Gas emission and wetlands - IMT Atlantique of... · Several investigations have shown that N 2 O and CH 4 emissions have high variability, staggering between -0.07 – 1000 mg m-2

ABSTRACTS - WETPOL 2013 - October 13-17, 2013 - Nantes - FRANCE

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Gas emission and wetlands


108

Methane and Nitrous Oxide Emissions from Constructed

Wetlands for Municipal Wastewater Treatment (O.7)

Kuno Kasaka, Ülo Mander

a

aDepartment of Geography, Institute of Ecology and Earth Sciences, University of Tartu,

Vanemuise 46, Tartu, 51014, ESTONIA. (E-mail: [email protected], [email protected])

INTRODUCTION

Constructed wetlands (CW) are increasingly used for water pollution treatment; as well as

natural wetlands, CW can also be a source of three important greenhouse gases (GHG):

carbon dioxide (CO2), methane (CH4), and nitrous oxide (N2O; Mander et al., 2008).

Constructed wetlands are widely investigated for wastewater treatment, however relatively

few studies have been carried out concerning N2O and CH4 fluxes from CW, especially for

municipal wastewater treatment.

Several investigations have shown that N2O and CH4 emissions have high variability,

staggering between -0.07 – 1000 mg m-2

h-1

and -32 – 43000 mg m-2

h-1

, respectively

(Mander et al., 2008; Søvik et al., 2006). The main objective of this research were to quantify

N2O and CH4 emission rates from horizontal subsurface flow constructed wetlands for

municipal wastewater treatment using the static closed chamber method and alternative filter

materials.

METHODS

The study system was located in the territory of the activated sludge treatment plant of

Nõo village in Estonia. Small amount of wastewater was pumped into CW before it reaches

the wastewater pre-filtration grid of the treatment plant. CW was divided into 5 parallel

systems (width 0.6 m, depth 0.6 m, length 1.5 m; each). Each parallel system was filled with

different combination of well mineralized peat and hydrated oil shale ash (industrial by-

product from Estonian thermal power plants, with high pH and Ca content): HF1 – peat (60

cm), HF2 – peat (50 cm) covered with 10 cm ash, HF3 – peat (40 cm) covered with 20 cm

ash, HF4 – peat (30 cm) covered with 30 cm ash, and HF5 – ash (60 cm). The fluxes of N2O

and CH4 were determined using the closed chamber method. Gas concentrations were

measured using white (to avoid heating during application) PVC chambers (diameter 50 cm,

height 50 cm, and volume 65 L) sealed with a water-filled ring on the surface. Ten replicate

chambers were installed 5 (of them) to the inflow part and 5 (of them) to the outflow part).

First gas sampling was made in January (2012) and henceforward once a month from April

2012 to November 2012. Due to snow cover, no field measurements were carried out from

February 2012 to March 2012. Gas samples were taken 5 times (at the beginning, after 10

min, after 20 min, after 40 min, and after 60 min) from the enclosure of samplers using

previously evacuated (0.3 mbar) gas bottles (100 mL).

Kolmogorov-Smirnov, Lilliefors’ and Shaphiro-Wilks’ tests were used to check normality

of variables. In all cases the gaseous distribution differed from the normal, hence the non-

parametric Wilcoxon Matched Pairs Test was used for statistical analysis.

RESULTS AND DISCUSSIONS

The median emission values of N2O-N and CH4-C from inflow part ranged between 21.0

to 72.3 μg m-2

h-1

and 16.4 to 556.2 μg m-2

h-1

, respectively (Fig 1.). The median emission

values of N2O-N and CH4-C from outflow part ranged between 12.9 to 40.2 μg m-2

h-1

to 11.0

to 237.3 μg m-2

h-1

, respectively (Fig 1.). Median gaseous fluxes from CWs were remarkably

lower compared with other studies (Mander et al., 2008; Søvik et al., 2006). According to

mailto:[email protected]



109

Le Mer & Roger, (2001) the activity of methanogens is usually optimum around neutrality or

under facile alkaline conditions as presented in filter systems HF2 and HF3 (median pH

around 8-9). Filter systems HF4 and HF5 had high pH (12-12.3) and Ca2+

concentration;

therefore, showing much lower CH4 emissions (Fig. 1), low CH4 emissions from peat filter

(HF1) was due to low pH. Low N2O emissions can be explained by the low Ntot removal from

all filter systems, by the high pH, and by the static water regime (Mander et al., 2011). The

highest N2O emissions originate from filter systems HF1 and HF4, where the Ntot removal

was highest.

Fig. 1. Median, 25%, and 75% percentiles, and min-max values of N2O and CH4 emissions from

constructed wetlands. Asterisks above the bars indicates significantly (p<0.05) differing values between

inflow and outflow. Letters above bars indicate significant differences (p<0.05) between parallel filter

systems.

CONCLUSIONS

In the CW, high pH and high Ca2+

content resulted in low methane emissions.

Methane emissions are highest in slightly alkaline conditions.

Nitrous oxide emissions from CW are affected by the low nitrogen removal, by

the high pH, and by the static water regime.

ACKNOWLEDGMENT

This study was supported by Ministry of Education and Science of Estonia grants nos.

SF0180127s08 and IUT13016, and the EU through the European Regional Development

Fund (Center of Excellence ENVIRON).

REFERENCES Le Mer, J., Roger, P. 2001. Production, oxidation, emission and consumption of methane by soils: A review.

Eur. J. Soil Biol. 37, 25-50

Mander, Ü., Lõhmus, K., Teiter, S., Mauring, T., Nurk, K., Augustin, J. 2008. Gaseous fluxes in the nitrogen

and carbon budgets of subsurface flow constructed wetlands. Science of the Totan Environment, 404, 343-353

Mander, Ü., Maddison, M., Soosaar, K., Karabelnik, K. 2011. The Impact of Pulsing Hydrology and Fluctuating

Water Table on greenhouse Gas Emissions from Constructed Wetlands. Wetlands 31:1023-1032, DOI

10.1007/s13157-011-0218-z

Søvik, A.K., Augustin, J., Heikkinen, K., Huttunen, J.T., Necki, J.M., Karjalainen, S.M., Kløve, A.K.,

Liikanen, A., Mander, U., Puustinen, M., Teiter, S., Wachniew, P.J., 2006. Emission of the greenhouse gases

nitrous oxide and methane from constructed wetlands in Europe. J. Environ. Qual, 2360-2373

HF1-IN

HF2-IN

HF3-IN

HF4-IN

HF5-IN

HF1-OUT

HF2-OUT

HF3-OUT

HF4-OUT

HF5-OUT

-100

0

100

200

300

400

500

600

700

μg N

2O

-N m

-2

h-1

Median 25%-75% Min-Max

*

*

HF4-IN

HF5-IN

HF2-IN

HF2-IN

HF1-IN

HF2-IN

HF3-IN

HF4-IN

HF5-IN

HF1-OUT

HF2-OUT

HF3-OUT

HF4-OUT

HF5-OUT

-2000

0

2000

4000

6000

8000

10000

12000

14000

16000

18000

20000

μg

CH

4-C

m-2

h

-1

Median 25%-75% Min-Max

HF2-IN

HF3-IN

HF1-IN

HF1-IN

HF2-OUT

HF3-OUT

HF1-OUT

HF1-OUT


110

Nitrous oxide emissions in natural and restored Danish wetlands (O.44)

Joachim Audeta, Carl C. Hoffmann

a, Peter M. Andersen

a,b, Annette Baattrup-

Pedersena, Jan R. Johansen

a, Søren E. Larsen

a, Linus Lind

a,c, Lars Elsgaard

b,

Charlotte Kjaergaardb, Karin Tonderski

c.

aAarhus University, Department of Bioscience, Vejlsøvej 25, 8600 Silkeborg, DENMARK

([email protected] – [email protected] – [email protected] – [email protected] –

[email protected] – [email protected] – [email protected]) bAarhus University, Department of Agroecology, P.O. Box 50, Blichers allé 20, 8830 Tjele,

DENMARK ([email protected] – [email protected]) cLinköping University, IFM Biology, section Ecology, 58183 Linköping, SWEDEN

([email protected])

INTRODUCTION

Riparian wetlands located in agricultural catchments are likely to receive groundwater rich

in nitrate (NO3-) derived from crop fertilizers. Nitrate can be removed in the riparian zones by

denitrification, i.e. the transformation from aqueous NO3- to the gaseous forms nitrous oxide

N2O and dinitrogen (N2). While the production of N2 has no harmful consequences on the

environment, the release of N2O to the atmosphere is more problematic as N2O is a potent

greenhouse gas. Hence there is some concern that riparian wetlands are hotspots of N2O

emissions in the landscape. Thus, in the context of wetland restoration, which is often used as

a way to mitigate aquatic pollution due to NO3-, the benefits obtained by improving the water

quality might be realised at the expense of air quality. Therefore to evaluate the risk of N2O

release from Danish riparian wetlands, several natural and restored sites were investigated.

We aimed at quantifying the fluxes of N2O and at identifying the controllers of these fluxes.

METHODS

We used static chambers to measure N2O fluxes 1) during a year at four natural riparian

wetlands each comprising three plots presenting contrasting characteristics regarding

groundwater level, water chemistry and soil characteristics; 2) height months before and one

year after restoration of a riparian wetland. Together with the gas monitoring, numerous

environmental variables such as dissolved N species, soil pH, and soil C:N ratio were

collected.

We also used laboratory microcosms to simulate an upward flow of NO3- enriched

groundwater through intact soil cores collected from four wetlands with contrasting soil

characteristics and measured the production of N2O consecutive to the reduction of NO3-.

RESULTS AND DISCUSSION

Individual fluxes of N2O in the natural sites ranged between -45 and 122 µg N2O-N m-2

h-1

while yearly fluxes ranged between -0.01 and 0.12 g N2O-N m-2

y-1

(Audet et al., submitted).

Modelling of the fluxes revealed a significant effect of ammonium in the soil water. Despite

these wetlands were located in catchments dominated by agricultural land use, these

measurements indicate that riparian wetlands were not a hotspot for N2O emissions.

But elevated fluxes (>2000 µg N2O-N m-2

h-1

) were recorded both before and after

restoration of a riparian wetland (Audet et al., 2013). No effect in the first year following the

restoration on N2O emissions could be demonstrated. However, the fluxes in the restoration

study appeared to be controlled by N content in the top soil and are therefore expected to

decrease as a consequence of increasing flooding and decreasing N availability related also to

the reduction of fertilization in the neighbouring fields.


111

Soil microcosms enabled to identify a significant positive effect of the NO3- load on N2O

emissions (Lind et al., 2013). The results showed a rapid reduction of the NO3- fluxes,

supporting the effectiveness of wetlands for removal of N. However, during the reduction of

NO3- transient accumulation of N2O was observed, but the N2O concentration decreased with

declining NO3- availability (<2 mg N L

-1). Still, in some cases the production of N2O was

substantial and accounted for 30% of the NO3- reduction. In this laboratory study, the NO3

-

load was revealed as the only significant factor controlling both NO3- reduction and N2O

production.

CONCLUSIONS

Overall, our results show that Danish riparian wetlands do not act as hotspot for N2O

emissions in the landscape. The restoration of riparian areas did not show any significant

effect on N2O emissions but only the first year after restoration was considered. Laboratory

experiments revealed that N2O production can be high but N2O is generally rapidly reduced

in the soil provided that NO3- concentration is decreasing. Hence, this research supports the

use of riparian wetlands as a way to mitigate aquatic N-pollution and show that the risk for

elevated N2O emissions is low.

ACKNOWLEDGEMENTS

This work is part of the MONITECH project supported by the Danish Council for

Strategic Research and by the Aarhus University Research Foundation.

REFERENCES Audet, J., Elsgaard, L., Kjaergaard, C., Larsen, S.E. and Hoffmann, C.C. (2013) Greenhouse gas emissions from

a Danish riparian wetland before and after restoration. Ecol. Eng. 57:170-182.

Audet, J., Hoffmann, C.C., Andersen, P.M., Baattrup-Pedersen, A., Johansen, J.R., Larsen, S.E., Kjaergaard, C.

and Elsgaard, L. (Submitted) Low fluxes of nitrous oxide in Danish riparian wetlands located in agricultural

catchments.

Lind, L.P.D., Audet, J., Tonderski, K. and Hoffmann, C.C. (2013). Nitrate removal capacity and nitrous oxide

production in soil profiles of nitrogen loaded riparian wetlands inferred by laboratory microcosms. Soil Biol.

Biochem. 60:156-164.


112

The impact of pulsing water table on wastewater purification and

greenhouse gas emission in a horizontal subsurface flow

constructed wetland (O.152)

Ülo Mander1,2

, Martin Maddison1, Kaido Soosaar

1, Helen Koger

1,

Alar Teemusk1, Reinhard Well

3

1 Institute of Ecology and Earth Sciences, University of Tartu, 51014 Tartu, Estonia

2 Hydrosystems and Bioprocesses Research Unit, National Research Institute of Science and

Technology for Environment and Agriculture (Irstea), 1 rue Pierre-Gilles de Gennes CS

10030, F92761 Antony cedex, France

3 Institut für Agrarrelevante Klimaforschung, Johann Heinrich von Thünen-Institut, 38116

Braunschweig, Germany

INTRODUCTION

A pulsing hydrological regime is often used in horizontal subsurface flow (HSSF)

constructed wetlands (CW) to enhance removal of BOD, COD, NH4 and total N (Vymazal

and Masa, 2003 WST) utilization by bacteria and to support more effective NH4, total N and

total P removal. Little is known of the impact of intermittent loading on greenhouse gas

(GHG) emissions from CWs. Several studies have shown that a pulsing regime decreases

both CH4 and N2O emissions from created riverine wetlands (Altor, Mitsch, 2008 Wetlands;

Hernández, Mitsch 2006 Wetlands). Mander et al (2011; Wetlands) have found that the lower

water table level in the HSSF bed of the Paistu-Sultsi hybrid CW in Estonia caused a

significant increase in CO2 and N2O emission and a decrease in CH4 emission. However, no

systematic experiments have been performed in HSSF CWs to study the impact of a

fluctuating water table.

The main aim of this paper was to estimate the impact of a fluctuating water table on the

water purification efficiency and GHG emissions in the HSSF bed of a hybrid CW in Sultsi-

Paistu, Estonia. To distinguish between denitrification and nitrification as source processes of

N2O, isotopologue studies have been conducted.

MATERIAL AND METHODS

The Sultsi-Paistu hybrid wetland system (constructed in 2002; 58o14’30.62’’N,

25o35’341.77’’E) treats the wastewater of 140 people (about 64 PE) and consists of a two-

chamber VSSF filter bed (12 m × 18 m) and a 216 m2 HSSF filter bed. The latter has a depth

of 0.9 m and is filled with 2–4 mm light-weight aggregates (LWA) and covered with reed

(see Öövel et al 2007 Ecol Eng for a detailed description). The whole system showed

outstanding purification effect: for BOD7 the average purification efficiency is 91%; for total

suspended solids (TSS) 78%, for total P 89%, for total N 63%, and for NH4-N 77% (Öövel et

al 2007).

In the period 2008-2010 the depth of the water table in the HSSF bed fluctuated from 0 to

70 cm. Since that period the water table has been kept constantly at a depth of 0-10 cm. In

October and November 2012, an experiment with fluctuating water table depth from 0-12

(high level) to 17-25 cm (low level) was conducted.

Gas fluxes were measured in 12 sampling sessions in 2008-2010 from the inflow, middle

and outflow parts of the HSSF bed, and in 16 sessions during the experimental period (7th

October to 7th

November 2012) from the inflow and outflow parts of the HSSF bed (5

replicates from each location). During the experimental period, water samples were taken

from the inflow and outflow of the HSSF once a week, during both high and low water table


113

level and analysed for pH, TOC, BOD7, NH4+-N, NO2

--N, NO3

--N, total N (TN), PO4

3--P and

total P (TP) in the lab of Estonian Environmental Research Ltd.

Water samples for analyses of NO3-, N2 and N2O and isotopologues of N2O were taken in

four sessions from April to December 2008. CO2, CH4 and N2O emission was measured

using the closed-chamber/gas-chromatographic technique (Mander et al., 2003 WST).

Isotopologue signatures of N2O such as δ18

O, average δ15

N (δ15

Nbulk

) and 15

N site preference

(SP = difference in δ15

N between the central and peripheral N positions of the asymmetric

N2O molecule) were measured using an enhanced IRMS in the Centre for Stable Isotope

Research and Analysis, University of Göttingen, Germany (Well et al 2005 Environ Sci). The

SP value has been used as an indicator of N2O from denitrification.


During the experimental period of a pulsing water table, water quality parameters did not

differ significantly between the inflow and outflow parts. Likewise, in the inflow part, there

were no significant differences between high and low water table periods. In the outflow part,

however, TOC and TN values were significantly higher at the deep water table. Average ±

standard values of TN and TP in the outflow (18±2.3 to 25±2.0, and 1.5±0.9 to 3.7±2.1 mg L-

1, respectively) were notably higher than during the normal management regime (Öövel et al

2007).

The emission of CO2 was significantly higher from the inflow part (between 18.2±3.6 and

20.3±3.0 mg CO2-C m-2

h-1

) than from the outflow part (from 6.4±0.8 to 12.2±1.4 mg CO2-C

m-2

h-1

), whereas at the higher water table in the outflow section, the CO2 flux was

significantly higher than at the deep water table level. While the N2O emission varied

between 1.6±1.5 and 4.9±2.0 μg N2O-N m-2

h-1

in the inflow and from 3.0±0.6 to 3.2±1.2 μg

N2O-N m-2

h-1

in the outflow, the inflow emission values at the deeper water table (in better

aerated conditions) were significantly higher than at the higher water table (in saturated

conditions). The dramatic difference in CH4 emission between the inflow (123±11.3 to

156.4±5.1 mg CH4-C m-2

h-1

) and outflow emission values (2.0±0.3 to 4.5±2.8 mg CO2-C m-2

h-1

) is probably due to the accidental killing of the vegetation in the outflow section due to the

use of herbicides in the adjacent field. We assume that the methanogenetic archea and

bacteria were also severely damaged. The GHG emission level in Paistu-Sultsi is comparable

to that from analogous studies on HSSF CWs (Mander et al 2003, 2011).

The value of δ15

Nbulk

N2O and δ18

O-N2O in water samples varied from -2 to 32 ‰ and

between 41 and 78 ‰, whereas the SP N2O value was from 15-41 ‰. There was a

significant positive correlation (p < 0.05) between the δ18

O-N2O and δ15

Nbulk

N2O values (R2

= 0.35) and between the δ18

O-N2O vs SP N2O values (R2 = 0.77). No significant relationship

was found between other isotopologue values. The SP N2O values, as well as the correlation

between the isotopologue parameters corroborate that the main source of N2O fluxes in the

studied HSSF CW bed is denitrification.

One can conclude that the short-term (one month) and short-range (up to 35 cm)

fluctuation of the water table in HSSFs can decrease CH4 emission and enhances CO2 and

N2O emission, whereas water purification efficiency may decrease.

ACKNOWLEDGEMENTS

This study was supported by the Estonian Research Council (grant IUT2-16) and the EU

through the European Regional Development Fund (Center of Excellence ENVIRON).


114

Molecular and microbial advances

related to pollutant fate, disposal

and removal in wetlands


115

Microbial nitrogen transformation in constructed wetlands

treating contaminated groundwater (O.40)

Oksana Voloshchenkoa, Kay Knoeller

a, Peter Kuschk

b

aDepartment of Catchment Hydrology, Theodor-Lieser-Strasse 4, 06120 Halle/Saale,

Germany ([email protected], [email protected])

bDepartment of Environmental Biotechnology, Permoserstrasse 15, 04318 Leipzig, Germany

([email protected])

INTRODUCTION

To improve the treatment of ground- and wastewater in constructed wetlands (CWs),

better understanding of the ongoing processes is vital. This research explores N-isotope

transformations in the removal of ammonium (N-transformation) from contaminated

groundwater in pilot-scale CWs downstream of the chemical industrial area Leuna/Germany.

This groundwater is contaminated mainly by organic (BTEX, MTBE) and inorganic (NH4+,

NO3-) chemicals.

We assume that in these horizontal subsurface flow CWs the anaerobic ammonium

oxidation (ANAMMOX) plays an important role in nitrogen removal. However, to date,

interactions between processes of aerobic and anaerobic ammonium oxidation in CWs still

have not been well explored. Especially, the importance of the ANAMMOX process for the

nitrogen removal is generally accepted, but its role in HSSF-CWs is quite unknown.

METHODS

Types of CWs:

(i) planted horizontal subsurface flow (HSSF-CW), (ii) unplanted HSSF CW, and (iii)

floating plant root mat (FPRM)

Sampling points:

Inflow, outflow and pore water samples at three distances from inlet (1, 2.5, and 4 m) and

at three depths (20, 30, and 40 cm) for HSSF CWs and 30 cm for FPRM

Physicochemical parameters:

pH, temperature, redox potential, inorganic ions (NH4+, NO2

-, NO3

-), N2O, organic

compounds (CH4, MTBE, BTEX), water balance and contaminated mass loads calculations

Stable isotope methods: 15

N/14

N isotope signatures of NH4+; 15

N/14

N and O18

/O16

isotope signatures of NO3-

Microbiological techniques:

DNA extraction, pyrosequencing, FISH, and confocal laser scanning microscopy


Samples of soil pore water, inflow and outflow were collected in a time interval from 1 to

6 weeks. Within the CWs spatial concentration gradients of the nitrogen species (ammonium

and nitrate) can be shown. After the isotope analysis (see Fig. 1), the δ15

N variations of

ammonium and nitrate will be interpreted according to the prevailing processes during the N-

transformation processes. By an isotope mass-balance approach, it is anticipated to quantify

the N-transformations as nitrification, denitrification, and ANAMMOX.


116

Fig. 1. δ

15N-NH4

+ and δ

15N-NO3

- versus to fraction of residual substrate with enrichment factors at

planted horizontal subsurface-flow constructed wetlands Leuna, Germany, 14.08.2012

DNA from biofilms at roots and gravel was extracted using FastDNA® Spin Kit For Soil

(MP Biomedicals). At next steps, pyrosequencing and specific FISH probes in connection

with confocal laser scanning microscopy will give information about structure and spatial

distribution of the microbial nitrogen transforming community.

CONCLUSIONS

According to the isotope results, ammonium oxidation seems to proceed in a

straightforward manner along the flow path. In contrast, no clear evidence is provided by the

isotope results for a significant impact of denitrification. This is probably due to the

superimposition of further N-transformation processes such as nitrification.

ACKNOWLEDGMENTS

This research was supported by: the framework of the Marie Curie Initial Training

Network ADVOCATE – Advancing sustainable in situ remediation for contaminated land

and groundwater, funded by the European Commission, Marie Curie Actions Project No.

265063, SAFIRA project and the Helmholtz Interdisciplinary Graduate School for

Environmental Research (HIGRADE).


117

Dynamics of bacterial communities in constructed wetlands from

modelling results (O.45)

Roger Samsó and Joan García

GEMMA - Group of Environmental Engineering and Microbiology, Department of

Hydraulic, Maritime and Environmental Engineering, Universitat Politècnica de Catalunya-

BarcelonaTech, C/ Jordi Girona, 1-3, Building D1, E-08034, Barcelona, Spain (E-mail:

[email protected], [email protected])

INTRODUCTION

Bacteria communities growing in constructed wetlands play a major role on the removal of

pollutants from wastewater and the presence of a stable community is a critical factor

affecting their performance. With this work we aimed at finding how long it takes for

bacterial communities to stabilise in constructed wetlands and at answering specific questions

regarding their abundance, spatial distribution and their relative importance on the treatment

processes. To our knowledge, this is the first time bacterial communities’ distribution and

dynamics are studied in SSF CWs from modelling results. This is also the first time that a

model is used to simulate the behaviour of a constructed wetland for 3 years of operation.

METHODS

The numerical model BIO_PORE (Samsó and García, 2013) was used to simulate the

dynamics of 6 functional bacteria groups (heterotrophic, autotrophic nitrifying, fermenting,

acetotrophic methanogenic, acetotrophic sulphate reducing and sulphide oxidising bacteria)

within a pilot wetland for a period of 3 years. Three indicators of bacterial stabilization were

used: 1) total biomass; b) effluent pollutant concentrations and c) Shannon’s diversity index.

Bacterial distribution was studied by plotting the concentration of the different bacterial

groups on a longitudinal cross-section of the simulated wetland at different times.


Results indicate that aerobic bacteria dominated the wetland until the 80th day of

operation. Anaerobic bacteria dominated the wetland from that moment and until the end of

the studied period (Fig 1).

Fig 1. Changes in the average concentration of a) heterotrophic (XH), nitrifiying (XA) and sulfide

oxidising (XSOB) bacteria and b) fermenting (XFB), sulphate reducing (XASRB) and methanogenic

(XAMB) bacteria (kgCOD m-3

) within the entire wetland through 3 years of operation. Total biomass

(XH+XA+XFB+XAMB+XASRB+ XSOB) is also represented in image b).

a) b)



118

Bacteria stability was reached between 400 and 700 days after starting operation (Figures

1b (Total biomass) and 2).

Fig. 2. Changes in Shannon’s diversity index (H’).

Once the wetland reached stability, sulphate reducing bacteria accounted for the highest

biomass of all bacterial groups (46%) (Fig. 1). The distribution of bacterial communities

obtained after bacterial stability is consistent with available experimental results, and was

clearly controlled by dissolved oxygen (SO) concentrations and H2S toxicity (results not

shown). After stability, the progressive accumulation of inert solids pushed the location of the

active bacteria zone towards the outlet section (results not shown).

CONCLUSIONS

In the present study we demonstrate based on simulations in a pilot system that bacteria

communities reach stability, but that it is a slower process than what has generally been

reported. The distribution of bacterial communities obtained after bacterial stability is

consistent with available experimental results. The results of this study coupled with previous

field results will give new insights and perspectives on wetland functioning.

ACKNOWLEDGEMENTS

This work was possible thanks to the funding from the Spanish Ministry of Innovation and

Science for the NEWWET2008 Project (CTM2008-06676-C05-01) and from the

NAWATEC FP7 Project. Roger Samsó also acknowledges the scholarship provided by the

Universitat Politècnica de Catalunya (UPC).

REFERENCES Samsó, R., García, J. BIO_PORE, a mathematical model to simulate biofilm growth and water quality

improvement in porous media: application and calibration for constructed wetlands. Ecol Eng 2013; 54:116-127.


119

Effect of plant diversity on microbial community function in

horizontal subsurface flow constructed wetlands (0.80)

Mark Buttona, Mariana Rodriguez

b, Jacques Brisson

b, Kela Weber

a

aDepartment of Chemistry and Chemical Engineering, Royal Military College of Canada,

Kingston (Ontario), K7K 7B4, Canada. ([email protected]) bInstitut de Recherche en Biologie Végétale, Département de sciences biologiques. Université

de Montréal. 4101 East, Sherbrooke St, Montreal (Quebec), H1X 2B2, Canada.

INTRODUCTION

Within treatment wetlands plant root systems provide mechanical support for microbial

community attachment, transfer oxygen from aerial tissues into the rhizosphere, and secrete

root exudates into the subsurface. It is through these processes that plants are thought to help

regulate the microbial community structure and function within the surrounding rhizosphere

(Weber and Legge, 2012). Ecological theory states that greater biodiversity (plant, animal,

microbial) provides for a more resilient and healthy ecosystem. In the context of treatment

wetlands it is hypothesized that through capitalizing on the complimentary nature of different

plant species, greater plant diversity can lead to greater microbial functional diversity, and

possibly an enhancement in the microbial community function in terms of water treatment

abilities. Phragmites australis and Phalaris arundinacea are commonly used species for

wastewater treatment with complementary traits (root system arrangement and seasonality)

(Vymazal et al. 2005). In this study horizontal subsurface flow (HSSF) mesocosm wetlands

planted with either Phragmites australis (A) or Phalaris arundinacea (B) were operated in

series under four possible combinations (ie. AA, AB, BA, BB), and the function of the

intrinsic microbial communities assessed via community level physiological profiling

(CLPP).

MATERIALS AND METHODS

Experimental set-up: Sixteen mesocosm scale experimental constructed wetland units

were set-up in a controlled greenhouse environment and run for the period April 2012-March

2013. Sampling of the microbial communities for CLPP analysis occurred in October 2012.

Each experimental unit consisted of two coupled mesocosms (L 70cm W 51cm H 36cm),

respectively planted according to each of the following four treatments: monocultures of P.

australis (AA) and P. arundinacea (BB) and the combination of the two plant species, P.

australis followed by P. arundinacea (AB), as well as P. arundinacea followed by P. australis

(BA). Experimental units were divided into two separate mesocosms to avoid any one

species dominating a polyculture. Each treatment was replicated four times following a

randomized block design. The mesocosms were filled with granitic river gravel (10-15 mm

diameter). Following a period of plant establishment, the experimental units were fed from

April 2012 with 15 L d1

of reconstituted wastewater from diluted fish farm sludge, at a

hydraulic loading rate of 42 L·m−2

·d−1

. Average influent concentration (in g·m−2

·d−1

) was

TSS 8.2; COD 14.3; PO4-P 0.6; NH4-N 1.1. The temperature of the greenhouse ranged from

35ºC in summer to 5ºC in winter. Community level physiological profiling (CLPP):

Microbial communities were sampled directly from each mesocosm via a central (vertically

mixed) sampling port. A 50 ml water sample was collected in a sterilized polyethylene tube

and transported to the laboratory on ice. Several microbial community functional (catabolic)

characterization metrics were extracted via CLPP according to Weber and Legge (2010)

including the overall catabolic activity (average measured activity for 31 different carbon

sources), catabolic richness (number of carbon sources out of 31 utilised), catabolic diversity


120

(Shannon index), and the carbon source utilisation pattern (CSUP) which was later used to

compare overall catabolic profile differences of the studied microbial communities.


Several differences were observed in the microbial community function between both

plant species P. australis (A) and P. arundinacea (B) and mesocosm position 1 or 2 (Table

1). When examining the results for communities from mesocosms in position 1 (with the

exception of treatment AA) the results are similar despite which plant is used (AWCD

between 1.12 and 1.19, richness between 26.17 and 27.67, diversity between 3.23 and 3.26).

The biggest difference in microbial community function between tank positions occurred

within a monoculture of P. australis (AA) with a decrease in AWCD and Richness of -40%

and -28% respectively between positions 1 and 2. This was followed by the polyculture of P.

arundinacea/P. australis (BA) with -30/-16 % reduction in AWCD/Richness, then P.

australis/P. arundinacea (AB) at -19/ -14%. Interestingly, an increase in AWCD (+8%) was

observed from mesocosm position 1 to 2 in the monoculture of P. arundinacea (BB) and only

a slight decrease in Richness (-3%). Multivariate analysis of the carbon source utilization

patterns (CSUPs) in the form Principal Component Analysis (PCA) reinforced the above

outlined differences between plant species and mesocosm position. Additional microbial

community characterization was performed in March 2013, with sampling planned for the

summer of 2013. These results will also be presented. Table 1: Summary of compiled catabolic activity (AWCD), richness and diversity results (values

represent means from 4 replicate systems)

AA

AWC

D

Richnes

s

Diversit

y AB

AWC

D

Richnes

s

Diversit

y

Position

1 P. australis 0.90 24.25 3.14 P. australis 1.19 27.67 3.26

Position

2 P. australis 0.53 17.50 3.00

P.

arundinacea 0.96 23.83 3.11

% Diff -40.51 -27.84 -4.55 % Diff -19.35 -13.86 -4.51

BB

AWC

D

Richnes

s

Diversit

y BA

AWC

D

Richnes

s

Diversit

y

Position

1

P.


P.


Position

2

P.

arundinacea 1.22 26.08 3.19 P. australis 0.77 22.08 3.10

% Diff 7.93 -2.80 -1.69 % Diff -30.66 -15.61 -4.05

CONCLUSIONS

Catabolic activity (AWCD), catabolic richness and CSUPs for the microbial communities

profiled indicate that microbial community function can be altered by choice of plant species

and spatial positioning along a nutrient gradient; however, at this stage specific advantages of

using plant polycultures over monocultures to enhance microbial function are not proven or

clear.

REFERENCES Vymazal, J. and Kropfelova, L. (2005) Growth of Phragmites australis and Phalaris arundinacea in constructed

wetlands for wastewater treatment in the Czech Republic. Ecol. Eng. 25:606–621.

Weber, K.P. and Legge, R.L. (2010) Community-level physiological profiling. Methods in molecular biology

(Clifton, N.J.). 599: 263-281.

Weber, K.P. and Legge, R.L. (2013) Comparison of the catabolic activity and catabolic profiles of rhizospheric,

gravel-associated and interstitial microbial communities in treatment wetlands. Water Sci. Technol. 67:886-

893.


121

Assessment of bacterial community structure in horizontal

subsurface flow constructed wetland units treating municipal

wastewater using next-generation sequencing (O.34)

K. Oopkaupa, J. Truu

a, M. Truu

a, H. Nõlvak

a, T. Sildvee

a, J.-K. Preem

a, Ü.

Mandera

a Department of Geography, Institute of Ecology and Earth Sciences, University of Tartu, 46

Vanemuise St., Tartu, 51014, ESTONIA, ([email protected])

INTRODUCTION

Constructed wetlands (CW) have become widely used for wastewater treatment over the

recent decade for their cost-effective building and maintenance. While physical and chemical

pollutant removal mechanisms in CWs are well known, understanding of microbial processes

related to pollutant removal in CWs is still low. Furthermore, the dynamics of microbial

community in CWs during the start-up period is still not clearly understood.

Recent advances in qualitative and quantitative microbial techniques have made possible

to detect more precisely the microbial community structure and abundance in CWs.

Consequently, assessing the data acquired by high-throughput sequencing and quantitative

PCR, the objectives of this study were: (1) to determine the diversity and dynamics of

bacterial communities in newly established horizontal subsurface flow filter units of

constructed wetlands treating municipal wastewater; (2) evaluate the relationships between

microbial community structure and purification efficiency.

METHODS

An experimental hybrid constructed wetland system was located on the territory of the

activated sludge wastewater treatment plant of Nõo borough in Estonia. The microbial

community on the filter material (lightweight expanded clay aggregates (LECA) with particle

size of 2–4 mm) of three horizontal subsurface flow mesocosms (HSSF MCs) and its influent

was analysed in current study. Total of five sampling occasions on day 25, 45, 67, 94, and

150 of the system performance, respectively were performed during the five month

experiment. Microbial community profiling was performed using Illumina® HiSeq 2000

sequencing combinatorial sequence-tagged PCR products of the V6 hypervariable region of

the 16S rRNA gene. Quantitative PCR (qPCR) was applied for the evaluation of changes in

bacterial community size by its 16S rRNA gene copy numbers and denitrification potential

by its nitrite reductase encoding nirS and nirK gene copy numbers. To assess the relationship

between microbial community structure and HSSF MCs treatment efficiency Molecular

Ecological Network Analyses was applied.


Bacterial community structure in HSSF filter

The structure of the bacterial community developing on filter material particles during the

treatment process differed from the initial community of the material and from the

community entering into the filter with wastewater (Table 1). Multivariate analysis indicated

that in general the differences between sampling occasions were substantially greater than

community differences among replicas.


122

Table 1. Relative abundance (%) of the five most dominant bacterial phyla in influent, filter material

(FM) and mesocosms (n=3) average value and standard deviation (in parentheses). d - days number of the

system performance.

Phylum Influent FM 25d 45d 67d 94d 150d

Proteobacteria 41.7 36.5 54.9 (±1.9) 51.3 (±1.2) 54.5 (±2.7) 46.1 (±7.4) 49.9 (±0.7)

Bacteroidetes 20.0 11.3 20.3 (±0.6) 17.0 (±1.1) 15.5 (±1.2) 21.0 (±6.0) 18.2 (±0.5)

Actinobacteria 2.1 24.6 5.4 (±1.1) 5.6 (±1.2) 5.0 (±0.6) 5.4 (±1.1) 4.7 (±0.3)

Firmicutes 25.3 0.8 2.8 (±0.2) 2.8 (±0.4) 2.9 (±0.1) 2.2 (±0.1) 1.9 (±0.1)

Verrucomicrobia 0.4 2.2 3.3 (±0.4) 3.5 (±0.4) 3.1 (±0.4) 3.3 (±0.6) 3.4 (±0.2)

Factors affecting bacterial community composition in HSSF MCs

According to the distance-based regression analysis, the system operational time had a

statistically significant (p<0.001) impact on bacterial community structure in HSSF filter

MCs. The influent quality characteristics such as nitrate (48% variation explained), total

nitrogen (14% variation explained), total organic carbon (9% variation explained), and pH

(6% variation explained) were related to formation of bacterial community structure and had

a cumulative effect explaining 77% of variation.

Bacterial community abundance and denitrification potential of the HSSF filter

The number of bacterial 16S rRNA genes in the filter material increased remarkably

during the first 94 days of the system operation on filter material and then stayed stable until

the end of the experiment.

The time of system performance had a positive impact on nirK gene copy number, their

proportion in bacterial community and in addition affected nirK/nirS ratio. The filter

temperature had negative impact on evaluated nirK parameters. NirS affected amounts of all

measured nitrogen fractions and BHT7 value in effluent. Both nir genes were related to the

effluent pH.

Relationship between HSSF filter treatment efficiency and microbial community

structure

Statistically significant (p<0.05) correlations between module-based eigengenes and HSSF

treatment efficiency were found: first and fourth submodule was positively related to NH4-N

(r=0.58), second submodule negatively to NH4-N (r=-0.64) and TOC (r=-0.64), and third

submodule to NO2-N (r=0.77) and NO3-N (r=0.59) removal.

CONCLUSIONS

Bacterial community had clear successional pattern in a newly established CW;

Microbial community structure differed between influent, initial filter material and HSSF

filter MCs;

Total bacterial community abundance reached its maximum after two months of system

operation and remained stable;

Environmental factors had a different effect on the abundance of the denitrifiers nirS and

nirK gene;

Different bacterial phylotypes had a co-effect on different chemical compounds removal;

ACKNOWLEDGEMENTS

This study was supported by Target Project No SF0180127s08 of the Ministry of

Education and Research of the Republic of Estonia, Estonian Science Foundation grant 9387

and by the European Regional Development Fund through ENVIRON (Centre of Excellence

in Environmental Adaptation).


123

Microfauna from different configurations of constructed wetlands (O.61)

Anna Pedescollab

, Lorena Rodrígueza, Aida A. Sarañana

a, Eloy Bécares

a

a Ecology Section, Department of Biodiversity and Environmental Management, University

of León, Campus de Vegazana s/n, 24071 León, Spain ([email protected]).

b Environmental Institute, c/ La Serna 56, 24007 León, Spain.

INTRODUCTION

Little data is available concerning microfauna in constructed wetlands (CWs) (Puigagut et

al., 2007). This study aimed at evaluating the microfauna community in horizontal CWs

according to its design configuration. In this first approach, abundances of different groups of

protozoa and metazoan were analysed.

METHODS

Eight mesocosms-scale CWs were built inside the facilities of the León WWTP, in the

Northwest of Spain. Each CW consisted of a fibreglass container measuring 80 cm wide, 130

cm long and 55 cm high, which differed from each other in the design configuration.

Characteristics of the CWs are listed in Table 1. The experimental plant was operated from

May 2007 to December 2010. The wetlands were fed with homogenised wastewater from the

primary settler of the León WWTP at a hydraulic loading rate of 50 mmd-1

(CW6’ received

100 mmd-1

) with a continuous flow rate. Table 1. Main design characteristics of the wetlands of the experimental plant.

CW Plant species Flow type

Gravel

matrix

(cm)

Water

depth

(cm)

Outlet

pipe

position

Organic load

(g BOD5m-2

d-1

)

CW1 Typha

angustifolia

Hydroponic (Floating

macrophytes)

Without

gravel 30 Top 3-10

CW2 Typha

angustifolia

Free water surface

(FWS) 25 50 Top 3-10

CW3 Typha

angustifolia FWS 25 50 Bottom 3-10

CW4 Unplanted FWS 25 50 Bottom 3-10

CW5 Phragmites

australis

Hydroponic (Floating

macrophytes)

Without

gravel 30 Bottom 3-10

CW6 Phragmites

australis Subsurface flow (SSF) 50 45 Bottom 3-10

CW6’ Phragmites

australis SSF 50 45 Bottom 6-20

CW7 Unplanted SSF 50 45 Bottom 3-10

Samples of outlet were taken in summer and winter campaigns and analysed for protozoa

(ciliates and amoeba) and metazoan (tardigrades, copepods, rotifers, nematodes, turbellarians,

cladocerans and acari) abundances.


Differences were observed, in terms of protozoa (which included ciliates and amoeba) and

metazoan, both depending on the system considered and the season sampling campaign.

Protozoa were mainly ciliates (amoeba represented 2% of the total protozoa in average)

while the most abundant groups of metazoan were rotifers and turbellarians. Ciliates


124

abundance in outlet samples was in the range of other studies in which ciliates were analysed

near the outlet (Puigagut et al., 2012). Protozoa were more abundant in winter in those

systems with free water table, this is CW1 and CW5 (hydroponic systems) and CW2, CW3

and CW4 (FWS systems). In contrast, for those CWs abundance of metazoan in winter was

lower than for SSF systems (Fig. 1). Thus, a negative relationship can be observed between

protozoa and metazoan.

Fig. 1. Abundance of microfauna in outlet samples of the experimental plant for winter and summer

campaigns.

Differences were also observed between metazoan groups composition depending on the

system considered. Therefore, hydroponic and strict SSF systems were dominated by

turbellarians and in unplanted systems rotifers were the most abundant metazoan.

CONCLUSIONS

Microfauna community seemed to be affected by the configuration of the wetland. The

presence of plants and flow type influenced the abundance of the groups within the

microfauna community.

ACKNOWLEDGEMENTS

This study was funded by the Spanish Ministry of Science through the projects CTM2005-

06457-C05-03 and CTM2008-06676-C05-03/TECNO. Anna Pedescoll acknowledges the

Juan de la Cierva Programme of the Spanish Ministry of Science and Innovation.

REFERENCES Puigagut, J., Salvadó, H., García, J. (2007) Effect of soluble and particulate organic compounds on microfauna

communities in subsurface flow constructed wetlands. Ecological Engineering 29, 280-286.

Puigagut, J., Maltais-Landry, G., Gagnon, V., Brisson, J. (2012) Are ciliated protozoa communities affected by

macrophyte species, date of sampling and location in horizontal sub-surface flow constructed wetlands? Water

Research 46, 3005-3013.

CW1 CW2 CW3 CW4 CW5 CW6CW6' CW7

Pro

tozoa (

indiv

iduals

L-1

)

0

100000

200000

300000

1000000

CW1 CW2 CW3 CW4 CW5 CW6CW6' CW7M

eta

zoa (

indiv

iduals

L-1

)0

1000

2000

3000

4000

10000

12000Winter

Summer


125

Factors affecting microbial community structure in a created

riverine wetland complex (O.47)

Teele Sildveea, Kristjan Oopkaup

a, Marika Truu

a, Jaak Truu

a, Hiie Nõlvak

a,

William J. Mitschb, Ülo Mander

a

aDepartment of Geography, Institute of Ecology and Earth Sciences, 46 Vanemuise Street,

Tartu, 51014, Estonia ([email protected])

bWilma H. Schiermeier Olentangy River Wetland Research Park, 352 W. Dodridge Street,

Columbus, OH 43202, USA

INTRODUCTION

Constructed wetlands are proposed as an effective and low-cost solution to decrease

nutrient concentrations in polluted surfaces and subsurface waters. Whereas purification

processes in constructed wetlands depend greatly on microbial organisms, it is important to

understand the factors affecting the composition of bacterial communities.

The purpose of this study was to determine the bacterial community structure and its

relationship with denitrification potential and site-specific characteristics such as chemical

parameters, water regime in the wetland, wetland type and soil type of the created riverine

wetland complex in Olentangy River Wetland Research Park (ORWRP) in Ohio, USA.

METHODS

The ORWRP study site includes two 1 ha experimental marshes (W wetlands) and a 3 ha

river diversion wetland (oxbow). W wetlands are fed with the Olentangy River water by

continuous pumping while into the oxbow water enters through the check valve only in the

cases when the water level in the river is higher than in the wetland.

29 soil and sediment samples were collected in March 2009. Samples from W wetland

open (Woo, n=6) and transitional water regime areas (Wtrans, n=6), the oxbow (Ox, n=7), and

upland (Up, n=4) were obtained from the 0-15 cm top soil layer. Mineral samples from W

wetland open areas (Wom, n=6) were collected below the organic layer from a depth of 15-30

cm.

The pHKCl values and total C and N, NH4-N, NO3-N, P, Ca, K, and Mg content of the soil

samples were determined. Bacterial community was profiled using Illumina HiSeq2000

sequencing of 16S rRNA gene V6 region. 16S rRNA, nirS, nirK, and nosZ gene copies in the

soils were quantified by using the qPCR method and proportions of denitrifying genes in the

total community were calculated.


The microbial communities of all the studied soil groups were dominated mostly by the

same phyla; however, their proportions were variable between the groups. In the microbial

communities of the Wtrans, Ox and Up the dominant phylum was Proteobacteria (35.6±4.9%,

37.7±9.5%, and 25.5±2.3%, respectively), followed by Acidobacteria (15.8±4.3%,

14.1±4.7%, and 24.0±3.2%), Actinobacteria (6.6±1.1%, 11.4±5.7%, and 21.2±4.6%), and

Bacteroidetes (6.2±0.9%, 7.9±1.8%, and 9.5±2.2%). Proteobacteria was the dominating

phylum also in the Woo and Wom microbial communities (54.0±3.2% and 49.8±7.9%,

respectively), followed by Bacteroidetes (9.3±1.5% and 10.2±2.3%) and Acidobacteria

(7.1±1.1% and 6.8±1.8%).

Nonmetric multidimensional scaling based on the 16S rRNA sequencing data displayed

clear clustering of the bacterial communities of studied soil groups (Fig. 1). The Up samples


126

situated more separately from the clusters of the Woo and Wom samples and the Wtrans and Ox

samples.

Fig. 1. The ordination plot of the soil samples according to the nonmetric multidimensional scaling of the

16S rDNA sequencing data.

The denitrification potential was different in the studied soil groups. While nirK and

nirK/16S rRNA values were higher in the Ox compared to the Wom samples (p<0.01), nirS

and nirS/16S rRNA values showed greater values in Woo compared to the Up samples

(p<0.01). NosZ gene abundance was lower in the Wom in comparison with the Wtrans and Ox

samples, but the proportion of this gene was not different in bacterial communities of the

studied soil groups.

All the quantified denitrification gene values were related to microbial community

structure. The microbial community diversity was strongly correlated with soil NO3-N and

NH4-N concentrations, whereas relationships were found also with pH, Ca, C/N, K and P

values.

CONCLUSIONS

Soil bacterial community structure was dependent on the water regime. The studied

samples shared the same dominant phyla; however, their proportions were different. The

microbial communities in the Up samples were clearly distinguishable from the W wetland

and oxbow samples. Denitrification genes were affected differently by environmental factors

in studied soil groups. Microbial community structure was related to denitrification genetic

potential, wetland soil chemical properties and its nitrogen removal processes.

ACKNOWLEDGEMENTS

This study was supported by Target Project No SF0180127s08 of the Ministry of

Education and Research of the Republic of Estonia, Estonian Science Foundation grant 9387,

and by the European Regional Development Fund through ENVIRON (Centre of Excellence

in Environmental Adaptation).


127

Characterization of metagenomes of microbial communities in

vertical and horizontal flow filter units of full scale constructed

wetland (O.64)

Jaak Truu, Marika Truu, Jens-Konrad Preem, Teele Sildvee, Kristjan Oopkaup,

Kertu Tiirik, Hiie Nõlvak, Ülo Mander

Department of Geography, Institute of Ecology and Earth Sciences, 46 Vanemuise Street,

Tartu, 51014, Estonia ([email protected])

INTRODUCTION

High-throughput sequencing methods, such as 454 pyrosequencing, Illumina sequencing

and Ion Torrent semiconductor technology have been applied to characterize phylogenetic

and functional genes in microbial community of different conventional water treatment

systems (Albertsen et al. 2011, Whiteley et al. 2012, Yu and Zhang, 2012). Metagenomic

approaches offer the ability to examine directly the genomic content of microbial

communities, complementing taxonomic information with functional capability.

The purpose of this study was to characterize metagenomes of microbial communities in

vertical and horizontal flow filter units of the full scale constructed wetland.

METHODS

Two pooled samples were obtained from vertical and horizontal flow filters of the hybrid

constructed wetland (CW) system treating wastewater from a basic school in Paistu, Estonia.

After DNA extraction form samples total community DNA was sequenced using Illumina

MiSeq platform in 2x250-base paired-end reads mode. Obtained data was processed using

different programs and software platforms.


Domain distributions in two samples showed the dominance of Bacteria (>80%), a small

fraction of Eukaryotes (1-2%) and Archaea (>1%) in biofilm samples of both filter units. The

most abundant bacterial groups in vertical flow unit were Proteobacteria (51%), followed by

Bacteroidetes (10%) and Actinobacteria (5%). Archaea had two dominant phyla -

Thaumarchaeota (46%) and Euryarchaeota (44%). All known species from phyla

Thaumarchaeota are chemolithoautotrophic ammonia-oxidizers. Around 25% of Archaeal

sequences in VF unit belonged to methanogens.

In case of horizontal flow unit the most abundant bacterial groups were Proteobacteria

(44%), followed by Bacteroidetes (10%) and Firmicutes (8%). Archaea had only one

dominant phylum Euryarchaeota (97%). Over 60% of Archaeal sequences in HF unit

belonged to methogenic species. There were differences in distribution of classes within

Proteobacteria phylum in two systems. While the proportion of Alphaproteobacteria was

same (9%) then Betaproteobacteria and Gammaproteobacteria were more abundant in

vertical flow unit.

Metabolism of carbohydrates, proteins, amino acids and derivatives were the three most

abundant gene categories in global metabolism of microbial community in vertical flow filter

unit. In case of nitrogen metabolism nitrate and nitrite ammonification related genes were

dominant (33%) followed by ammonia assimilation (25%) and nitrogen fixation (16%).

Abundance of denitrification related genes (nitrous oxide reductases and nitrite reductases)

was 7% and 6% respectively in nitrogen metabolism related gene sequences. Phosphorous

metabolism was dominated by phosphate metabolism genes and sulphur metabolism by


128

inorganic and inorganic sulphur assimilation. Further data analysis to characterise and

compare metagenomes of two filter units is in progress.

CONCLUSIONS

Sequencing of full metagenomes of constructed wetland microbial communities allows to

explore microbial metabolic potential and characterize key processes involved in

transformation of carbon, nitrogen and sulphur within constructed wetland different

compartments. In addition metagenomic analysis provides taxonomic information about

microbial community structure in these treatment systems. Metagenome sequence data could

be applied for discerning presence of sequences originating from putative pathogenic bacteria

as well as for assessment of antibiotic resistance determinants.

ACKNOWLEDGEMENTS

This study was supported by the Ministry of Education and Research of the Republic of

Estonia (grant IUT2-16), state program „Aid for research and development in environmental

technology“ grant 3.2.0801.11-0026, and by the European Regional Development Fund

through ENVIRON (Centre of Excellence in Environmental Adaptation).

REFERENCES Albertsen, M., Hansen, L. B. S., Saunders, A. M., Nielsen, P. H., & Nielsen, K. L. (2011) A metagenome of a

full-scale microbial community carrying out enhanced biological phosphorus removal. ISME J. 6: 1–13.

Whiteley, A. S., Jenkins, S., Waite, I., Kresoje, N., Payne, H., Mullan, B., O’Donnell, A. (2012) Microbial 16S

rRNA Ion Tag and community metagenome sequencing using the Ion Torrent (PGM) Platform. J. Microbiol.

Meth. 91: 80–8.

Yu, K., and Zhang, T. (2012) Metagenomic and metatranscriptomic analysis of microbial community structure

and gene expression of activated sludge. PloS One, 7: e38183.


129

Energy production with microbial fuel cells implemented in

horizontal subsurface flow treatment wetlands (O.108)

Clara Corbellaa, Jaume Puigagut

a

a GEMMA – Group of Environmental Engineering and Microbiology, Department of

Hydraulic, Maritime and Environmental Engineering, UniversitatPolitècnica de Catalunya-

BarcelonaTech, C/JordiGirona, 1-3, Building D1, E-08034, Barcelona, Spain (Email:

[email protected], [email protected])

INTRODUCTION

Microbial Fuell Cells (MFC) are bioelectrochemical systems that use microorganisms to

oxidize organic and inorganic matter and generate current (Logan et al. 2006). In a MFC

bacteria reduce an electrode (anode) under lower redox conditions and the electrons flow

through an electrical circuit toward a high redox scenario where it reduces an electron

acceptor such as oxygen, at the cathode. Natural redox gradients along the depth in horizontal

subsurface flow treatment wetlands could be exploited to produce energy via MFC

implementation. The purpose of the present work was to quantify the energy production that

can be obtained by implementing MFC technology in constructed wetlands treating urban

wastewater.

METHODS

Energy production was measured in a pilot plant that consisted of 2 horizontal subsurface

flow treatment wetlands (SSF TW) of 0,4 m2

each and a wetted depth of 25 cm planted with

common reed at an initial density of 16 plants m-2

. Wetlands were fed with primary settled

urban wastewater at a continuous flow of 21 L.day-1

.

MFC electrodes consisted of 35 graphite rods (1 cm long x 0,5 cm diameter) wrapped with

stainless steel iron mesh. Projected surface of electrodes was that of 0.0012 m2 and they were

placed at 5 cm (cathode) and 15 cm (anode) of wetland depth in a cylindrical plastic mesh

filled up with gravel media between electrodes. The circuit was closed using isolated copper

wires with an external resistance of 1000 Ω. Voltage generated was measured for three

months by means of a datalogger (DATA TAKER DT50 series 3). Redox potentials at the

cathode and anode locations were also measured using two redox probes (Digimed TH-404)

equipped with a platinum electrode (Ag/AgCl reference 114 system - accuracy: ±10 mV).


Redox potentials measured at 5 cm and 15 cm depth were significantly different.

Accordingly, at 5 cm depth redox conditions were higher (average daily redox of 196±29 mV

respect to the standard hydrogen electrode) than those recorded at 15 cm depth (average daily

redox of -124±118 mV). Therefore, the average daily redox gradient generated was, in

average, that of 318±112 mV, with a maximum gradients being that of 512 mV at the end of

the study period (late summer). Results showed a very similar pattern regardless the

experimental campaigns considered. Figure 1a shows a typical redox profile along 24 hours

of monitoring for the late summer campaign.


130

Time (hours)

10h

11h

12h

13h

14h

15h

16h

17h

18h

19h

20h

21h

22h

23h

24h

01h

02h

03h

04h

05h

06h

07h

08h

09h

Eh

(m

V)

-400

-300

-200

-100

0

100

200

300

5 cm

15 cm

Time (weeks)

0 2 4 6 8 10 12

Volt

age

(mV

)

0

20

40

60

80

100

120

Replicate 1

Replicate 2

Fig. 1. a) Redox potential measured at 5 cm and 15 cm depth during late summer campaign b) Electric

output obtained during summer campaign.

Voltage produced with MFC implemented in SSF TW varied from ca. 100 mV to ca. 20

mV, to the beginning and end of the study period, respectively (Figure 1b). Although the

theoretical voltage attainable (Eemf) corresponds to the redox gradient between anode and

cathode (Logan et al., 2006), the actual energy production (Ecell) is usually lower due to both

overpotentials and/or ohmic losses (Logan et al. 2006). To this regard, results showed a

maximum Ecell of ca. 100 mV that corresponded roughly to a 25% of the maximum attainable

voltage (Eemf) and about 4 mW.m-2

of anode. Maximum power production here recorded is in

the range of that described in literature (Rabaey et al., 2005). However, maximum Ecell was

only sustainable for 3 weeks (Figure 1b). Accordingly, we detected a drop in Ecell from the

third week on (Figure 1b) that lead to a Ecell of ca. 20 mV. The decrease was probably due to

the observed corrosion of the metal mesh wrapping the graphite rods that difficulted the

electrons flow from anode to cathode. This phenomena, known as electrode passivation, has

been previously described in the context of MFC in marine environments (Reimers et al.,

2006).

CONCLUSIONS

Constructed wetlands are suited to successfully implement microbial fuel cells for energy

production due to marked redox gradients along the depth. In our study, maximum cell

voltages of ca. 100 mV were recorded that corresponded to ca. 25% of the cell electromotive

force (maximum attainable cell voltage). However, electrodes materials that avoid corrosion

are a key factor to provide high sustainable cell voltages. .

ACKNOWLEDGEMENTS

This study was funded by the Spanish Ministry of Science and Innovation (MICINN)

(projectCTM2010-17750)

REFERENCES Logan, B.E., Hamelers, B., Rozendal, R., Schr¨der, U., Keller, J., Freguia, S., Aelterman, P., Verstraete, W.,

Rabaey, K. (2006) Microbial Fuel Cells: Methodology and Technology. Environmental Science & Technology

Vol.40 17:5181-5192

Rabaey, K., Verstraete, W. (2005) Microbial fuel cells: novel biotechnology for energy generation. TRENDS in

Biotechnology. Vol.23 No.6

Reimers, C.E., Girguis, P., Stecher III, H.A., Tender, L.M., Ryckeliynck, N., Whaling, P. (2006) Microbial fuell

cell energy from an ocean cold seep. Geobiology. 4:123-136

(a) (b)


131

Comparison of the microbial community metabolic function in

different field-scale constructed wetland designs: A spatial

dynamics study (O.116)

Mark Buttona, Jaime Nivala

b, Kela Weber

a, Thomas Aubron

b, Roland Mueller

b

aDepartment of Chemistry and Chemical Engineering, Royal Military College of Canada,

Kingston (Ontario), K7K 7B4, Canada. ([email protected])

bHelmholtz Center for Environmental Research (UFZ), Environmental and Biotechnology

Center (UBZ), Permoserstrasse 15, 04318 Leipzig, Germany ([email protected])

INTRODUCTION

Current constructed wetland (CW) technology spans a diverse range of designs from

completely passive horizontal surface or subsurface flow to intermediate systems such as

unsaturated vertical flow with pulse loading to increasingly sophisticated and technologically

dependent systems utilizing increased pumping, water level fluctuation or active aeration [1].

The overall efficacy of CWs for wastewater treatment is determined by a combination of

wetland design, vegetation type and microbial processes. Microorganisms are central to the

biogeochemical processes in wetlands, and as such to improve the performance of

constructed wetlands it is essential to have some understanding of the functional diversity and

metabolic properties of the microbial community [2]. In this study we spatially explore the

microbial community functional dynamics in several different CW design types by assessing

the rate and ability of microbial communities to utilize a range of different carbon sources.

MATERIALS AND METHODS

Experimental site: The pilot-scale constructed wetlands at the research facility in

Langenreichenbach, Germany differ in terms of flow direction, degree of media saturation,

media type, loading regime and aeration mechanism. A detailed description can be found in

Nivala et al. (2013)[1]. Table 1 briefly describes the systems surveyed in the present study. Table 1: Details of the CW systems and sampling locations.

System

Abbreviationa

System Type Depth of

Main Media

(m)

Saturation

Status

Approximate

Daily Flow

(L/d)

Internal

Sampling

Points

Horizontal Flow

H25, H25p HF 0.25 Saturated 90 4 locationsb

H50, H50p HF 0.50 Saturated 180 4 locationsb

Vertical Flow

VS, VSpe VF 0.85 Unsaturated 565 3 locationsc

Intensified

VAc VF + Continuous Aeration (with plants) 0.85 Saturated 565 3 locationsd

HAc HF + Continuous Aeration (with plants) 1.00 Saturated 750 4 locationsb

HAw HF + Windmill Aeration (diaphragm pump) 1.00 Saturated 750 4 locationsb

aSystems planted with Phragmites australis are denoted with “p” in the system abbreviation, other systems are unplanted. bSampling tees located mid-depth at 12%, 25%, 50%, and 75% fractional length (4.7 m); named 1, 2, 3, and 4, respectively. cSampling pans located at approximately 12.5%, 25%, and 50% of the effective depth; named 1, 2, and 3, respectively. dSampling tees located at the mid-point of the upper, middle, and lower thirds of the bed, named 1, 2, and 3, respectively. eThe VS1 and VS1p beds were operated as a second stage of a two-stage system; all other beds received primary-treated

domestic wastewater.

Community Level Physiological Profiling (CLPP): Microbial communities were

sampled directly from each CW system via a (horizontally mixed) sampling port. The water

sample was collected in a sterilized polyethylene tube or glass beaker and transported to the

laboratory on ice. Several microbial community functional (metabolic) characterization


132

metrics were extracted via community level physiological profiling according to Weber and

Legge (2010)[3] including the overall metabolic activity (average measured activity for 31

different carbon sources - AWCD), metabolic richness (number of carbon sources utilized out

of 31 total), and the carbon source utilisation pattern (CSUP).


Overall microbial community metabolic activity (AWCD) decreased dramatically along

the flow path in the horizontal sub surface flow wetlands (Figure 1A). The biggest drops in

activity occurred from 25 to 50% along the flow path for most HF systems with the exception

of H50p and HAw. Figure 1B presents a principle component analysis (PCA) ordination

based on the carbon source utilisation patterns from the CLPP analysis. Five groupings are

apparent differentiating the first two sampling points of the HF systems (sample points 1 and

2), the last two sampling points of the HF systems (sample points 3 and 4), the vertical flow

systems (VAc, VS and VSp), the continuously aerated system (HAc), and system with

intermittent windmill aeration (HAw).

Figure 1: (A) Microbial activity as function of distance along the flow path, and (B) PCA ordination based

on the CLPP carbon source utilisations patterns (CSUPs) for the different CW systems.

CONCLUSIONS

The CLPP method was useful in helping to understand and quantify the spatial dynamics

of microbial community function in pilot-scale CWs. The microbial community function

varied based on both system design and spatial positioning within the system. Additional

system comparisons and results will be presented including metabolic richness, carbon source

guild utilization analysis, and water quality data.

REFERENCES [1] Nivala, J., et al. (in press) Comparative analysis of constructed wetlands: The design and construction of the

ecotechnology research facility in Langenreichenbach, Germany. Ecological Engineering.

http://dx.doi.org/10.1016/j.ecoleng.2013.01.035

[2] Deng, H., et al., Analysis of the Metabolic Utilization of Carbon Sources and Potential Functional Diversity

of the Bacterial Community in Lab-Scale Horizontal Subsurface-Flow Constructed Wetlands. J. Environ. Qual.,

2011. 40(6): p. 1730-1736.

[3] Weber, K.P. and R.L. Legge, Community-level physiological profiling. Methods in molecular biology

(Clifton, N.J.), 2010. 599: p. 263-281.

HAw-1HAw-2

HAw-3

HAw-4

HAc-1HAc-2HAc-3

HAc-4H25p-1H25p-2

H25p-3H25p-4

H25-1H25-2

H25-3H25-4

H50p-1H50p-2H50p-3

H50p-4

H50-1H50-2

H50-3H50-4

VAc-1

VAc-2

VAc-3

VSp-1

VSp-2

VSp-3

VS-1

VS-2

VS-3

-2.5

-2

-1.5

-1

-0.5

0

0.5

1

-2.5 -2 -1.5 -1 -0.5 0 0.5 1 1.5 2

PC

2 (

25

.03

%)

PC 1 (44.72 %)

B


133

Modeling of Carbon and Nitrogen removal kinetics in a biofilter

under different hydraulic conditions for wastewater treatment (O.166)

Audrey Soric, Ming Zeng, Nicolas Roche Aix Marseille University, Centrale Marseille, CNRS, M2P2 UMR 7340, Europole de

l’Arbois, 13545 Aix en Provence Cedex 4, FRANCE

([email protected] - [email protected] - nicolas.roche@univ-

amu.fr)

INTRODUCTION

Nowadays biofilm reactors are widely used in wastewater treatment. There exist many

configurations of these reactors. In these reactors removal efficiency of organic carbon and

nitrogen pollutants is due to microbial processes and activity. Because in Constructed

Wetlands (CW) removal of most pollutant is due to microbial processes of the biofilm

attached on the media they can be studied as biofilm reactors too (J.L. Faulwetter et al, 2009).

In this study the influence of hydrodynamic behavior on carbon and nitrogen removal

kinetic of a biofilm reactor was experienced and modeled as a function hydraulic conditions

(trickling bed, submerged or half submerged bed).

METHODS

Two bioreactors - cylinder columns with surface area of 78 cm2 and height of 110 cm -

working in parallel inoculated with WWTP sludge were continuously fed with a surrogate

wastewater containing organic carbon, ammonia, inorganic carbon and trace elements. Plastic

rings Anox Kaldnes K1 were used as biofilm carriers. Effluents were sampled respectively at

the depths of 10 cm (1), 30 cm (2), 70 cm (3), 90cm and 110 cm of filter bed. Compressed air

was injected at the filter depth of 70 cm at the flow rate of 0.4 L/min to produce aerobic and

anoxic conditions respectively in the top and the bottom of the reactor (Figure 1).

Figure 1 Biofim reactor with its layout modules in GPS-X software

Two Hydraulic Loading Rate (HLR) of 0.5 and 1 L.h-1

were applied to the reactors in

different hydraulic configurations (full saturated, half-saturated and unsaturated) and different

C/N ratio in the range of 8 to 1. Periodically the hydrodynamic behavior of the reactors were


134

studied by Retention Time Distribution (RTD) analysis by the means of a pulse injection of

10-40 mL of concentrated NaCl solution in the influent of the reactor. Then the electrical

conductivity of the outlet of the reactor was recorded for several hours. The Break Through

Curves (BTCs) analysis helped to calibrate the hydrodynamic model. In parallel biofilm

activity and biodegradation efficiency were determined by respirometric tests, batch tests and

TOC/TN (Total Organic Carbon and Total nitrogen) analyses of the effluent for different

positions.

The number of tanks calculated from the calibration of the hydrodynamic model and the

biokinetic parameters from respirometric and batch tests were then implemented in GPS-X

software. This software is a user-friendly simulator of WWTP units which is capable of

predicting the removal of suspended solid and conventional aquatic chemical indicators in

suspended sludge reactor as well as in biofilm reactor.


Firstly in situ batch experiments led to determine Monod kinetic parameters for carbon

biodegradation as a function of the position of the biofilm in the height of the reactor. Then

experimental data of TOC concentration of the effluent were compared to simulations for

different hydraulic conditions (Table 1). Table 1: Comparison of experimental and simulated TOC concentrations and removal rate under

different hydraulic conditions

Hydraulic

conditions

Influent TOC

(mg/L)

Effluent TOC

(mg/L)

TOC removal rate

(%)

Exp. Sim. Exp. Sim.

Full saturated 228±8 57±10 75±5 75 67

Half saturated 252±25 82±14 81±15 68 68

Unsaturated 296±34 93±7 113±3 69 62

Sensitivity analysis of filter depth to carbon removal and nitrification were carried out. It

showed that at any ammonia inlet concentration TOC was completely removed in the first

part of the filter depth (about 10 cm). These results explain the very low influence of

hydraulic conditions on TOC removal rate. Nevertheless, nitrification efficiency gradually

increased with the depth of the filter and ideal partial nitrification ratio of NH4:NO2 of 1

appeared at a depth of 20 and 30 cm for low and high NH4 concentrations respectively.

Comparison of experimental and simulated data on NH4 removal and NO2 production

showed a very good agreement between experimental and simulated results.

CONCLUSION

Experimental results led to calibrate a model of biofilm reactor associating hydrodynamic

and kinetic parameters. The simulated results demonstrated that this model is an efficient tool

for carbon and nitrogen removal simulation in biofilm reactors. Concerning operating

parameters, hydraulic conditions have low effect on TOC removal whereas nitrification is

more influenced by the depth of the filter.

REFERENCES Jennifer L. Faulwetter et al (2009) Microbial Processes influencing performance of treatment wetlands: A

review. Ecol. Eng. 35:987-1004

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