et al., 2004; Gu et al., 2007; Kay et al., 2004; Kulshrestha et al., 2004; MacKay and
Canterbury, 2005; ter Laak et al., 2006). Tetracyclines are adsorbed by clay minerals via
exchanging the inorganic cations on clay surfaces and/or complexing with the exchanged
multivalent cations associated with clays (Loke et al., 2002; Sithole and Guy, 1987a; Sithole and
Guy, 1987b; Wessels et al., 1998). Tetracycline manifest a greater sorption to Namontmorillonite in solution at pH 1.5 and 5.0 than that in alkaline solutions, and no apparent
difference was observed for sorption at pH 1.5 and 5.0 (Kulshrestha et al., 2004). Hence, cation
exchange reaction is believed to be the dominant mechanism for tetracycline sorption by soils
and minerals (Figueroa et al., 2004; Kulshrestha et al., 2004; Sassman and Lee, 2005). When pH >
7, tetracycline anion is the predominant species in solution, which could complex with metal ions
on clay surfaces (Aristilde et al., 2010; Figueroa and Mackay, 2005; Pils and Laird, 2007).
Depending on types of ligands and ratio of metal to ligand, stoichiometric 1:1, 1:2 and 2:1 metaltetracycline complexes could be formed in solution phase (Berthon et al., 1983; Lambs et al.,
1988; Martin, 1979; Newman and Frank, 1976; Schmitt and Schneider, 2000; Wessels et al.,
2+

1998). Tetracycline can chelate with metal ions at several sites. For example, Ca
tetracycline via O12-O1 and O10-O11, whereas Mg

2+

chelate with

chelate with tetracycline via O11-O12 and

O3-N4 to form Mg-tetracycline complexes (Wessels et al., 1998). Soil organic matter and
aqueous dissolved organic matter are other major domains for interacting with tetracycline in the
environment, and the corresponding mechanisms involve cation exchange, hydrogen bonding
and metal-tetracycline complexation (Gu et al., 2007; MacKay and Canterbury, 2005; McCarthy
69 
 

 

and Zachara, 1989; Pils and Laird, 2007; Sithole and Guy, 1987a).
The objective of this study was to improve the understandings of tetracycline sorption
mechanism by smectites from aqueous phase. Cation exchange process (i.e. replacement of
inorganic cations associated with clay minerals by tetracycline) is hypothesized to be the primary
driving force for sorption of tetracycline by clay minerals. To validate this hypothesis,
tetracycline sorption from water was measured for smectite with the negatively charged sites
+

2+

compensated by K and Ca . Meanwhile, the release of exchanged inorganic cation was also
determined to demonstrate the cation exchange process. In addition, molecular dynamic
simulation and X-ray diffractions of tetracycline-sorbed clay films were obtained to provide
further insight into the tetracycline-clay interaction mechanisms.

70 
 

 

Experiment Section
Smectite Clays and Chemicals. A reference smectite clay (SWy-2) was obtained from
the Source Clays Repository of the Clay Minerals Society. The clay has a cation exchange
2

capacity of 800 mmol/kg, and the theoretical surface area of 800 m /g. The < 2 μm-sized clay
fractions were collected using wet sedimentation method. The cation exchange sites on smectite
+

surfaces were saturated with K and Ca

2+

separately by washing with 0.5 M KCl or CaCl2

solution three times. The clay suspensions were subsequently washed using Milli-Q water five
-

times to remove the excess KCl and CaCl2 salts. The removal of Cl was determined by the
negative test result of AgNO3 solution. The clay suspensions were then quickly frozen, and
freeze dried to receive clay powders.
Tetracycline hydrochloride (purity > 95%) was purchased from Sigma-Aldrich Chemical
Company (St. Louis, MO), and was used as received. Sodium acetate, sodium hydroxide,
ammonium acetate, heptafluorobutyric acid, KCl, CaCl2, methanol, sodium
ethylenediaminetetraacetate (EDTA), formic acid, and sulfuric acid were purchased from J.T.
Baker (Phillipsburg, NJ, USA).
Sorption Isotherm Measurements. Tetracycline sorption to K- and Ca-smectites from
aqueous solution was measured using a batch equilibration method. A series of tetracycline
solutions were prepared with concentrations ranging from 0 to 0.5 μmol/mL in KCl and CaCl2
solutions. The background KCl and CaCl2 solutions were prepared at the concentrations of 0.0,
0.01, 0.10 and 0.80 M. A certain amount of clay was weighed in glass centrifuge tubes, and
mixed with a known volume of tetracycline solutions. The tubes were placed on a rotatory shaker
at 40 rpm for 18 h at room temperature (~23 ºC). The preliminary study indicated that sorption
71 
 

 

reached equilibration within 12 h. The tubes were subsequently centrifuged at 3000 g for 20 min
at 23 ºC. The supernatant pH was measured using an Accumet AB15 pH meter (Fisher Scientific,
Pittsburgh, PA, USA) equipped with a combination electrode. Experimental controls consisted of
the initial tetracycline solutions in the absence of clay. Within the sorption equilibration period
(18 h), the loss of tetracycline in these controls was < 5%. Therefore, the sorbed concentration
was calculated from the difference between the concentration of initial tetracycline control
solution (without clay) after 18 h and the amount remained in aqueous solution after sorption by
clay at equilibration. All experimental samples were prepared in duplicate.
An aliquot of supernatant was collected and analyzed using a Perkin-Elmer highperformance liquid chromatography (HPLC, Waltham, MA, USA) for tetracycline
concentrations ranging from 0.004 to 0.35 μmol/mL or a Shimadzu HPLC (Columbia, MD, USA)
fully integrated with an Applied Biosystems Sciex 3200 triple quadrupole mass spectrometer
-4

(MS/MS, Foster City, CA, USA) for tetracycline concentrations from 2 × 10 to 0.004 μmol/mL.
The Perkin-Elmer HPLC system consisted of a Series 200 pump, a Series 200 UV/vis detector, a
+

Series 200 autosampler, and a Supelco ABZ column (15cm × 4.6mm, 5 μm). The wavelength of
the UV/vis detector was 370 nm. Isocratic binary mobile phase (A:B = 40:60) was used with a
flow rate of 1.0 mL/min. Mobile phase A was methanol, and phase B was a aqueous solution (pH
= 6.5) containing 50mM sodium acetate, 25 mM calcium chloride, 12.5 mM EDTA and 2 M
sodium hydroxide. The samples of lower tetracycline concentrations were analyzed by LCMS/MS. A binary mobile phase was used, and consisted of phase A water and methanol (v/v =
95/5) and phase B methanol and water (v/v = 95/5) both of which contained 2 mM ammonium
acetate and 20 mM heptafluorobutyric acid. A Thermo Hypersil Gold column (50 × 2.1 mm,
particle size 5 µm) was used with the flow rate of 0.15 mL/min. Tetracycline concentration was
72 
 

 

quantified using multiple reaction monitoring mode with the precursor/product transition of
445.0/410.0.
+

2+

Release of Exchanged Cation. A certain amount of exchanged cation (i.e., K and Ca )
associated with smectite surfaces was observed to release into solution during tetracycline
sorption. The released cation concentration in aqueous solution was measured only for the
experimental setting of tetracycline sorption by smectite without adding the electrolyte in the
+

background solution. K and Ca

2+

concentrations ware measured using a Perkin-Elmer

AAnalyst 400 Atomic Absorption Spectrophotometer (AAS, Branford, CT, USA). To do so, a
+

portion of supernatant (~10 mL) was collected after centrifugation and measured for K and
2+

+

2+

Ca . The measured concentrations was subtracted by the amounts of K or Ca

released from

K- or Ca-clays devoid of tetracycline, and normalized on the basis of clay weight. This
normalized concentration is referred to as the released exchanged concentration (μmol/g) from
clay.
X-ray Diffraction Measurement. After collecting the supernatant for the measurement
of tetracycline concentration and the amount of released cation, the remaining supernatant was
removed until ~1 to 2 mL of clay residue remained in the tubes. The remaining clay slurry was
re-suspended using a vortex mixer, dropped on glass slide, and air dried to obtain oriented clay
film. The prepared clay films were analyzed using a Philips APD 3720 automated X-ray
diffractiometer (Philips Electronic Instrument Co., Mahwah, NJ) equipped with Cu-Kα radiation,
an APD 3521 goniometer and a diffracted-beam monochromator. The scanning angle (2θ) ranged
from 2 to 14º at step of 0.02º, and the scanning time was 2s per step.
Molecular Dynamic Simulations. The interaction of tetracycline with smectite in
73 
 

 

interlayer was conducted using molecular dynamic simulations. A model was created for Casmectite (SWy-2) with the composition of Ca4(Al42Mg6)(Si95Al1)O240(OH)48Cl and the
corresponding CEC of 790 mmol/kg. Tetracycline was loaded between clay interlayers with the
amount equivalent to the concentration of 104 µmol tetracycline per gram of clay. The charge
distribution of the added tetracycline was calculated using CHELPG method (Breneman and
Wiberg, 1990), and structural configuration was optimized using HF/6-31G method (M. J. Frisch,
2003). Water molecules were randomly added into clay interlayers to maintain the clay basal
spacing at 14.9 Å that was experimentally measured. The overall system energies were computed
using force filed developed specifically for clay minerals (Teppen et al., 1997) combined with
the pcff force field for tetracycline and water molecules (Maple et al., 1994). These molecular
dynamics simulations were run in NPT ensemble for 0.1 ns, which is sufficient for the system to
reach volume and energy equilibration.

74 
 

 

Results and Discussion
Cation Exchange. Figure III-2 shows tetracycline sorption to K- and Ca-smectites from
the solution without adding KCl or CaCl2 background electrolytes, and the corresponding release
+

2+

of exchanged K and Ca

from the minerals. It is showed that the amount of the exchanged

cations released into solution from clay increased with increasing tetracycline sorption (Figure
III-2). This clearly indicates that tetracycline replaces the inorganic cation associated with
n+

mineral surfaces, that is, cation exchange process: tetracycline + M -clay ↔ tetracyline-clay +
n+

M

in which M

n+

2+

= Ca

+

or K . For K-smectite clay, the amount of tetracycline sorption is

equivalent to that of K+ released from clay (Figure III-2A).
The similar trend of tetracycline sorption vs. the release of exchanged cation was also
observed for tetracycline sorption on Ca-clay, but a relatively less amount of exchanged Ca

2+

was released compared to that magnitude tetracycline sorption (Figure III-2B). It is unlikely to
occur that sorption of two tetracycline molecules results in release of one Ca

2+

due to the

significant distinction in size for organic vs. inorganic cations. At the lower tetracycline sorbed
concentrations (<350 μmol/g), the apparent distinction between the amount of tetracycline
sorption vs. the release of Ca

2+

was observed (Figure III-2B). In addition, Ca-smectite

demonstrated a relatively stronger affinity with tetracycline than K-clay with the aqueous
concentration <0.10 μmol/mL. For instance, Ca-smectite approached the sorption of about 200
μmol/g at aqueous tetracycline concentration of 0.01 μmol/mL, whereas K-smectite reached the
same level of sorption at aqueous tetracycline concentration of 0.04 μmol/mL. Taken together, in
addition to cation exchange reaction, the enhanced sorption by Ca-smectite is plausibly attributed
75 
 

 
2+

to formation of cation-organic complexes between tetracycline and Ca

associated with mineral

surfaces (described below).

500

500

A: K-Smectite
400

300

300

200

200

100

100

tetracycline
exchanged cation
0
0.00

0.05

0.10

0.15

0
0.25

0.20

500

500

B: Ca-Smectite
400

400

300

300

200

200

100

Released exchanged cation from clay (μmol/g)

Tetracycline sorbed concentration (μmol/g)

400

100

0
0.00

0.02

0.04

0.06

0.08

0.10

0.12

0.14

0
0.16

Aqueous tetracycline equilibrium concentration (μmol/mL)

Figure III-2. Relationship between tetracycline sorption and exchanged cation released from
clays.

76 
 

 

Effects of Competitive Inorganic Cations. Figure III-3 shows tetracycline sorption
isotherms to K-smectite from the aqueous KCl solution and to Ca-smectite from CaCl2 solution
with concentrations of 0.0, 0.01, 0.10 and 0.80 M, along with sorption from water without added
electrolyte for comparison. The sorption isotherms were essentially nonliner for both K- and Casmectite. The presence of inorganic cations suppressed sorption of tetracycline by K- or Casmectite clays, and the greater suppression was observed with increasing inorganic cation
concentrations in solution. For the instance of K-smectite, at aqueous tetracycline concentration
of 0.20 μmol/mL tetracycline sorption reduced from 432 μmol/g in solution without the added
KCl down to 263, 153 and 36 μmol/g in 0.01 M, 0.10 M and 0.80 M KCl solution, respectively.
For Ca-clay similar suppressive effects was also observed; at aqueous tetracycline concentration
of 0.15 μmol/mL sorption decreased from 440 μmol/g in the solution without the added CaCl2 to
390, 250 and 79 μmol/g in 0.01, 0.10 and 0.80 M CaCl2 solution. In the experiments conducted,
+

2+

the molar concentrations of added inorganic cations (i.e. K and Ca ) were 50 to 80000 times
the tetracycline molar concentrations in aqueous solution (Figure III-3). Such vast amount of
inorganic cation present in aqueous phase effectively competed with tetracycline for the
negatively charged sites on mineral surfaces, resulting in a shift of cation exchange equilibration
toward less sorption of tetracycline. It is noted that Ca-smectite also manifested a greater
sorption for tetracycline compared to K-smectite in the presence of the same level of CaCl2 and
KCl. The view of cation exchange reaction would predict that, a priori, more reduced sorption by
2+

Ca-smectite (vs. K-smectite) since divalent Ca

is less favorable for release from mineral
+

surfaces, and is more competitive for cation exchange sites compared to monovalent K .

77 
 

 

However, the fact that greater tetracycline sorption was observed for Ca-smectite relative to Ksmectite suggests that the formation of Ca-tetracycline complexes on mineral surfaces enhances
tetracycline sorption by Ca-smectite.

78 
 

 

600

25
20
15
10
5
0
0.000

K-Smectite

0.008

0.016
6.5

Tetracycline sorbed concentration (μmol/g)

400

0.00 M
0.01 M
0.10 M
0.80 M

5.4

7.4
7.9

4.6

7.6
200

600

5.8

8.3
8.4

0

5.7

6.7

7.0

80
60
40
20
0
0.000

400

6.5

6.5

5.4
6.0

6.9

4.4

5.9

Ca-Smectite

0.003
6.1

7.3

0.006

5.5
4.5

5.4

5.6
4.6
6.3
5.5
5.8
7.0
6.0
4.7
6.5
5.3
6.3
7.2 7.0

4.1

7.8
200

0
0.00

4.5

4.7

5.6

0.05

0.10

0.15

3.8

4.4

0.20

0.25

0.30

Aqueous tetracycline equilibrium concentration (μmol/mL)

 

Figure III-3. Tetracycline sorption isotherms for (A) K-smectite from KCl solution and (B) Casmectite from CaCl2 solution with the background electrolyte concentrations of 0.0, 0.01, 0.10

79 
 

 

and 0.80 M. The pH values measured after sorption equilibrium are marked close to the
corresponding sorption data.
pH Changes during Tetracycline Sorption. In this study, no buffer was added into the
solution to control pH in order to exclude the possible disturbances from buffer species on
tetracycline sorption. The pH values after tetracycline sorption equilibration (labeled in Figure
III-3) ranged from 4.5 to 8.4 for K-smectite in which tetracycline primarily manifested zwitterion
and anions. For Ca-smectite the pH values ranged from 3.8 to 7.8, and the associated species in
aqueous solution were cation, zwitterion, anions and Ca-tetracycline complexes. For the initial
tetracycline solution, the pH values ranged from 3.5 to 8.5 corresponding to tetracycline
concentrations from 0.34 to 0.01 μmol/mL. As tetracycline was sorbed by the smectite clay, pH
values increased from minimal at very lower sorption (< 20 μmol/g) to ~1.5 pH unit at higher
sorption (> 250 μmol/g). We estimated the reduction of aqueous tetracycline concentration due to
sorption, which is plotted against the corresponding consumption of proton concentration (Figure
III-4). It is noted that increased tetracycline sorption (i.e. greater decrease of aqueous tetracycline
concentration) concurred with increased consumption of protons, which is consistent with the
fact of increase of pH values after tetracycline sorption.

80 
 

 

0

A: K-Smectite
-1

Logarithm of proton consumption in solution (μmol/mL)

-2
-3

0.00 M
0.01 M
0.10 M
0.80 M

-4
-5
-6
-7
0

B: Ca-Smectite

-1
-2
-3
-4
-5
-6
-7
-3.0

-2.5

-2.0

-1.5

-1.0

-0.5

Logarithm of reduction of aqueous tetracycline concentration (μmol/mL)

Figure III-4. Relationship between reduction of tetracycline concentration in aqueous phase due
to sorption by smectite and the consumption of proton in solution. (A) Sorption by K-smectite in
the presence of 0.0, 0.01, 0.10, and 0.80 M KCl, and (B) sorption by Ca-smectite in the presence
of 0.0, 0.01, 0.10, and 0.80 M CaCl2.

81 
 

 

At lower pH (e.g. < 4.7), tetracycline primarily manifested cationic and zwitterionic
species in solution. Cationic tetracycline sorbed onto clay surfaces in concomitant with
replacement of exchangeable cations on mineral surfaces. To maintain tetracycline speciation
equilibrations in solution, other species such as zwitterion and anions are protonated and shift to
cationic species in solution, which consumes proton resulting in increasing solution pH. In
addition, zwitterion could be also sorbed by clay through cation exchange reaction, but needs to
accompany with proton to balance the negatively charged moiety in tetracycline. Sorption of
both tetracycline species consume the protons in aqueous solution leading to an increasing pH.
As pH continued to increase, the major species of tetracycline shifted to zwitterions and anions.
2+

For CaCl2 solution, tetracycline zwitterion and anions as well as the formed Ca -tetracycline
complexes were present in the solution with the distributions of each fraction dependent of pH,
2+

tetracycline concentration and the amount of Ca . Recall that sorption of tetracycline zwitterion
via cation exchange reaction consumes the proton from solution to compensate the negatively
charged functional moiety of tetracycline molecule as indicated by an increased pH (Figure III-4).
Ca-tetracycline complexes usually form at relatively higher pH values in the presence of
sufficient amount of CaCl2. For example, at pH 7 approximately 43% of tetracycline formed
2+

complexes with Ca , 46% and 10% of tetracycline were zwitterions and anion, respectively, in
0.01 M CaCl2. As CaCl2 concentration increased, Ca-tetracycline fractions increased to 95% in
0.10 M CaCl2 solution, and to 99% in 0.80 M CaCl2. Meanwhile, tetracycline sorption was
observed to significantly decrease (Figure III-3). The formed Ca-tetracycline complexes
manifested a lower affinity for mineral surfaces compared to tetracycline zwitterion due

82 
 

 

plausibly to the larger sized metal-tetracycline complexes inhibited the intercalation in smectite
interlayers.
It is also noted from Figure III-4 that the presence of KCl and CaCl2 consumed more
protons for tetracycline to approach similar sorption levels (i.e., similar reduction of aqueous
tetracycline concentrations) compared to sorption from water. The consumed amount of proton
increased with increasing electrolyte concentration. In aqueous solution, inorganic cations
approximate the negatively charged functional groups of tetracycline. Protons must compete with
+

2+

inorganic cations (K or Ca

in this study) for these negative charges to convert the species

facilitated for sorption. Therefore, more protons are needed to be consumed for achieving the
similar sorption levels in the solutions containing a higher electrolyte concentration (Figure III-4).
Complexation of Tetracycline with Ca

2+

on Mineral Surfaces. Many previous studies
2+

have indicated that anionic tetracycline can complex with Ca

in aqueous phase (Schmitt and

Schneider, 2000; Tongaree et al., 1999; Wessels et al., 1998), and minimally complex with
monovalent cation such as K+ (Coibion and Laszlo, 1979). To illustrate the interactions of
tetracycline with Ca-smectite, molecular dynamic simulation was used with tetracycline sorption
at 104 µmol per gram of Ca-clay which is within the range of our measured sorption. After the
computing system reached the static state, tetracycline lay between smectite interlayer with BCD
ring system parallel with siloxane surfaces (Figure III-5A). Positively charged
dimethylammonium group is positioned towards tetracycline planar structure in the extended
conformation (Othersen et al., 2006). Comparing with the twisted conformation in which
dimethylammonium group is located above the BCD ring system, tetracycline adopts the
extended conformation that could be more ready for intercalation into smectite interlayers. The
83 
 

 
2+

enlarged view of tetracycline and Ca

interaction is shown in Figure III-5B. The simulation

results indicated that O11 and O12 of tetracycline approached to Ca
suggesting tetracycline forms inner-sphere complexes with Ca

2+

2+

with a distance of 2.2 Å,

in smectite interlayers.

To experimentally verify the possible formation of tetracycline-metal cation complexes
on mineral surfaces, we measured sorption by Cu- and Fe-smectite clays, and compared the
sorption with those by K- and Ca-smectite. The results revealed that sorption by Cu- and Fe2+

smectite clays was much greater than that by Ca and K-clay. It is known that Cu
2+

could form a much stronger complexation with tetracycline than Ca

3+

and Fe

(Gu and Karthikeyan,

2005; Pei et al., 2010). The magnitude of sorption order implies that tetracycline forms
2+

complexes with Ca , but the interaction is weaker than those with Cu

84 
 

2+

and Fe

3+

in interlayers.

 

 

Figure III-5. Molecular dynamic simulations of tetracycline sorption by Ca-smectite. Elements
are labeled with color: C (grey), H (white), N (blue), O (red), Ca (green), K (purple), Al (pink),
Si (yellow). (A) Tetracycline molecule sits between smectite interlayer with BCD ring system
2+

parallel with siloxane sheet. (B) Enlarged view of tetracycline interaction with exchanged Ca
on mineral surfaces. The distance of 2.2 Å (solid arrows) between Ca and O11, O12 on
2+

tetracycline structure indicates tetracycline forms inner-sphere complex with Ca . (For
interpretation of the references to color in this and all other figures, the reader is referred to the
electronic version of this dissertation.)
85 
 

 

Intercalation of Tetracycline in Smectite Interlayers. X-ray diffraction pattern for Ksmectite at several tetracycline sorption levels is presented in Figure III-6A. A diffraction peak
was observed at ~7.9º (2θ) for K-smectites without or with low tetracycline sorption (< 56
μmol/g), indicating that smectite platelets maintain relatively parallel structures with the basal
spacing of 11.2 Å. As the tetracycline loading increased, the diffraction peak at 7.9º (2θ)
gradually became broader with reduced intensity indicating less ordering of K-clay layers. As the
sorption reached > 218.4 μmol/g, the diffraction peak at 7.9º became very weak, whereas an
additional peak at ~5.0º appeared corresponding to a basal spacing of 17 Å. The similar changes
of XRD patterns were also observed for Ca-smectite with varying tetracycline loadings; the
diffraction peak shifted from original 6º to 5.2º corresponding to the basal spacing expanding
from 14.7 to 17.1 Å (Figure III-6B). For Cu- and Fe-clays, the basal spacings expanded to ~18 Å
with tetracycline sorption > 250 μmol/g. These results clearly establish that tetracycline
molecules enter interlayers when sorbed by smectite.

86 
 

 

A
300.1 mmol/g

Relative intensity

218.4 mmol/g
169.5 mmol/g
135.1 mmol/g
76.8 mmol/g
55.9 mmol/g
26.7 mmol/g
19.4 mmol/g
14.4 mmol/g
7.6 mmol/g
10.1 mmol/g
Control

2

4

20

6
8
Angles (2 Theta)

10

12

B
basal spacing (A)

18
16
14

K-Smectite
Ca-Smectite
Cu-Smectite
Fe-Smectite

12
10
0

100

200

300

400

500

600

Tetracycline sorbed concentration (mmol/g clay)
 

Figure III-6. (A) X-ray diffraction patterns of K-smectite with increasing loading of tetracycline,
(B) increase of basal spacing as a function of tetracycline sorbed concentration
87 
 

 

Considering that the thickness of smectite is ~9.6 Å, and the basal spacing of smectites
were 14.7 Å with tetracycline loadings < 135 μmol/g, and 17.1 Å with loadings > 300 μmol/g,
the distance between two adjacent clay sheets was 5.1 Å and 7.5 Å respectively. Recall that the
molecular dynamics simulation results indicate that at low loading rate (< 135 μmol/g),
tetracycline lays parallel to clay surfaces with ABCD ring systems on the same planar and
dimethylammonium group locates beside ring system with extended formation (Figure III-5A).
The thickness of the extended structure is around 5.9 Å, which is close to XRD observed result
of 5.1 Å. When sorption approaches to 420 μmol/g, which is the sorption plateau in Figure III-2,
the clay sheets expands to 7.5 Å. At this distance tetracycline adapts a tilted position in clay
interlayers. Molecular dynamic simulation results for sorption at 420 μmol/g indicate that
tetracycline molecules adapt vertically tilted position. The disappearance of peak at 7.9º and
formed new peak at 5º with increasing tetracycline sorption from 135 to 218 μmol/g (Figure III6A) indicates that as tetracycline molecules enter clay interlayers, and adapt the position from
parallel to vertically tilted positions.

88 
 

 

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89 
 

 

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Sassman, S.A., and L.S. Lee. 2005. Sorption of three tetracyclines by several soils: Assessing the
role of pH and cation exchange. Environmental Science & Technology 39:7452-7459.
Schmitt, M.O., and S. Schneider. 2000. Spectroscopic investigation of complexation between
various tetracyclines and Mg2+ or Ca2+. Physchemcomm.
Sithole, B.B., and R.D. Guy. 1987a. Models for tetracycline in aquatic environments .2.
Interaction with humic substances. Water Air and Soil Pollution 32:315-321.
Sithole, B.B., and R.D. Guy. 1987b. Models for tetracycline in aquatic environments .1.
Interaction with bentonite clay systems. Water Air and Soil Pollution 32:303-314.
Smith, M.S., R.K. Yang, C.W. Knapp, Y.F. Niu, N. Peak, M.M. Hanfelt, J.C. Galland, and D.W.
Graham. 2004. Quantification of tetracycline resistance genes in feedlot lagoons by realtime PCR. Applied and Environmental Microbiology 70:7372-7377.
Teppen, B.J., K. Rasmussen, P.M. Bertsch, D.M. Miller, and L. Schafer. 1997. Molecular
dynamics modeling of clay minerals .1. Gibbsite, kaolinite, pyrophyllite, and beidellite.
Journal of Physical Chemistry B 101:1579-1587.
ter Laak, T.L., W.A. Gebbink, and J. Tolls. 2006. Estimation of soil sorption coefficients of
veterinary pharmaceuticals from soil properties. Environmental Toxicology and
Chemistry 25:933-941.
Tongaree, S., D.R. Flanagan, and R.I. Poust. 1999. The interaction between oxytetracycline and
divalent metal ions in aqueous and mixed solvent systems. Pharmaceutical Development
and Technology 4:581-591.

93 
 

 

Wessels, J.M., W.E. Ford, W. Szymczak, and S. Schneider. 1998. The complexation of
tetracycline and anhydrotetracycline with Mg2+ and Ca2+: A spectroscopic study.
Journal of Physical Chemistry B 102:9323-9331.
Witte, W. 1998. Medical consequences of antibiotic use in agriculture. Science 279:996-997.
Zhang, W., B.S.M. Sturm, C.W. Knapp, and D.W. Graham. 2009. Accumulation of tetracycline
resistance genes in aquatic biofilms due to periodic waste loadings from swine lagoons.
Environmental Science & Technology 43:7643-7650.
 
 

 

94 
 

 

CHAPTER IV
TETRACYCLINE SORPTION TO DIFFERENT CLAY MINERALS FROM WATER

95 
 

 

Abstract
Tetracyclines are a class of antibiotics commonly administered to control diseases and
improve health for humans and animals, especially used in the practice of livestock feeding
supplement at subtherapeutic levels. Tetracyclines enter the environment via land application of
biosolids derived primarily from animal manures and sewage sludge of wastewater treatment
plants. Understanding environmental fate and mobilization of tetracyclines is critical to assessing
their exposure and potential threats to at-risk populations. Tetracycline usually manifested a
great affinity to soil component, such as clay minerals and humic substances. In this study, we
investigated sorption of tetracycline onto 2:1 phyllosilicate smectite minerals (including
montmorillonite: SW-2, SAz-1; Beidellite: SBCa-1, Panther Creek) and non-expandable 1:1
kaolinite clay. Overall significantly greater sorption of tetracycline was observed for smectite
compared to kaolinite. No significant difference in sorption was observed for tetracycline
sorption on K- and Ca-smectite in acidic solution (e.g, pH 6.3 and 5.3). However, Ca-smectite
+

manifested much stronger sorption in alkaline solutions (e.g., pH 8.5). The presence of K or
Ca

2+

in aqueous solution suppressed tetracycline sorption because tetracycline was primarily

sorbed to clays via cation exchange reactions. In alkaline solution, sorption on Ca-smectite clays
was greater than that by K-smectite, which was attributed to formation of complexes of the
2+

negatively charged functional groups of tetracycline with Ca

on clay siloxane surfaces. K-

Panther Creek and K-SWy-2 clay has reached sorption plateau around 450 and 420 µmol/g clay
after aqueous equilibrium concentration reaches 0.15 µmol/mL. This sorption amount occupied
only ~50% of cation exchange sites on clay surface, but almost all surface area (> 100%) on clay

96 
 

 

minerals were occupied, which suggested that limit factor for tetracycline sorption onto smectite
was surface area not cation exchange capacity.

97 
 

 

Introduction
Tetracyclines are a class of antibiotics widely administered to treat diseases and improve
health for humans and animals, especially used routinely as supplementary components in
livestock feeds (Chopra and Roberts, 2001). Large fractions of tetracycline used in animal
treatment and feeding operations are excreted into livestock liquid and solid manures. After a
simple treatment, these manures are land applied to agriculture fields for plant fertilizer values.
This process is the primary pathway to disseminate tetracyclines into the environment. Kolpin et
al. (2002) found that tetracyclines such as chlortetracycline, oxytetracycline and tetracycline
were frequently detected in stream water with concentrations ranging from 0.11 to 0.42 µg/L.
Although tetracycline concentrations in water samples were not sufficient high to cause
immediate impacts to ecosystems, considering their high affinity to soils/sediments, substantial
amount of tetracyclines are still present in sediment/soil and water systems, potentially causing
chronic exposure and impacts to at-risk populations. Therefore, understanding the fate and
transport of tetracycline in soils is an important factor to assess the risks of tetracycline exposure
in the ecosystems.
Sorption to soil/sediment is an essential process affecting the fate, transport and
bioavailability of pollutants. It has been viewed that many nonpolar organic contaminants
primarily partition to soil organic matter. For tetracyclines and many other pharmaceuticals,
those compounds containing multiple types of polar and/or ionic functional groups commonly
leading to several speciation existing in aqueous environment; these species could develop
several specific interactions with soil organic matter and/or mineral fractions (Gu and
Karthikeyan, 2005; MacKay and Canterbury, 2005; Pils and Laird, 2007; ter Laak et al., 2006;
Wang et al., 2009). Among soil minerals the 2:1 aluminosilicate smectites are especially
98 
 

 

important clay fraction interacting with many classes of organic contaminants including
pharmaceuticals (Aggarwal et al., 2006; Boyd et al., 2001; Chappell et al., 2005; Charles et al.,
2008; Laird, 1999; Li et al., 2004; Sheng et al., 2002; Wang et al., 2009). Sorption by smectite
clay is influenced by many intrinsic properties such as interlayer swelling, high cation exchange
capacity (CEC), and large siloxane surface area.
Tetracycline structure and speciation chemistry in aqueous phase indicate that within
environmentally-relevant pH range of 5 to 8, tetracycline commonly exists in three species of
zwitterion, anion and a minor fraction of cation (Figure IV-1). These species determine sorption
mechanisms and driving forces for tetracycline sorption to clay minerals (Aristilde et al., 2010;
Figueroa et al., 2004; Sithole and Guy, 1987). Porubcan et al. (1978) studied sorption of
tetracycline by montmorillonite, and concluded that at low pH range (1.5 to 5.0) tetracycline
sorption to montmorillonite was driven primarily by cation exchange reaction with the major
species of cationic (at pH 1.5) and zwitterionic (at pH 5.0) tetracycline to montmorillonite. At pH
5.0, the major species of tetracycline is zwitterion; however, sorption was believed to cation
exchange reaction assuming that zwitterion simultaneously accompanied with proton to be
sorbed by mineral. Kulshrestha et al. (2004) also observed that tetracycline sorption decreased
with increasing pH from 1.5 to 11.0, which was attributed to decreasing cationic fraction of
tetracycline in aqueous phase hence reducing the contributions of cation exchange reaction. In
addition to cation exchange-driven sorption, formation of complexes between tetracycline and
metal ions is another mechanism responsible for tetracycline sorption, particularly at alkaline
condition in which minimal amount of cationic tetracycline is present in aqueous phase (Aristilde
2+

et al., 2010; MacKay and Canterbury, 2005; Pils and Laird, 2007; Porubcan et al., 1978). Ca +

saturated soils or clays usually manifested a greater sorption for tetracycline than K -saturated
99 
 

 

sorbents (Pils and Laird, 2007), which was attributed to the formation of complexes between
negatively charged functional groups in tetracycline and Ca

2+

(MacKay and Canterbury, 2005).

Using spectroscopic approaches, Aristilde et al. (2010) observed the formation of complexes
+

between deprotonated phenol groups of oxytetracycline and exchanged Na on clay surfaces,
though it is more common to accept that the complexes are more readily formed between
tetracycline with multivalent metal cations, which have been reported to occur in aqueous
solution (Benet and Goyan, 1965; Berthon et al., 1983; Brion et al., 1981; Lambs et al., 1988;
Ohyama and Cowan, 1995; Wessels et al., 1998).
The objective of this study was to compare tetracycline sorption to phyllosilicate
smectites and kaolinite under varying solution pH and electrolyte background. Factors such as
clay characteristics, type of exchangeable cation, cation exchange capacity, surface area and
locations of deficit negative charges were evaluated for their effects on tetracycline sorption in
aqueous systems.

100 
 

 

A
HO

CH3

7
D

C
11

9
OH

O
pKa2=7.7
diketon

B

H(TC)-

H2(TC)

H3(TC)+

100
Speciation fraction (%)

H

HN(CH 3)2
4
OH
5
3
B
A
2
12
1
C NH2
pKa1=3.3
OH
OH
O
O
tricarbonyl methane

6

10

8

pKa3=9.7
dimethylammonium

TC2-

80
60
40
20
0
0

2

4

6

pH

8

10

12

Figure IV-1. (A) Tetracycline structure, and (B) speciation distribution as a function of solution
+

-

2-

pH. H3(TC) , H2(TC), H(TC) , and TC represent the cation, zwitterion, anion, and divalent
anion of tetracycline species.

101 
 

 

Materials and Methods
Clay Minerals and Chemicals
Reference clay minerals were purchased from the Source Clays Repository of the Clay
Minerals Society. Selected clay mineral characteristics are listed in Table IV-1. Sedimentation
+

method was used to collect < 2 µm-sized clay fractions, which were subsequently K or Ca

2+

saturated by washing the clay fractions with 0.5 M KCl and CaCl2 solution three times.
Excessive KCl and CaCl2 were removed by washing clay with Mili-Q water five times. The
-

removal of Cl was examined by negative test result of AgNO3 solution. Suspensions were
quickly frozen and freeze dried to receive clay powders.
Tetracycline hydrochloride (≥ 95%) and heptafluorobutyric acid were purchased from
Sigma-Aldrich Chemical Company (St. Louis, MO), and was used as received. Methanol, CaCl2,
KCl, and ammonium acetate were purchased from J.T. Baker (Phillipsburg, NJ, USA). All
aqueous solutions were prepared using water obtained from a Millipore Milli-Q Water
purification system followed by 0.2 μm membrane filtration.

102 
 

 

Table IV-1. Selected properties of clay mineral used in this study

Minerals

SWy-2
SHCa-1
Panther
Creek
SAz-1
SBCa-1
KGa-1

Clay type
montmorillonite,
dioctahedral
hectorite,
trioctahedral
beidellite,
dioctahedral
montmorillonite,
dioctahedral
beidellite,
dioctahedral
kaolinite,

Tetrahedral
charge (%)

Cation
Exchange
Capacity
(cmol/kg)

Surface
2
area(m /g)

Surface
charge
density
2
(µmol/m )

0

78

766

1.02

0

44

743

0.59

52

88

619

1.42

12

130

768

1.69

98

66

750

0.88

2

20

1.00

Soprtion Isotherm Measurement.
Batch equilibration method was used to measure sorption of tetracycline by clay minerals
from aqueous solution. Tetracycline initial solutions were prepared with concentrations ranging
from 0 to 340 µmol/L. A certain amount of clay was mixed with a known volume of tetracycline
initial solution in glass centrifuge tubes on a rotator shaker at 40 rpm for 18 h at room
temperature (~23 ºC). The tubes were subsequently centrifuged at 6000 g for 30 min at 23 ºC.
The supernatant pH was measured using a pH meter equipped with a combination electrode. All
samples were prepared in duplicate. An aliquot of supernatant was collected and analyzed for
tetracycline concentrations using a liquid chromatography coupled to a tandem mass
spectrometer (LC-MS/MS) system consisting of a Shimadzu high-performance liquid
chromatography (HPLC, Columbia, MD, USA) fully integrated with an Applied Biosystems
Sciex 3200 triple quadrupole mass spectrometer (Foster City, CA, USA). A PEAK Scientific gas
generator (NM20ZA, Billerica, MA, USA) was used to supply gases required for the mass
103 
 

 

spectrometer. A binary mobile phase was used consisting of solutions with water and methanol
(v/v = 95/5), and methanol and water (v/v = 95/5), both containing 2 mM ammonium acetate and
20 mM heptafluorobutyric acid. A Thermo Hypersil Gold column (50 × 2.1 mm, particle size 5
-1

µm) was used with a flow rate of 0.15 mL min . Tetracycline concentration was quantified
using multiple reaction monitoring mode with the precursor/product transition of 445.0/410.0.
Sorbed concentrations were calculated from the difference between the initial and equilibrium
concentrations in aqueous solution.
Results and Discussion
Tetracycline contains three ionic functional groups which can be ionized in water to
produce four species (Figure IV-1). This speciation chemistry of tetracycline governs sorption
mechanism to clay minerals. In addition, anionic species can form metal-tetracycline complexes
in aqueous solution. The distributions of the four species in water as a function of solution pH
were calculated based on the relationship between solution acidity and tetracycline pKa values
(Figure IV-1). In this study, equilibrium solution pH values ranged from 3.3 to 9.0 depending on
the type of clays and the amount of tetracycline present. Within this pH range, tetracycline exists
in the forms of cation, zwitterion and anion containing one negative charge. At pH > 7.7, anionic
form is the most abundant species, whereas at pH between 3.3 and 7.7, zwitterion of tetracycline
is the dominant species.
Tetracycline sorption on Wyoming montmorillonite (SWy-2), Arizona montmorillonite
(SAz-1), Panther Creek beidellite, California beidellite (SBCa-1) and kaolinite (KGa-1) from
water was measured with pH value ranging from 3.8 to 8.4 (Figure IV-2). At lower aqueous
equilibrium concentrations (< 0.007 µmol/mL), tetracycline sorption to K-Panther Creek mineral
104 
 

 

demonstrated a liner relationship. As aqueous tetracycline concentration increased, sorption
isotherm became nonliner, and approached to sorption plateau at ~460 µmol/g. Wyoming
montmorillonite and California beidellite manifested similar nonlinear sorption isotherms.
Arizona montmorillonite demonstrated a S-shape sorption isotherm. Arizona montmorillnite has
2

the highest CEC (130 cmol/kg) and comparable surface area (768 m /g), but did not manifest the
greatest sorption for tetracycline. This could be attributed to that negatively charged sites on
smectite surface is not the limited factor for sorption of relatively large sized pharmaceuticals
such as lincomycin (Wang et al., 2009).
+

Sorption of tetracycline to K -saturated Panther Creek clay from aqueous solution
containing 0.0, 0.001, 0.01 and 0.10 M KCl (Figure IV-3). Increasing sorption suppressive effect
was observed with increasing KCl concentration. Given pH ranges from 3.9 to 6.6, the major
species of tetracycline were cation and zwitterion, and the sum of these two species was > 94%
in the solutions. This speciation of tetracycline resulted in sorption on smectite mineral to be
primarily driven via cation exchange reaction. The protonated dimethylamino functional moiety
in tetracycline plausibly replaces the exchangeable potassium and interact with negatively
charged sites on mineral surfaces. The results are consistent with previous studies that cation
exchange mechanism was predominant mechanism controlling tetracycline sorption to soil
minerals and soils (Figueroa et al., 2004; Sassman and Lee, 2005).

105 
 

Sorbed tetracycline on clay (mmol/g)

 

600
500
400
300

K-Panther Creek
K-Montmorillonite, Wyoming
K-Beidellite, California
K-Kaolinite
K-Montmorillonite, Arizona
6.5
5.2
6.1
7.4
5.3
7.9
5.9
7.6

200
100

6.1 8.3
6.9 8.4

4.6

3.8

6.2
6.2

5.5

5.9 7.2
6.0
6.3

5.4

7.2
5.4

4.4

0
0.00
0.05
0.10
0.15
0.20
0.25
Aqueous tetracycline equilibrium concentration (mmol/mL)

 

+
Figure IV-2. Tetracycline sorption to K -saturated clay minerals. The solid curves represent the

fitting results of Langmuir equation. The equilibrium solution pH values are marked close to
sorption data
Tetracycline sorption to clay minerals was fit to the Langmuir equation assuming only
one type of surface sites is available sorption (Figueroa et al., 2004). The sorption sites are
assumed to be cation exchange sites on mineral surfaces and to demonstrate the same affinity for
tetracycline.

kQ Cw
Cs = 0
1+ kCw
106 
 

(1)

 

where Cs (μmol/g) and Cw (μmol/mL) are tetracycline concentration sorbed by clay and in
aqueous phase, respectively. Q0 (µmol/g) is the maximal sorption capacity, and k (mL/µmol) is
Langmuir sorption coefficient. The fitting results are shown as curves in Figures IV-2 and IV-3,
and the corresponding parameters are listed in Table IV-2. For different clay minerals, SWy-2
and SBCa-1 clay minerals were estimated to manifest higher sorption (Q0 = 700 and 768 µmol/g,
respectively) than that by Panther Creek mineral (Q0 = 533 µmol/g). But the fitted k values
indicate that tetracycline is more strongly sorbed to Panther Creek clay than that by SWy-2 and
SBCa-1. The presence of increasing KCl in solution did not result in decreased Q0 values, but
+

significantly reduced k values, suggesting that K ions present in aqueous solution effectively
competed with tetracycline for cation exchangeable sites on clay surfaces.

107 
 

 

Table IV-2. Langmuir equation fittings for tetracycline sorption onto clay minerals
Clays and solution conditions Q0 (µmol/g) k (mL/µmol)

r

2

K-Panther Creek
0.0 M KCl
0.001 M KCl
0.01 M KCl
0.10 M KCl

a

533 ± 41
637 ± 37
976 ± 167
882 ± 378

63 ± 12
35 ± 4
10 ± 2
3±2

0.99
0.99
0.99
0.98

700 ± 94

9±2

0.97

768 ± 219

4±1

0.98

8.5 ± 1

248 ± 88

0.90

K-SWy-2
0.0 M KCl
K-SBCa-1
0.0 M KCl
K-KGa-1
0.0 M
a

± standard error

108 
 

 

Sorbed tetracycline on clay (umol/g)

600
0M
0.001 M
0.01 M
0.1 M

500
400
300
200
100
0
0.00

0.02

0.04

0.06

0.08

0.10

0.12

0.14

0.16

Aqueous tetracycline equilibrium concentration (umol/mL)
Figure IV-3. Tetracycline sorption to K-Panther Creek from aqueous solution containing 0.0,
0.001, 0.01 and 0.10 M KCl.

109 
 

 
+

For K -saturated clay minerals, solution pH change only affects the distribution of
tetracycline speciation in aqueous phase hence the extent of cation exchange equilibration.
2+

However, for Ca -saturated clays as solution pH increases to alkaline condition, formation of
Ca-tetracycline complexes on mineral surfaces is believed to enhance sorption. For instance of
hectorite, the aqueous solution pH values after tetracycline sorption equilibrium were 8.4 to 9.2
-

for K- and Ca-SHCa-1 clay. Within this pH range, most tetracycline exists as anion (HTC ) in
aqueous solution. For K-SHCa-1 sorption reached to a plateau at 205 µmol/g when aqueous
tetracycline concentration was 0.9 µmol/mL. At the similar pH and aqueous tetracycline
concentration, Ca-SHCa-1 sorption of tetracycline reached to 570 µmol/g (Figure IV-4A). The
-

2+

dominant HTC species could interact with exchanged Ca

on mineral surfaces to form

complexes, hence demonstrate an enhanced sorption vs. sorption by K-SHCa-1. As for the claywater system at relatively lower pH such as K- and Ca-Panther Creek clay with pH ranging from
5.2 to 7.3 (Figure IV-4B), there was only a little distinction for tetracycline sorption by K- vs.
Ca-Panther Creek clays. Within this pH range, the primary tetracycline species is zwitterion;
cation exchange is invoked as the dominant sorption mechanism for tetracycline sorption.

110 
 

Tetracyclien sorbed on clay (mmol/g)

 

A
600

K-Hectorite, pH 8.5
Ca-Hectorite, pH 8.8 600

400

400

200

B

200

0
0.00

0.05

0.10

0.15

0
0.20 0.00

K-Panther Creek, pH 6.3
Ca-Panther Creek, pH 5.3

0.01

0.02

0.03

Aqueous tetracycline equilibrium concentration (mmol/mL)

0.04

Figure IV-4. Tetracycline sorption to (A) K- and Ca-SHCa-1 and (B) K- and Ca-Panther Creek clays. The pH values for
hectorite/water systems were averaged at ~8.6, and the pH values for Panther Creek clays ranged from 5.2 to 7.3.

111 
 

0.05

 
-

It has been reported that in aqueous solution tetracycline anionic form (HTC ) forms
2+

complex with Ca

at metal-to-ligand ratio of 1:1, the corresponding formation constant (log K)

was measured at 12.73 (Martin, 1979). Other complexes with metal-to-ligand ratios of 2:1 or 1:2
also formed with the formation constants of 8.67 and 17.62, respectively (Brion et al., 1981;
Lambs et al., 1988). These results indicate that strong complexation occurs between tetracycline
and Ca

2+

in water. The formed complexes with varying metal-to-ligand ratios are determined by

metal cation and tetracycline concentration, and pH in water. For the complexation occurring on
mineral surfaces, Ca

2+

present on mineral surfaces usually manifests a more condensed scenario

compared to that in solution phase. In addition, sorbed tetracycline on clay surfaces enhances the
contact probability with exchanged cation due to the condensed sorbent phase. Furthermore, the
subaqueous clay interlayer environment plausibly enhances the interaction strength between the
2+

negatively charged moiety in tetracycline and Ca

by formation of inner-sphere complexes.

Tetracycline complexation with exchanged cations on mineral surfaces is expected to be
2+

enhanced for transition metal cations such as Cu

3+

and Fe

solution the complex formation constant between tetracycline and Cu
2+

magnitudes greater than that for Ca

and Mg

2+

2+

compared to Ca . In aqueous
2+

is six order of

of 1:1 metal-to-ligand complex (MacKay and

Canterbury, 2005; NovakPekli et al., 1996). Therefore, it is expected that Cu-clay would
manifest a strong affinity for tetracycline than Ca-clay. Our measured results are consistent with
this expectation that tetracycline sorption onto Cu- and Fe-SWy-2 was higher than Ca-SWy-2
(Data not shown here).

112 
 

 

In addition to the factors of solution acidity/alkalinity and exchangeable cations
compensating negative charges on mineral surfaces, clay type also influences tetracycline
sorption. In general, swelling clay minerals (i.e. smectites) displayed much greater sorption for
tetracycline than non-expandable kaolinite mineral (Figure IV-2). Kaolinite is a 1:1 tetrahedral :
octahedral silicate with a small amount of cation exchange sites due to dissociation of –OH
groups on broken edges. Strong hydrogen bonding between the adjacent clay layers prevents
kaolinite from expanding when hydrated (Schroth and Sposito, 1997). Therefore, kaolinite has
the least sorption sites for tetracycline due to the insufficient amount of cation exchange sites and
small surface area available for sorption. On the other hand, smectites are 2:1 phyllosilicates
containing high cation exchange capacity (i.e. negatively charged sites) and large surface area
available for tetracycline sorption when expanded. As expected, tetracycline displayed a much
stronger sorption for smectite clays compared to sorption by kaolinite (Figure IV-2).
For tetracycline sorption on smectites, sorption approached close to plateaus at 420, 450
and 280 µmol/g for K-SWy-2, K-Panther Creek and K-SBCa-1, respectively (Figure IV-2). The
2

molecular surface area of tetracycline was estimated at ~178 Å . Assuming cation exchange sites
and surface area are fully accessible for tetracycline sorption, and the sorbed tetracycline are
positioned with ABCD ring systems flat on mineral surfaces, the occupation of cation exchange
sites and surface area was calculated for tetracycline sorption by K-SWy-2, K-Panther Creek
and K-SBCa-1 as reaching to plateaus (Table IV-3). Approximately 42 to 54% of cation
exchange sites are occupied by tetracycline assuming that tetracycline is sorbed singly by cation
exchange reaction. And ~40 to 78 % of smectite surface area was estimated to be occupied by
tetracycline molecules. The sorbed tetracyclines plausibly adapt a flat position with the
molecular plane parallel to clay siloxane sheets at lower sorption rate. As sorption increases, the
113 
 

 

accessible mineral surface areas become a limiting factor for tetracycline to adapt that parallel
position, though sufficient amount of cation exchange sites are still available. To compromise
this limiting factor, tetracycline molecules may form a double-layer or vertically tilted
configurations in clay interlayers, resulting in further expansion of smectite basal spacing. For
many large-sized cationic pharmaceuticals sorption by smectites is not limited by the amount of
negatively charged sites (i.e. cation exchange sites) on mineral surfaces but by accessible surface
2

area, despite smectite has very large surface area of ~800 m /g.

Table IV-3. Cation exchange site and surface area of clay mineral occupied by tetracycline
Minerals
SWy-2

Clay Type

CEC Occupied (%) Surface area occupied (%)

montmorillonite, dioctahedral

54

58

Panther Creek

beidellite, dioctahedral

51

78

SBCa-1

beidellite, dioctahedral

42

40

 

Conclusions
Tetracycline structure and solution chemistry determine the formation of cationic,
zwiterrionic and anionic species in solution, which pose significant impacts to sorption by
soil/sediment. Results from this study revealed that tetracycline sorption to clay minerals was
highly dependent on solution pH. Cation exchange reaction and metal-ligand complexation are
114 
 

 

the two major mechanisms responsible for tetracycline sorption on clay surfaces. For acidic and
neutral solution, replacement of metal cation associated with clay with tetracycline is the
dominant mechanism responsible for the overall sorption process. In alkaline solution complex
formation between tetracycline anion and multivalent metal cation on mineral surface also
contributes to the overall sorption. Theoretically, tetracycline could be sorbed on smectites up to
reaching cation exchange capacity when cation exchange reaction dominates the overall sorption;
however, the large molecular size of tetracycline results in that the accessible mineral surface
area becomes the limiting factor in spite of the large surface areas of smectite.

115 
 

 

REFERENCE
 

 

116 
 

 

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Aristilde, L., C. Marichal, J. Miehe-Brendle, B. Lanson, and L. Charlet. 2010. Interactions of
oxytetracycline with a smectite clay: A spectroscopic study with molecular simulations.
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