environmental effectors on the inactivation of human adenoviruses in water

ORIGINAL PAPER

Environmental Effectors on the Inactivation of HumanAdenoviruses in Water

Anna Carratala • Marta Rusinol • Jesus Rodriguez-Manzano •

Laura Guerrero-Latorre • Regina Sommer •

Rosina Girones

Received: 5 June 2013 / Accepted: 2 August 2013 / Published online: 17 August 2013

� Springer Science+Business Media New York 2013

Abstract Environmental factors are highly relevant to

the global dissemination of viral pathogens. However, the

specific contribution of major effectors such as temperature

and sunlight on the inactivation of waterborne viruses is

not well characterized. In this study, the effect of temper-

ature (7, 20, and 37 �C), UVB and UVA radiation on viral

inactivation was evaluated in phosphate buffered saline

(PBS), mineral water, wastewater, 1,000-fold diluted

wastewater and seawater. The stability of human adeno-

viruses infectivity, known as human pathogens and indi-

cators of fecal contamination, was monitored during 24 h,

both in the dark and exposed to UV radiation by immu-

nofluorescence assays. In the dark, no Human adenovirus

(HAdV) inactivation was observed in PBS and mineral

water at any of the temperatures studied, whereas at 37 �C

in reactors with higher microbial concentration (wastewa-

ter, diluted wastewater, and seawater), decays between 2.5

and 5 log were recorded. UVB radiation showed a dramatic

effect on HAdV inactivation and 6-log were achieved in all

reactors by the end of the experiments. The effect of UVA

showed to be dependent on the water matrix analyzed. At

20 �C, HAdV showed a 2-log decay in all reactors radia-

tion while at 37 �C, results in wastewater, diluted waste-

water, and seawater reactors were equivalent to those

observed in the dark. These results suggest UVB radiation

as the major environmental factor challenging viral inac-

tivation, followed by biotic activity indirectly associated to

higher temperatures and finally, by UVA radiation.

Keywords UVA � UVB � Temperature �Microorganisms � Inactivation � Viruses � Water �Infectivity

Introduction

Wastewater and contaminated water used for irrigating,

drinking, and recreational purposes has been widely linked

with the transmission of infectious viral diseases among

human populations (McKinney et al. 2006; ter Waarbeek

et al. 2010; Riera-Montes et al. 2011; Nenonen et al. 2012).

Viral pathogens are frequently released in stool and urine

in high numbers, and may contaminate surface and coastal

waters establishing an important route for their transmis-

sion and a public health concern.

It is widely accepted that the stability of a virus is key to its

capability to remain infectious over time in the environment

or after disinfection treatments and consequently, to spread

and cause disease. In particular, a high stability has been

reported for certain enteric viruses in a range of environ-

mental conditions and after various disinfection treatments

(Carter 2005). However, it is known that viral pathogens may

be naturally inactivated, by both abiotic and biotic factors

(Gordon and Toze 2003; Bertrand et al. 2012).

Sunlight is known as a challenge to the persistence of

viruses (Viau et al. 2011; Boehm et al. 2009) and it has been

A. Carratala � M. Rusinol � J. Rodriguez-Manzano �L. Guerrero-Latorre � R. Girones (&)

Department of Microbiology, Faculty of Biology, University of

Barcelona, Av. Diagonal 643, 08028 Barcelona, Catalonia, Spain

e-mail: [email protected]

Present Address:

A. Carratala

Laboratory of Environmental Chemistry, School of Architecture,

Civil and Environmental Engineering (ENAC), 1015 Lausanne,

Switzerland

R. Sommer

Water Hygiene, Medical University Vienna, Institute for

Hygiene and Applied Immunology, Vienna, Austria

123

Food Environ Virol (2013) 5:203–214

DOI 10.1007/s12560-013-9123-3

exploited in the last 15 years as a low-cost alternative pro-

cedure to inactivate pathogens in drinking water (McGuigan

et al. 2012). UVB (280–320 nm) and UVA (320–400 nm)

are the main sunlight wavelengths inactivating viral patho-

gens. While it is known that UVB is highly efficient on the

inactivation of microorganisms by direct photo-inactivation,

its fluence may be affected by weather conditions, water

quality and depth. However, UVA is less influenced by these

factors and may also be implicated on solar disinfection,

mainly by indirect photo-inactivation mechanisms. Simi-

larly, temperature is an important factor for the stability of

viruses in the environment and heat treatments have shown

to inactivate viral pathogens in water and food (Bertrand

et al. 2012; Carratala et al. 2013).

There is evidence that enteric viruses may persist after

water disinfection treatments that eliminate bacteria (Tree

et al. 2003; Sirikanchana et al. 2008; Simmons and Xago-

raraki 2011; Rodriguez-Manzano et al. 2012) and concerns

have been raised regarding the use of bacterial parameters as

indicators of fecal contamination in the environment (Gerba

et al. 1979; Lipp et al. 2001). Due to their epidemiologic

characteristics, human adenoviruses have been proposed as

alternative indicators of human fecal contamination and

have been used to evaluate the efficiency of water disinfec-

tion treatments (Pina et al. 1998; Formiga-Cruz et al. 2003;

Albinana-Gimenez et al. 2009). Human adenovirus type 2

(HAdV) are excreted by healthy and symptomatic humans at

high concentrations and are widely prevalent in urban

wastewater in different geographical areas year round (Pina

et al. 1998; Bofill-Mas et al. 2006; Fong et al. 2010; Kok-

kinos et al. 2011; Rigotto et al. 2010).

In order to achieve a better understanding of the dissem-

ination of viral pathogens in water, the role of different

environmental factors on the inactivation of viruses must be

characterized. The main objective of the present study is to

investigate the effect of temperature, UVB and UVA radia-

tion on the natural inactivation of HAdV in a diversity of

water types that are relevant to the dissemination and

transmission of viral pathogens. The studied matrices are

phosphate buffered saline (PBS), mineral water, wastewater,

diluted wastewater, and seawater. Experimental data on viral

inactivation in a wide range of environmental conditions is

needed to develop mathematical models predicting viral

dissemination in superficial waters and to implement effi-

cient control measures in case of waterborne outbreaks.

Materials and Methods

Viral Suspensions

Human adenovirus type 2 (HAdV2) stocks were produced

by infecting A549 cells cultured in Earl’s minimum

essential medium (EMEM, Gibco, Frederick, MD) sup-

plemented with 1 % glutamine, 50 lg of gentamycin per

ml and 10 % (growth medium) or 2 % (maintenance

medium) of heat-inactivated fetal bovine serum (FBS).

Viruses were released from cells by freezing and thawing

the cultures three times. Then, a centrifugation step at

3,0009g for 20 min was applied to eliminate cell debris.

Viruses were ultra-centrifuged at 34,5009g and the

obtained pellet was re-suspended in PBS without further

purification. Finally, viral suspensions were quantified and

stored in 10 ml aliquots at -20 �C until use.

UV Radiation

The UV radiation applied in the experiments was obtained

by exposure to a 400-W Philips MSR400 HR hydrargyrum

medium-arc iodide lamp focus (Koninklijke Philips Elec-

tronics, The Netherlands). The lamp was installed inside a

thermo-static chamber. As previously described, this lamp

has an emission spectrum that approximates the fraction of

UVA, UVB, and photosynthetically active radiation of

natural sunlight, a continuous spectrum from 250 to

800 nm, and a irradiance plateau from 400 to 700 nm

(Dick et al. 2010). The total output fluence was controlled

during our experiments using a pyranometer (LP Pyra03,

Delta Ohm Srl, Italy) and UVA (320–400 nm) and UVB

radiation (280–320 nm) were measured by means of

spectral-specific sensors (LP UVA and UVB, Delta Ohm

Srl, Italy). As a reference measure, the total irradiance and

the irradiances of UVA and UVB were measured outdoors

on a spring sunny day at 23 �C; the total irradiance was

97 W/cm2 and the irradiances of UVA and UVB were 506

and 40 lW/cm2 respectively.

The mean total irradiance emitted by the lamp was

35.7 W/cm2. The wavelength distribution in the UV region

emitted by this lamp is highly dependent on the tempera-

ture and consequently, in our experiments there was no

need to use radiation filters to expose the tested reactors to

UVB and UVA. While at 7 �C, UVA fluence was

29.74 lW/cm2 and UVB fluence was 0.13 lW/cm2, at 20

and 37 �C no UVB fluence was registered and UVA flu-

ence was 38.81 and 31.14 lW/cm2, respectively. The mean

values of radiation fluence applied to the reactors at each

sampling time are presented in Table 1.

Water Matrices and Experimental Design

Different types of water were selected on the basis of their

potential role in the dissemination of viral pathogens in the

environment. The water matrices studied in these experi-

ments are PBS as a control matrix, mineral water, waste-

water, and wastewater 1,000-fold diluted in mineral water

as a representation of superficial waters showing two

204 Food Environ Virol (2013) 5:203–214

123

different levels of fecal contamination and seawater. The

1,000-fold dilution for wastewater was selected based on

previous studies where 3 log of reduction in the concen-

tration of human adenoviruses were commonly observed

between raw sewage (Bofill-Mas et al. 2006) and the river

water samples analyzed in the area (Albinana-Gimenez

et al. 2009).

Commercial mineral water was acquired from a local

retailer. Raw wastewater was collected at the entry of a

treatment plant in Catalonia that receives sewage from

2,275,000 inhabitants. Seawater samples were collected on

a beach located south of Barcelona, taking care to avoid

sand and macro-algae. Turbidity, conductivity, pH, and

light absorbance were determined for each water type at the

beginning of the assay.

Reactors were prepared by distributing 150 ml of water

sample in Pyrex glass beakers standing on magnetic stirrers

in a thermostatic chamber where the temperature was

controlled during the experiments at 7, 20, and 37 �C. In

this work, we intended to study the effect on virus inacti-

vation of different temperatures representing conditions

that may be found in diverse geographical areas at different

times of the year. It is known that cold temperatures are

optimal for virus conservation and consequently, no tem-

peratures below 7 �C were included in this work.

Inside the chamber the effect of both dark and light

conditions was investigated at each temperature. Dark

conditions were created in parallel and achieved by cov-

ering the Pyrex glass beakers with an opaque paper-box. In

all experimental conditions, water reactors were covered

with 0.05 mm-thick FEP-Teflon films, 96 % transparent to

UVA and UVB light (Dupont Corporation, Wilmington,

DE) to protect the reactors from external inputs (Dick et al.

2010).

Human adenovirus suspensions were spiked in the

reactors at a concentration of 106 foci forming U/ml (FFU/

ml). Each single experiment was conducted in duplicate for

24 h collecting samples at five time points (0, 2, 4, 8, and

24 h). Ten-ml samples were taken at each sampling time

and filtrated through 0.22 lm filters (Millex-GP, Millipore,

Ireland) to avoid contamination that might interfere with

the infectivity assays. Immunofluorescence assays (IFA) to

evaluate HAdV infectivity after the inactivation experi-

ments were performed within 24 h, keeping the samples at

4 �C until that time.

Infectivity Assays

IFA for the quantification of infectious HAdV2 were per-

formed as described in previous work (Calgua et al. 2011).

In brief, A549 monolayers were incubated overnight in

8-well chamber slides (Lab-Tek II, Nagle Nunc Interna-

tional, Naperville, IL) at 37 �C in 5 % of CO2 until

90–100 % of confluence. One hundred microliters of direct

or diluted samples were inoculated in triplicate into each

well and incubated for 90 min at 37 �C in 5 % CO2. After

this, 2 % FBS-supplemented EMEM was added and incu-

bated for 3 days at 37 �C in 5 % CO2. After 3 days’

incubation, the monolayers were fixed with chilled absolute

methanol for 10 min and rehydrated by soaking in PBS for

Table 1 Fluences of total

radiation, UVA and UVB

received by the samples at each

sampling time during the

experiments

Sampling times (h) Fluence

Total radiation (kJ/m2) UVA (J/m2) UVB (J/m2)

7 �C 0 0 0 0

2 1.803 2.185 5.35

4 3.567 4.277 11.28

8 7.059 8.623 18.96

24 20.954 25.413 122.91

20 �C 0 0 0 0

2 2.081 2.080 0

4 4.162 4.162 0

8 8.258 8.257 0

24 24.875 24.875 0

37 �C 0 0 0 0

2 1.686 2.246 0

4 3.246 4.492 0

8 6.742 8.985 0

24 20.377 26.957 0

Food Environ Virol (2013) 5:203–214 205

123

5 min. The monolayers were blocked for 1 h in blocking

solution containing PBS with 1 % BSA (w/v) and 0.05 %

Tween (v/v). After removing the blocking solution, the

monolayers were incubated for 1 h with their first antibody

solution. Cells were then stained for 15 min at RT with a

1:100 dilution of goat anti-mouse IgG-FITC (Sigma-

Aldrich, Steinheim, Germany) in blocking solution. The

IgG antibody was removed and cells were rinsed with PBS

for 15 min. Finally, chambers were mounted by adding

UltraCruzTM mounting medium (Santa Cruz Biotechnol-

ogy, Inc.) and observed under UV light in an epifluores-

cence microscope.

Data Analysis and Estimation of Inactivation Times

(T90, T99, T99.9, and T99.99)

Viral concentrations were estimated as measured in tripli-

cate by IFA. The decay of infectious viruses was modeled

by different approaches as previously described (De Roda

Husman et al. 2009; Romero et al. 2011; Tuladhar et al.

2012; Kovac et al. 2012; Verhaelen et al. 2012). Linear

regression analysis, monophasic, and biphasic models were

used to estimate inactivation times and the corresponding

confidence intervals for each condition. Constant expo-

nential decay was modeled as described by Chick (1910):

log10 ct½ � ¼ log10 c0exp �ktð Þ½ �

In this equation, ct is the infectious virus concentration

after a certain time (t) exposed to the specific experimental

conditions. C0 corresponds to the initial concentration and

k the inactivation rate constant (h-1). Second, viral

inactivation was modeled as biphasic (De Roda Husman

et al. 2009):

log10 ct½ � ¼ log10 c0 x exp �k1tð Þ þ ð1� xÞ exp �k2tð Þð Þ½ �

The parameter x refers to mixing and k1 and k2 are the

two inactivation rate constants. The parameters estimates

were obtained by the Least squares estimation and the best

fitting model was chosen using the Aikaike Information

Criterion (Hurvich and Tsai 1989). The comparison of

inactivation parameters between different experiments was

performed by F test. When no significant differences were

observed, the monophasic model was chosen as the most

parsimonious. All analyses were performed using Prism5

(Version 5.0, GraphPad Software 2007).

The value T90 is defined as the time required for 1-log

inactivation of the initial viral concentration. Generation of

regression curves and prediction of the times required for a

viral inactivation from 1 to 4-log were accomplished by

including data points from all experiments for each virus

and water reactor at different temperatures. The predicted

(discontinuous line) and observed (individual data points)

values were plotted on figures (Prism5 version 5.0.

GraphPad Software 2007), representing log inactivation

[log (Nt/N0)] versus time (Figs. 1, 2, 3).

Results

Characterization of the Water Types

Each water type tested in the experiments was characterized

by measuring turbidity, conductivity, pH, and light absor-

bance. The data obtained are shown in Table 2. From our

measurements, the highest turbidity values were detected in

wastewater (142 NTU) and seawater (3.69 NTU). Conduc-

tivity values measured in PBS and seawater samples were the

highest of all, being 15.29 and 54.5 mS, respectively. pH

ranged from 7.35 in PBS to 8.8 in seawater samples. As

expected, reactors containing wastewater showed the highest

absorbance values in the range of UV wavelength. Light

screening can be calculated according to the following equa-

tion as previously described (Schwarzenbach et al. 2003):

IðiÞ ¼ 10�a�L

In this equation, a corresponds to the absorbance value

in the UVB range and L the depth in the beaker (5 cm).

From this, only a 17 % of the UVB radiation was absorbed

in the reactors containing wastewater. Considering the

experimental error in the assays, UV light absorption by the

suspended particles in the tested waters may be neglected

to analyze the results obtained in this study.

Effect of Temperature on HAdV2 Inactivation in Water

The effect of temperature on the inactivation of HAdV was

assayed performing the experiments at 7, 20, and 37 �C in

dark conditions. The obtained inactivation curves are pre-

sented in Fig. 1. In wastewater reactors, more than 5-log

inactivation of HAdV were achieved after 24 h at 37 �C,

approximately corresponding to the complete inactivation

of the initial viral load, while under the same conditions, a

3-log decay was observed in seawater reactors.

The correlation values (R2), inactivation times (T90, T99,

T99.9, and T99.99) and inactivation rate constants (k) are

detailed in Table 3. The inactivation rate constants were

obtained by linear regression and did not differ statistically

from 0 (p value [ 0.05) in 24 h in PBS or in mineral water, at

the temperatures studied in the absence of light. Similarly, in

the reactors containing wastewater and seawater, no signif-

icant inactivation of HAdV2 was observed at 7 and 20 �C.

However, higher inactivation rate constants of infectious

HAdV were identified both in wastewater (k & 0.24 ±

0.05 h-1) and seawater (k & 0.13 ± 0.06 h-1) reactors at

37 �C, probably caused by the action of biota naturally

present in the samples. Interestingly, equivalent results were


123

observed in reactors containing diluted wastewater in min-

eral water (k & 0.24 ± 0.05 h-1). From our results, the

inactivation times needed to achieve a reduction of 4-log in

HAdV initial concentrations (T99.99) are 15.82 h in waste-

water, 34.09 h in diluted wastewater in mineral water

(1:1,000), and 31.76 h in reactors containing seawater.

Mineral water

0 10 20 30

-2

0

2

4

6

8

Time (h)Lo

g10

(Nt/N

0)

Wastewater

0 10 20 30

-2

0

2

4

6

8

Time (h)

Lo

g10

(Nt/

N0)

Wastewater 1/1000

0 10 20 30

-2

0

2

4

6

8

Time (h)

Log 10

(Nt/N

0)

Log 10

(Nt/N

0)

Seawater

0 10 20 30

-2

0

2

4

6

8

Time (h)

Fig. 1 Effect of temperature on

the stability of HAdV in water

reactors, at 7 (circles), 20

(squares) or 37 �C (crosses),

placed at dark. Data points

represent mean values obtained

from duplicate experiments,

each quantified in triplicate

Mineral water

0 10 20 30

0

2

4

6

8

Time (h)

Lo

g 10(N

t/N

0)

Wastewater

0 10 20 30

0

2

4

6

8

Time (h)

Lo

g10

(Nt/N

0)

Wastewater 1/1000

0 10 20 30

0

2

4

6

8

Time (h)

Lo

g 10

(Nt/N

0)

Seawater

0 10 20 30

0

2

4

6

8

Time (h)

Lo

g10

(Nt/N

0)

Fig. 2 Effect of UVB on the

stability of HAdV on water

reactors. Data points represent

mean values obtained from

duplicate experiments, each

quantified in triplicate


123

Effect of UVB Radiation on HAdV2 Stability in Water

To evaluate the effect of direct photo-inactivation by UVB

radiation on HAdV inactivation, duplicate experiments

were performed at 7 �C with reactors irradiated and reac-

tors hold in the dark in parallel, as controls. At 7 �C the

emission in the UVB wavelength reached a fluence of

122.91 J/m2 by the end of the experiment. Under these

experimental conditions, HAdV were rapidly inactivated in

all reactors. As shown in Fig. 2, the decay observed in

these experiments was of *4-log after 5 h of exposure to

UVB. At the end of the experiment, a complete decay

between 5 and 6-logs was achieved in all reactors, inde-

pendently on the water type tested. On the contrary, no

HAdV2 inactivation was observed in the dark in any of the

water types analyzed.

The complete inactivation of the initial viral concen-

trations was close to the detection limit of the infectivity

assay used. Consequently, a plateau was shown in the

inactivation kinetics from 8 h of exposition to UVB until

the end of the experiment (Fig. 2). Therefore, data could

not be adjusted to linear regression but to monophasic and

biphasic exponential models as detailed above. The cor-

relation values (R2), inactivation times (T90, T99, T99.9 and

T99.99) and inactivation rate constants (k) for these exper-

iments were obtained with a monophasic exponential

model and are further detailed in Table 4. In particular,

HAdV2 inactivation rate constants in PBS (k & 0.14 h-1)

were significantly different from those obtained in the rest

of studied reactors, k & 0.29 h-1 in mineral water,

0.24 h-1 in wastewater and seawater and 0.26 h-1 in

diluted wastewater.

Mineral water

0 10 20 30

0

2

4

6

8

Time (h)Lo

g 10(N

t/N0)

Wastewater

0 10 20 30

0

2

4

6

8

Time (h)

Log 10

(Nt/N

0)

L og 10

(Nt/N

0)Lo

g 10(N

t/N0)

Wastewater 1/1000

0 10 20 30

0

2

4

6

8

Time (h)

Seawater

0 10 20 30

0

2

4

6

8

Time (h)

Fig. 3 Effect of UVA and

temperature on the stability of

HAdV on water reactors at

20 �C (squares) or 37 �C

(crosses). Data points represent

mean values obtained from

duplicate experiments, each

quantified in triplicate

Table 2 Characterization of the water types used in the experiments, measured at room temperature. Light absorbance measures correspond to

mean values obtained from different wavelengths in each light spectrum range

Water type PH Conductivity

(mS)

Turbidity

(NTU)

Light absorbance

UVB

(290–315 nm)

UVA

(315–400 nm)

Visible

(400–700 nm)

Infrared

(700 nm–1 mm)

PBS 7.35 15.29 0.25 0.0006 0.0010 0.0029 0.0027

Mineral water 7.57 0.28 0.15 0.0062 0.0044 0.0041 0.0056

Raw sewage 7.64 3.56 142 0.0156 0.0061 0.0031 0.0032

Sea water 8.8 54.5 3.69 0.0036 0.0024 0.0028 0.0032


123

From the obtained experimental data, under the expo-

sition to UVB, the T99.99 was 14.72 h in PBS and 5.52 h in

mineral water. In wastewater the T99.99 was 5.41 h, and

4.90 and 5.55 h in diluted wastewater and seawater,

respectively.

Effect of UVA on HAdV2 Inactivation in Water

At 20 and 37 �C, the studied reactors were exclusively

exposed to UVA radiation with similar fluence at both

temperatures after 24 h (24.875 and 26.957 J/m2, respec-

tively). Dark controls were conducted in parallel. These

experimental conditions were performed in order to char-

acterize the relative contribution of UVA on viral inacti-

vation by indirect photo-inactivation mechanisms in the

absence of direct damage caused by UVB radiation.

At 20 �C, as presented in Fig. 3, decays of around 2

logs were achieved at the end of the experiment in all

reactors (PBS data not shown). The inactivation param-

eters (R2 inactivation rate constants and inactivation

times) were obtained by linear regression and are

detailed in Table 5. In particular, the inactivation rate

constant for wastewater, diluted wastewater, and seawa-

ter were similar (k & 0.11, 0.06, and 0.11 h-1, respec-

tively). The T99.99 value was not achieved in PBS and in

mineral water under these conditions and was 266.30 h

in wastewater, 63.87 h in diluted wastewater and 32.87 h

in seawater reactors.

Interestingly, as observed in Fig. 3, at 37 �C a signifi-

cant inactivation of HAdV was observed in wastewater,

diluted wastewater, and seawater reactors. In reactors

prepared with PBS (data not shown) and mineral water,

HAdV inactivation was similar to what’s observed from

results obtained at 20 �C (2-log for both matrices). How-

ever, in wastewater reactors the decays observed were

around 6-log. In the reactors prepared with wastewater

diluted in mineral water, a 4-log decay was reached after

24 h. A 3-log decay was achieved in seawater reactors at

both 20 and 37 �C. The inactivation parameters for these

experiments are presented in Table 3, as compared to those

obtained in the dark controls. The inactivation rate con-

stants were similar between all matrices with higher

microbial and organic content; wastewater (k & 0.22 ±

0.08 h-1), diluted wastewater in mineral water

(k & 0.15 ± 0.03 h-1) and seawater (k & 0.12 ±

0.04 h-1). Furthermore, these values are generally equiv-

alent to those observed in the dark controls, except in

reactors containing diluted wastewater (k & 0.15 ±

0.03 h-1 in reactors exposed to UVA, while k & 0.24 ±

0.05 h-1 at dark). At 37 �C in reactors exposed to UVA

radiation, the T99.99 was 16.34 h in wastewater, 27.29 h in

diluted wastewater, and 29.94 h in seawater.Ta

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.58

14

.00

–2

5.8

9T

99

13

.68

11

.05

–1

6.8

2

T99.9

26

.29

20

.41

–3

6.8

3T

99.9

20

.48

17

.26

–2

4.8

6

T99.9

93

4.0

92

6.5

2–

48

.07

T99.9

92

7.2

92

3.1

9–

33

.15

Sea

wat

er6

.92

91

05

Lin

ear

reg

ress

ion

0.7

90

.07

–0

.18

0.1

3T

90

8.1

42

.69

–1

3.8

9L

inea

rre

gre

ssio

n0

.88

0.1

20

.88

T90

5.6

31

.52

–9

.35

T99

16

.02

10

.74

–2

4.8

8T

99

13

.73

9.9

8–

18

.72

T99.9

23

.89

17

.35

–3

7.3

8T

99.9

21

.84

17

.16

–2

9.3

7

T99.9

93

1.7

62

3.4

0–

50

.43

T99.9

92

9.9

42

3.8

1–

40

.55


123

Discussion

Our main objective in this study was to characterize the

relative contribution of temperature and UV radiation from

sunlight as major effectors on the inactivation of viral

pathogens and consequently, in their potential dissemina-

tion in water environments, predicting quantitative

inactivation parameters for a decay of 4-log under each

experimental condition.

While high temperatures are known to inhibit host-cell

recognition or viral binding (Wigginton et al. 2012), solar

inactivation occurs via three distinct mechanisms due to the

action of UV radiation on viral particles: directly through

genome damage (direct photo-inactivation), and indirectly

Table 4 Inactivation parameters of human adenoviruses in water reactors exposed to UVB radiation at 7 �C

Water reactor Initial concentration

(FFU/ml)

Best fitting

model

R2 k (h-1) 95 % CI Inactivation

times (h)

95 % CI

PBS 5.25 9 105 Monophasic 0.82 0.14 0–0.29 T90 1.91 0–4.06

T99 4.25 2.29–7.99

T99.9 7.73 4.38–15.66

T99.99 14.72 7.47–NA

Mineral water 7.7 9 105 Monophasic 0.71 0.29 0–0.67 T90 0.98 0–2.87

T99 1.95 0–4.58

T99.9 3.29 1.34–4.38

T99.99 5.52 2.31–NA

Wastewater 4.75 9 105 Monophasic 0.78 0.24 0–0.51 T90 0.94 0–2.43

T99 1.97 0.27–4.19

T99.9 3.34 1.62–6.52

T99.99 5.41 2.73–NA

Wastewater (1/1,000) 5.45 9 105 Monophasic 0.82 0.26 0–0.51 T90 0.81 0–2.01

T99 1.77 0.43–3.53

T99.9 3.03 1.56–5.55

T99.99 4.90 2.64–10.33

Seawater 3.2 9 105 Monophasic 0.80 0.24 0–0.50 T90 0.93 0–2.32

T99 1.98 0.50–4.05

T99.9 3.39 1.72–6.39

T99.99 5.55 2.88–NA

NA not achieved

Table 5 Inactivation parameters of human adenoviruses in water reactors exposed to UVA radiation at 20 �C

Water reactor Initial concentration

(FFU/ml)

Best fitting

model

R2 K (h-1) 95 % CI Inactivation

times (h)

95 % CI

Wastewater 3.05 9 105 Linear regression 0.85 0.11 0.07–0.15 T90 62.34 5.31–14.44

T99 130.33 13.84–26.30

T99.9 198.31 21.22–39.30

T99.99 266.30 28.23–52.68

Wastewater (1/1,000) 3.47 9 105 Linear regression 0.79 0.06 0.03–0.09 T90 13.56 8.99–22.31

T99 30.33 21.79–52.02

T99.9 47.10 33.60–82.73

T99.99 63.87 45.23–113.60

Seawater 4.45 9 105 Linear regression 0.94 0.11 0.09–0.14 T90 6.66 3.95–9.27

T99 15.40 12.60–18.94

T99.9 24.14 20.38–29.46

T99.99 32.87 27.85–40.29


123

by endogenous or exogenous photo-inactivation (Davies-

Colley et al. 1999, 2000; Kohn and Nelson 2007; Romero

et al. 2011). Sunlight wavelengths in the range of UVC

(100–280 nm) are extinct before reaching surface and

coastal waters. Consequently, direct damage caused by

sunlight is mainly due to the effect of UVB (wavelengths

of 280–320 nm) and derived photoproducts such as

pyrimidine dimers (Schuch and Menck 2010) that may

block genome replication. Both UV and visible sunlight

may initiate indirect photo-inactivation mechanisms that

are based on the excitation of sensitizer compounds in oxic

conditions. This reaction represents a source of reactive

oxygen species that can inactivate viral particles by various

mechanisms. Sensitizers may be located within the

microorganisms and lead to endogenous inactivation, or

outside the cell in the water, in which case they contribute

to exogenous inactivation (Kohn and Nelson 2007).

In order to evaluate the specific contribution of tem-

perature to the inactivation of HAdV in water, experiments

were performed at three distinct temperatures in the dark.

Viruses are known to be more stable at low temperatures

and we selected 7, 20, and 37 �C as representative tem-

peratures of different climatological conditions. Our results

showed, as expected, that low temperatures in water are

optimal for viral persistence. No relevant inactivation was

observed after 24 h at 7 and 20 �C in any of the water types

tested. However, at 37 �C, a linear reduction of between 4

and 6-log was identified in wastewater, diluted wastewater,

and seawater reactors, containing a higher microbial con-

centration as compared to PBS and mineral water. As

shown in other studies, this reduction is probably caused by

the microorganisms naturally present in the samples that

may grow or become more active at higher temperatures

(Fujioka et al. 1980; Girones et al. 1989a, b; Gordon and

Toze 2003). Interestingly, Ward et al. (1986) showed that,

at 27 �C, echovirus inactivation in freshwater samples was

caused by proteolytic bacterial enzymes cleaving viral

proteins VP4 and VP1. In consistence with our data, the

stability of enteric viruses in water at low temperatures has

been described in several matrices (Enriquez et al. 1995;

Raphael et al. 1985; reviewed by Rzezutka and Cook

2004). One significant conclusion of these experiments is

that the inactivation of viruses in water at warm tempera-

tures (37 �C), increases with the presence of a natural

bacterial flora (seawater) and especially with the presence

of urban sewage.

The direct effect of UVB radiation on viral inactivation

in water was studied by monitoring the persistence of

infectious HAdV2 during 24 h, corresponding to an accu-

mulated fluence of 122.91 J/m2. The experiments per-

formed in parallel in the dark were used as controls.

Currently, there is strong evidence that direct damage due

to UVB radiation present in solar light is highly effective

for the inactivation of enteric viruses and other fecal

indicators (Sinton et al. 2002; Kohn and Nelson 2007; Love

et al. 2010). The results obtained in our experiments are

consistent with these observations. At 7 �C, under the

exposition to UVB radiation, HAdV decayed equally in all

reactors achieving an inactivation of 5-log within 8 h of

experiment. The decay observed in these experiments was

approximately equal to the complete inactivation of the

initial viral concentration in the reactors and close to the

detection limit of our immunofluorescence assays. From

our results, UVB inactivation of viruses is not dependent

on the water type tested in the conditions of the assays,

however the characteristics of the water matrix will have an

important influence on the level of UVB penetration in the

water column. The equation for the quantification of light

screening may be used to estimate the potential effect of

UVB on the viruses in a natural system, calculating down

to which depth there may be any significant UVB light.

This approach may be informative about how deep the

photoactive zone is. It is known that the intensity and

spectral quality of UV radiation reaching the Earth’s sur-

face vary according to atmospheric conditions and solar

geometry which is function of the time of the day, the

season or the geographical position. There are different

studies analyzing the UV-B penetration in aquatic eco-

systems (Huovinen et al. 2003; Kirk 1994; Whitehead et al.

2004). Generally, the attenuation of UV-B radiation is

considered to be more efficient in lakes than in oceans. The

UV-B penetration depths may vary from only few centi-

meters in highly humic lakes to dozens of meters in the

oceans (Kirk 1994). It has been described that dissolved

organic carbon (DOC) strongly governs the UV attenuation

in humic lakes, whereas in oceans and in clear lakes with

low DOC concentration the contribution of phytoplankton

to UV attenuation can be significant (Huovinen et al.

2003). The deep of 90 % attenuation of UVB (305 nm)

radiation in different water bodies present then large dif-

ferences and could be from more than 50 meters deep to

few centimeters (Whitehead et al. 2004). As mentioned

above, no inactivation was observed in the dark. These

results show that when UVB radiation is present, particu-

larly in shallow water layers, the inactivation of HAdV is

mainly caused by direct damage of UVB to viral particles

and that indirect photo-inactivation may be negligible. In

fact, similar findings have been reported in other studies

where UVB was the main responsible for the observed

inactivation of MS2, rotavirus, adenovirus, and poliovirus,

in water exposed to the full sunlight spectrum (Kohn and

Nelson 2007; Love et al. 2010; Romero et al. 2011).

In our experiments, the indirect photo-inactivation

damage caused by UVA was investigated in experiments

performed with different water types at two temperatures

(20 and 37 �C). These conditions represent those found in


123

deeper layers of water environments where UVB radiation

may be extinct. Under these conditions, HAdV inactivation

followed a first-order kinetic in all reactors. At 20 �C in

reactors exposed to UVA radiation, a 2-log inactivation

was observed despite different physicochemical and bio-

logical characteristics between the water types used in the

experiments. Wastewater and seawater have a higher

microbial content than PBS and mineral water, and a sig-

nificant presence of organic matter containing substances

that may act as sensitizers in exogenous indirect photo-

inactivation processes. Hence, in these reactors, photo-

inactivation may be explained by both endogenous and

exogenous mechanisms. However, at 20 �C the inactiva-

tion of HAdV in wastewater, diluted wastewater and sea-

water reactors was similar to that observed in clear mineral

water, achieving approximately a 2-log decay in all cases.

These results do not elucidate the relative contribution of

endogenous and exogenous indirect photo-inactivation and

further studies are needed to shed light on this question.

The experiments performed at 37 �C under UVA

exposition allowed us to determine the potential viral

inactivation caused by microbial activity and to evaluate its

relative contribution as compared to indirect photo-inacti-

vation by UVA radiation. From the obtained results,

HAdV2 inactivation was 2-log in PBS and mineral water

while in complex matrices achieved 4-log in diluted

wastewater, 6-log in raw wastewater, and 3-log in seawater

reactors. The indirect photo-inactivation caused by UVA is

clearly identified in our experiments conducted at 20 �C

and at 37 �C in PBS and mineral water reactors. However,

at 37 �C, in all reactors with higher natural microbial

concentration, the effect of biotic factors on the inactiva-

tion of infectious HAdV overcomes the effect of indirect

photo-inactivation mechanisms caused by UVA.

Overall, our results show that the natural inactivation of

HAdV in water and by extension of other viral pathogens is

a highly complex process that may be governed by a wide

range of interacting factors. As expected, UVB radiation is

related to a major direct viral damage and under the studied

conditions was mainly responsible of viral inactivation

independently of the water physicochemical or biologic

characteristics. Second, as previously described our results

support the importance of biotic factors in viral inactivation

that at higher temperatures in matrices with high microbial

content are more relevant than indirect photo-inactivation

caused by UVA radiation in water layers where UVB is

extinguished. The quantitative parameters generated are

representative of the time needed for the efficient disin-

fection of HAdV. In water matrices at 7 �C when solar

irradiation and UVB are present, T99.99 values were

4.90–5.55 h. High temperatures have been identified as a

significant factor enhanced by the presence of biotic fac-

tors. In the water matrices tested at 37 �C from 16.34 h

(wastewater) to 29.94 h (diluted wastewater 1/1,000) are

needed for 4 log reduction (T99.99) of HAdV2 in the

developed experiments.

In general, our results indicate that viruses may persist

for longer periods in simple water matrices such as those

used as drinking sources or for irrigation, especially at

night or when UVB or UVA radiation are not present, and

at temperatures up to 20 �C. Therefore, outbreaks initially

establish during the colder months by contamination events

in water distribution systems should be of particular con-

cern regarding viral persistence and their potential

dissemination.

Despite the relevance of water in the transmission of

pathogens worldwide, few efforts have been made to

characterize the relative contribution of environmental

agents threating viral persistence and thus, determining

their dissemination. A detailed characterization of the

inactivation mechanisms driving viral inactivation is nee-

ded to achieve sustainable improvements in water safety by

developing efficient prediction tools and management

strategies to remediate waterborne outbreaks. The inacti-

vation parameters calculated here will be useful to develop

mathematical models to predict viral dissemination in

different water environments.

Acknowledgments This work was partially supported by the Euro-

pean Commission Framework Program 7 project ‘‘Integrated moni-

toring and control of foodborne viruses in European food supply chains

(VITAL)’’ (Grant No. KBBE 213178) led by the coordination team of

Nigel Cook (FERA, UK), Martin D’Agostino (FERA, UK), and Franco

M Ruggeri (ISS, Italy), and partly by the European project, VIROC-

LIME (Grant No. 243923), coordinated by David Kay and Peter Wyn-

Jones (University of Aberystwyth, UK). During the development of the

study Anna Carratala was the recipient of a fellowship from the Spanish

Ministry of Science and Innovation and Marta Rusinol was the recipient

of a fellowship from the Catalan Government ‘‘AGAUR’’ (FI-DGR).

We thank the Serveis Cientıfico-Tecnics and the Camps Experimentals

of the University of Barcelona for their kind support during the

experiments. We are grateful to Tamar Kohn (Ecole Polytechnique

Federale de Lausanne) and Miquel Calvo (Universitat de Barcelona) for

interesting discussions during the study.

References

Albinana-Gimenez, N., Miagostovich, M., Calgua, B., Huguet, J. M.,

Matia, L., & Girones, R. (2009). Analysis of adenoviruses and

polyomaviruses quantified by qPCR as indicators of water

quality in source and drinking water-treatment plants. Water

Research, 43, 2011–2019.

Bertrand, I., Schijven, J., Sanchez, G., Wyn-Jones, P., Ottoson, J.,

Morin, T., et al. (2012). The impact of temperature on the

inactivation of enteric viruses in food and water: a review.

Journal of Applied Microbiology, 112, 1059–1074.

Boehm, A. B., Yamahara, K. M., Love, D. C., Peterson, B. M.,

McNeill, K., & Nelson, K. L. (2009). Covariation and photo-

inactivation of traditional and novel indicator organisms and

human viruses at a sewage-impacted marine beach. Environ-

mental Science and Technology, 43(21), 8046–8052.


123

Bofill-Mas, S., Albinana-Gimenez, N., Clemente-Casares, P., Hun-

desa, A., Rodriguez-Manzano, J., Allard, A., et al. (2006).

Quantification and stability of human adenoviruses and polyo-

mavirus JCPyV in wastewater matrices. Applied and Environ-

mental Microbiology, 72(12), 7894–7896.

Calgua, B., Barardi, C. R., Bofill-Mas, S., Rodriguez-Manzano, J., &

Girones, R. (2011). Detection and quantitation of infection human

adenoviruses and JC polyomaviruses in water by immunofluores-

cence assay. Journal of Virological Methods, 171(1), 1–7.

Carratala, A., Rodriguez-Manzano, J., Hundesa, A., Rusinol, M.,

Fresno, S., Cook, N., et al. (2013). Effect of temperature and

sunlight on the stability of human adenoviruses and MS2 as fecal

contaminants on fresh produce surfaces. International Journal of

Food Microbiology, 164(2–3), 128–134.

Carter, M. J. (2005). Enterically infecting viruses: pathogenicity,

transmission and significance for food and waterborne infection.

Journal of Applied Microbiology, 98, 1354–1380.

Chick, H. (1910). The process of disinfection by chemical agencies and

hot water. The Journal of Hygiene (London), 10(2), 237–286.

Davies-Colley, R. J., Donnison, A. M., & Speed, D. J. (2000).

Towards a mechanistic understanding of pond disinfection.

Water Science and Technology, 42, 149–158.

Davies-Colley, R. J., Donnison, A. M., Speed, D. J., Ross, C. M., &

Nagels, J. W. (1999). Inactivation of faecal indicator microor-

ganisms in waste stabilization ponds: interactions of environ-

mental factors with sunlight. Water Research, 33, 1220–1230.

de Roda Husman, A. M., Lodder, W. J., Rutjes, S. A., Schijven, J. F.,

& Teunis, P. F. (2009). Long-term inactivation study of three

enteroviruses in artificial surface and groundwaters, using PCR

and cell culture. Applied and Environmental Microbiology,

75(4), 1050–1057.

Dick, L. K., Stelzer, E. A., Bertke, E. E., Fong, D. L., & Stoeckel, D.

M. (2010). Relative decay of Bacteroidales microbial source

tracking markers and cultivated Escherichia coli in freshwater

microcosms. Applied and Environmental Microbiology, 76(10),

3255–3262.

Enriquez, C. E., Hurst, C. J., & Gerba, C. P. (1995). Survival of the

enteric adenoviruses 40 and 41 in tap, sea, and wastewater.

Water Research, 29(11), 2548–2553.

Fong, T. T., Phanikumar, M. S., Xagoraraki, I., & Rose, J. B. (2010).

Quantitative detection of human adenoviruses in wastewater and

combined sewer overflows influencing a Michigan river. Applied

and Environmental Microbiology, 76(3), 715–723.

Formiga-Cruz, M., Allard, A. K., Conden-Hansson, A. C., Henshil-

wood, K., Hernroth, B. E., Jofre, J., et al. (2003). Evaluation of

potential indicators of viral contamination in shellfish and their

applicability to diverse geographical areas. Applied and Envi-

ronmental Microbiology, 69(3), 1556–1563.

Fujioka, R. S., Loh, P. C., & Lau, L. S. (1980). Survival of human

enteroviruses in the Hawaiian ocean environment: evidence for

virus-inactivating microorganisms. Applied and Environmental

Microbiology, 39(6), 1105–1110.

Gerba, C. P., Goyal, S. M., LaBelle, R. L., Cech, I., & Bodgan, G. F.

(1979). Failure of indicator bacteria to reflect the occurrence of

enteroviruses in marine waters. American Journal of Public

Health, 69, 1116–1119.

Girones, R., Jofre, J., & Bosch, A. (1989a). Isolation of marine

bacteria with antiviral properties. Canadian Journal of Micro-

biology, 35(11), 1015–1021.

Girones, R., Jofre, J., & Bosch, A. (1989b). Natural inactivation of

enteric viruses in seawater. Journal of Environmental Quality,

18(1), 34–39.

Gordon, C., & Toze, S. (2003). Influence of groundwater character-

istics on the survival of enteric viruses. Journal of Applied

Microbiology, 95, 536–544.

Huovinen, P. S., Penttil, H., & Soimasuo, M. R. (2003). Spectral

attenuation of solar ultraviolet radiation in humic lakes in

Central Finland. Chemosphere, 51, 205–214.

Hurvich, C. M., & Tsai, C. L. (1989). Regression and time series

model selection in small samples. Biometrika, 76(2), 297–307.

Kirk, J. T. O. (1994). Optics of UV-B radiation in natural waters.

Archiv fur Hydrobiologie, 43, 1–16.

Kohn, T., & Nelson, K. L. (2007). Sunlight-mediated inactivation of

MS2 coliphage via exogenous singlet oxygen produced by

sensitizers in natural waters. Environmental Science and Tech-

nology, 41(1), 192–197.

Kokkinos, P. A., Ziros, P. G., Mpalasopoulou, A., Galanis, A., &

Vantarakis, A. (2011). Molecular detection of multiple viral

targets in untreated urban sewage from Greece. Virology

Journal, 27(8), 195.

Kovac, K., Bouwknegt, M., Diez-Valcarce, M., Raspor, P., Hernan-

dez, M., & Rodrıguez-Lazaro, D. (2012). Evaluation of high

hydrostatic pressure effect on human adenovirus using molecular

methods and cell culture. International Journal of Food

Microbiology, 157(3), 368–374.

Lipp, E. K., Farrah, S. A., & Rose, J. B. (2001). Assessment and

impact of microbial fecal pollution and human enteric pathogens

in a coastal community. Marine Pollution Bulletin, 42, 286–293.

Love, D. C., Silverman, A., & Nelson, K. L. (2010). Human virus and

bacteriophage inactivation in clear water by simulated sunlight

compared to bacteriophage inactivation at a southern California

beach. Environmental Science and Technology, 44(18),

6965–6970.

McGuigan, K. G., Conroy, R. M., Mosler, H. J., du Preez, M.,

Ubomba-Jaswa, E., & Fernandez-Ibanez, P. (2012). Solar water

disinfection (SODIS): a review from bench-top to roof-top.

Journal of Hazardous Materials, 235–236, 29–46.

McKinney, K. R., Gong, Y. Y., & Lewis, T. G. (2006). Environmental

transmission of SARS at Amoy Gardens. Journal of Environ-

mental Health, 68, 26–30.

Nenonen, N. P., Hannoun, C., Larsson, C. U., & Bergstrom, T.

(2012). Marked genomic diversity of norovirus genogroup I

strains in a waterborne outbreak. Applied and Environmental

Microbiology, 78(6), 1846–1852.

Pina, S., Puig, M., Lucena, F., Jofre, J., & Girones, R. (1998). Viral

pollution in the environment and in shellfish: human adenovirus

detection by PCR as an index of human viruses. Applied and

Environmental Microbiology, 64, 3376–3382.

Raphael, R. A., Sattar, S. A., & Springthorpe, V. S. (1985). Long-term

survival of human rotavirus in raw and treated river water.

Canadian Journal of Microbiology, 31(2), 124–128.

Riera-Montes, M., Brus Sjolander, K., Allestam, G., Hallin, E.,

Hedlund, K. O., & Lofdahl, M. (2011). Waterborne norovirus

outbreak in a municipal drinking-water supply in Sweden.

Epidemiology and Infection, 139(12), 1928–1935.

Rigotto, C., Victoria, M., Moresco, V., Kolesnikovas, C. K., Correa,

A. A., Souza, D. S., et al. (2010). Assessment of adenovirus,

hepatitis A virus and rotavirus presence in environmental

samples in Florianopolis, South Brazil. Journal of Applied

Microbiology, 109(6), 1979–1987.

Rodriguez-Manzano, J., Alonso, J. L., Ferrus, M. A., Moreno, Y.,

Amoros, I., Calgua, B., et al. (2012). Standard and new faecal

indicators and pathogens in sewage treatment plants, microbi-

ological parameters for improving the control of reclaimed

water. Water Science and Technology, 66(12), 2517–2523.

Romero, O. C., Straub, A. P., Kohn, T., & Nguyen, T. H. (2011). Role

of temperature and Suwannee River natural organic matter on

inactivation kinetics of rotavirus and bacteriophage MS2 by

solar irradiation. Environmental Science and Technology,

45(24), 10385–10393.


123

Rzezutka, A., & Cook, N. (2004). Survival of human enteric viruses

in the environment and food. FEMS Microbiology Reviews,

28(4), 441–453.

Schuch, A. P., & Menck, C. F. (2010). The genotoxic effects of DNA

lesions induced by artificial UV-radiation and sunlight. Journal

of Photochemistry and Photobiology B, 99(3), 111–116.

Schwarzenbach, R. P., Gschwend, P. M., & Imboden, D. M. (2003).

Environmental organic chemistry. Hoboken: Wiley.

Simmons, F. J., & Xagoraraki, I. (2011). Release of infectious human

enteric viruses by full-scale wastewater utilities. Water

Research, 45(12), 3590–3598.

Sinton, L. W., Hall, C. H., Lynch, P. A., & Davies-Colley, R. J.

(2002). Sunlight inactivation of fecal indicator bacteria and

bacteriophages from waste stabilization pond effluent in fresh

and saline waters. Applied and Environmental Microbiology,

68(3), 1122–1131.

Sirikanchana, K., Shisler, J. L., & Marinas, B. J. (2008). Inactivation

kinetics of adenovirus serotype 2 with monochloramine. Water

Research, 42(6–7), 1467–1474.

ter Waarbeek, H. L., Dukers-Muijrers, N. H., Vennema, H., & Hoebe,

C. J. (2010). Waterborne gastroenteritis outbreak at a scouting

camp caused by two norovirus genogroups: GI and GII. Journal

of Clinical Virology, 47(3), 268–272.

Tree, J. A., Adams, M. R., & Lees, D. N. (2003). Chlorination of

indicator bacteria and viruses in primary sewage effluent.

Applied and Environmental Microbiology, 69(4), 2038–2043.

Tuladhar, E., Bouwknegt, M., Zwietering, M. H., Koopmans, M., &

Duizer, E. (2012). Thermal stability of structurally different

viruses with proven or potential relevance to food safety. Journal

of Applied Microbiology, 112(5), 1050–1057.

Verhaelen, K., Bouwknegt, M., Lodder-Verschoor, F., Rutjes, S. A.,

& de Roda Husman, A. M. (2012). Persistence of human

norovirus GII.4 and GI.4, murine norovirus, and human adeno-

virus on soft berries as compared with PBS at commonly applied

storage conditions. International Journal of Food Microbiology,

160(2), 137–144.

Viau, E. J., Goodwin, K. D., Yamahara, K. M., Layton, B. A.,

Sassoubre, L. M., Burns, S. L., et al. (2011). Bacterial pathogens in

Hawaiian coastal streams–associations with fecal indicators, land

cover, and water quality. Water Research, 45(11), 3279–3290.

Ward, R. L., Knowlton, D. R., & Winston, P. E. (1986). Mechanism

of inactivation of enteric viruses in fresh water. Applied and

Environmental Microbiology, 52(3), 450–459.

Whitehead, R. F., de Mora, S. J., & Demers, S. (2004). Enhanced UV

radiation-a new problem for the marine environment. In S. J. De

Mora, S. Demers, & M. Vernet (Eds.), The effects of UV

radiation in the marine environment (pp. 1–34). Cambridge:

University Press. ISBN 0511-03586-1.

Wigginton, K. R., Pecson, B. M., Sigstam, T., Bosshard, F., & Kohn,

T. (2012). Virus inactivation mechanisms: impact of disinfec-

tants on virus function and structural integrity. Environmental

Science and Technology, 46(21), 12069–12078.


123

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