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Influence of hydrological connectivity of riverine wetlands on nitrogen removal via denitrification

E. Racchetti • Marco Bartoli • E. Soana •

D. Longhi • R. R. Christian • M. Pinardi •

P. Viaroli

Received: 30 April 2009 / Accepted: 25 May 2010 / Published online: 17 June 2010

� Springer Science+Business Media B.V. 2010

Abstract Wetland ecosystems in agricultural areas

often become progressively more isolated from main

water bodies. Stagnation favors the accumulation of

organic matter as the supply of electron acceptors

with water renewal is limited. In this context it is

expected that nitrogen recycling prevails over nitro-

gen dissipation. To test this hypothesis, denitrifica-

tion rates, fluxes of dissolved oxygen (SOD),

inorganic carbon (DIC) and nitrogen and sediment

features were measured in winter and summer 2007

on 22 shallow riverine wetlands in the Po River Plain

(Northern Italy). Fluxes were determined from incu-

bations of intact cores by measurement of concen-

tration changes or isotope pairing in the case of

denitrification. Sampled sites were eutrophic to

hypertrophic; 10 were connected and 12 were

isolated from the adjacent rivers, resulting in large

differences in nitrate concentrations in the water

column (from \5 to 1,133 lM). Benthic metabolism and denitrification rates were investigated by two

overarching factors: season and hydrological

connectivity. SOD and DIC fluxes resulted in respi-

ratory quotients greater than one at most sampling

sites. Sediment respiration was coupled to both

ammonium efflux, which increased from winter

to summer, and nitrate consumption, with higher

rates in river-connected wetlands. Denitrification

rates measured in river-connected wetlands (35–

1,888 lmol N m-2 h-1) were up to two orders of magnitude higher than rates measured in isolated

wetlands (2–231 lmol N m-2 h-1), suggesting a strong regulation of the process by nitrate availabil-

ity. These rates were also significantly higher in

summer (9–1,888 lmol N m-2 h-1) than in winter (2–365 lmol N m-2 h-1). Denitrification supported by water column nitrate (DW) accounted for 60–

100% of total denitrification (Dtot); denitrification

coupled to nitrification (DN) was probably controlled

by limited oxygen availability within sediments.

Denitrification efficiency, calculated as the ratio

between N removal via denitrification and N regen-

eration, and the relative role of denitrification for

organic matter oxidation, were high in connected

wetlands but not in isolated sites. This study confirms

the importance of restoring hydraulic connectivity of

riverine wetlands for the maintenance of important

biogeochemical functions such as nitrogen removal

via denitrification.

Keywords Denitrification � Benthic respiration � N-regeneration � Hydrological connectivity � Riverine wetlands

E. Racchetti � M. Bartoli (&) � E. Soana � D. Longhi � M. Pinardi � P. Viaroli Department of Environmental Sciences, University

of Parma, Viale G.P. Usberti 33/A, 43100 Parma, Italy

e-mail: [email protected]

R. R. Christian

Department of Biology, East Carolina University,

Greenville, NC, USA

123

Biogeochemistry (2011) 103:335–354

DOI 10.1007/s10533-010-9477-7

Introduction

Coastal eutrophication and dystrophy have become

common globally as nutrient inputs from rivers,

particularly nitrogen loads, result in profound modi-

fication of coastal biogeochemistry (Nixon 1995;

Smith et al. 2006; Diaz and Rosenberg 2008). Thus,

both the scientific and environmental management

communities interested in nutrient loadings have

increased focus on those inland areas where nutrients

are generated and on nutrient transport and transfor-

mations in freshwater bodies (Schaller et al. 2004;

Arango et al. 2006; Boyer et al. 2006). In northern Italy

nitrate contamination of surface and groundwater is

particularly relevant in the Po River Plain, and the

potential for removal of nitrogen has significant

consequences to eutrophication of receiving coastal

ecosystems. The Po River is one of the most significant

freshwater source to the Mediterranean basin and

affects the trophic state of the shallow northern

Adriatic Sea (Provini and Binelli 2006).

Nitrogen transformations result from different bio-

geochemical processes including biological uptake,

abiotic sedimentation, nitrification, ammonification,

nitrate reduction to ammonium, denitrification and

anammox, the anaerobic oxidation of ammonium

(Brandes et al. 2007; Burgin and Hamilton 2007;

Jetten 2008). Among these, microbially mediated

denitrification and anammox are the only processes

that remove inorganic nitrogen from land and water

and displace it into the atmosphere (Schubert et al.

2006; Seitzinger et al. 2006). Studies reporting simul-

taneous measurements of denitrification and anammox

in eutrophic freshwater environments are very limited

but indicate that the contribution of denitrification to

N2 fluxes is generally [90% (Trimmer et al. 2003; Schubert et al. 2006; Koop-Jakobsen and Giblin 2009).

Among freshwater and marine ecosystems, the most

efficient nitrate removal has been measured in wet-

lands, largely due to the anoxic and organic wetland

sediments (Bachand and Horne 2000). But not all

wetlands function similarly (Brinson 1993a, b; Brin-

son et al. 1998; Mitsch and Gosselink 2000; Lin 2006),

and there is a need to determine how different classes

of wetlands contribute to nitrogen transformations and

removal.

Hydrogeomorphic conditions of wetlands affect

the physicochemical characteristics of water and

sediments which in turn regulate denitrification and

nitrogen removal. Hydrology, climate and basin

geomorphology are main components that determine

establishment and maintenance of a wetland’s

hydrogeomorphology (Brinson 1993a, b; Mitsch and

Gosselink 2000). Among hydrogeomorphic condi-

tions, hydrological connectivity is relevant in those

heavily exploited geographical areas where (1) the

control of water flow in rivers impedes flooding and

lateral interaction with riparian areas and (2) land

use by agriculture has turned wetlands into isolated

patches. In waterlogged soils, nitrate availability,

autochthonous organic matter and anaerobic condi-

tions are essential factors for the denitrification

process, but they are not sufficient to establish a

significant nitrogen removal efficiency. Morphology

and topographic position in the landscape, together

with the distribution of water sources (groundwater,

surface inflow or precipitation), turnover time and

hydrodynamics influence the degree to which chem-

icals are transported to or from wetlands (Brinson

1993a, b; Brinson et al. 1998).

Understanding how denitrification is regulated in

wetlands is a difficult target as environmental

features often interact dynamically, making it diffi-

cult to identify a prevailing factor, if indeed one

exists. Nitrate is supplied to sediments from nitrifi-

cation or by diffusion from the water column and its

concentration is in general positively correlated with

denitrification rates (Reinhardt et al. 2006; Sirived-

hin and Gray 2006). In anaerobic wetland sediments,

coupled nitrification–denitrification is limited, so the

anaerobic respiration may be supported by external

nitrate input—e.g. by agricultural runoff or previ-

ously nitrified sewage effluent (Verhoeven et al.

2006). Nitrate advected through the sediment from

groundwater is potentially an important source,

especially in areas with contaminated aquifers (Sei-

tzinger et al. 2006). Co-occurring processes often

make it difficult to separate the effect of temperature

on denitrification rates alone. Increasing tempera-

tures stimulate ammonification, nitrification and

benthic respiration that could promote denitrification

process (Bachand and Horne 2000; Sirivedhin and

Gray 2006; Hernandez and Mitsch 2007). Simulta-

neously, at elevated temperatures, low oxygen

solubility and high rates of aerobic respiration result

in oxygen deficits and enhanced anaerobic mineral-

ization processes, such as sulfate reduction, that

may inhibit denitrification and favor DNRA, the

336 Biogeochemistry (2011) 103:335–354

123

dissimilatory reduction of nitrate to ammonium

(Brunet and Garcia-Gil 1996; Piña-Ochoa and

Álvarez-Cobelas 2006). Increasing temperatures

enhance also macrophytic or microphytobenthic N

uptake, a process that competes with nitrification and

denitrification (Sundbäck et al. 2000; Bartoli et al.

2003; Risgaard-Petersen 2004). Nitrate removal in

wetlands is also regulated by the availability of

organic matter and its macromolecular quality

(Bastviken et al. 2005; Sirivedhin and Gray 2006).

Further, nitrate removal efficiency is inversely

correlated with the ratio between litter C and

NO3 -

concentration in the water column, while

other pathways, such as DNRA, are favored (Tiedje

1988; Ingersoll and Baker 1998; Gardner and

McCarthy 2009). Water pH is rarely examined as

it is usually not perceived as detrimental to denitri-

fying metabolism (Kadlec and Knight 1996).

The overall aim of this study was to evaluate the

factors regulating denitrification and assess the

importance of hydrological connectivity of riverine

wetlands on nitrogen removal efficiency. We hypoth-

esize that wetlands with limited hydrological con-

nectivity with adjacent rivers are characterized by

low nitrogen removal via denitrification. We mea-

sured nitrate removal via denitrification in conjunc-

tion with benthic respiration, inorganic nitrogen

exchange and, potentially regulating, environmental

factors. Denitrification was measured consistently

by the isotope pairing technique (IPT) in a number

of shallow riverine wetlands located in the Po

River Plain (Northern Italy). The wetlands were

characterized by degree of connectivity with the

adjacent water bodies (the Po, the Oglio and the

Mincio Rivers). To our knowledge, no studies have

reported denitrification and benthic flux measure-

ments in shallow freshwater wetlands of the Po Plain,

despite their recognized importance as nitrogen

receivers and transformers. This study represents

one of the most spatially extensive investigation of

nitrogen cycling within the wetlands of a large

watershed using the same methodological approach.

Materials and methods

Study area

All 22 sites from this study are small-sized and

shallow aquatic wetlands located in the Po River

Plain, within the basins of the Po, Oglio and Mincio

Rivers, in the provinces of Cremona and Mantova,

Northern Italy (Fig. 1; Table 1). The Po River Plain is

characterised by intensive agricultural practices, live-

stock farming and breeding (Marchetti 1993). The

studied wetlands are eutrophic to hypertrophic and are

undergoing rapid infilling; most of them are included

in the Oglio Nord, Oglio Sud, and Mincio Natural

Parks. We sampled seven oxbow lakes or old mean-

ders, six ponds, four freshwater marshes of which

three are old peat bogs, two riverine wetlands

and three shallow eutrophic lakes. All are colonized

by aquatic plants (emergent, submersed, floating

macrophytes and pleustophytes) and have permanent

Oglio

Mincio

MI1 MI3

MI4 OG3

OG11

OG7

OG2

PO7

OG6 OG8

OG9

OG10

OG1

N 0 10 20 km

PO2

MI2

PO1

PO3 PO4

PO5 PO6

OG4 OG5

Oglio

Mincio

MI1 MI3

MI4 OG3

OG11

OG7

Fig. 1 Location of the 22 sampling sites in the Po

River Plain (Northern Italy)

Biogeochemistry (2011) 103:335–354 337

123

T a

b le

1 M

a in

fe a tu

re s

o f

th e

sa m

p li

n g

si te

s; in

th e

se c o

n d

c o

lu m

n ,

C a n

d I

in d

ic a te

c o

n n

e c te

d a n

d is

o la

te d

w e tl

a n

d s

L o

c a ti

o n

(c o

d e )

H y

d ro

lo g

ic a l

c o

n n

e c ti

v it

a n

d w

a te

r so

u rc

T y

p o

lo g

y G

e o

g ra

p h

ic a l

c o

o rd

in a te

(G a u

ss –

B o

a g

a )

S u

rf a c e

(m 2 )

M e a n

d e p

th (m

)

D o

m in

a n

t m

a c ro

p h

y te

L a ti

tu d

e L

o n

g it

u d

G a b

b io

n e ta

(O G

1 )

C O g

li o

R iv

e r,

su rf

a c e

in fl

o w

O x

b o

w la

k e

5 0

0 8

5 1

5 9

8 2

8 0

,7 0

0 0

.5 P

h ra

g m

it e s

a u

st ra

li s,

T y p

h a

la ti

fo li

a ,

C a

re x

sp p

G a m

b a ra

(O G

2 )

C O g

li o

a n

d G

a m

b a ra

R iv

e rs

, su

rf a c e

in fl

o w

P o

n d

5 0

0 4

8 7

5 1

6 0

2 1

9 2

4 ,1

0 0

0 .5

T y p

h a

la ti

fo g

li a

, C

a re

x sp

p .

G e rr

e G

a v

a z z i

(O G

3 )

C O g

li o

R iv

e r,

su rf

a c e

in fl

o w

O x

b o

w la

k e

5 0

0 2

8 8

0 1

6 0

5 9

6 9

3 8

,5 0

0 1

.5 P

h ra

g m

it e s

a u

st ra

li s

R u

n a te

D x

(O G

4 )

C O g

li o

R iv

e r,

su rf

a c e

in fl

o w

O x

b o

w la

k e

5 0

0 2

8 0

9 1

6 0

6 4

9 4

1 ,8

0 0

0 .8

N u

p h

a r

lu te

u m

R u

n a te

S n

(O G

5 )

C O g

li o

R iv

e r,

su rf

a c e

in fl

o w

O x

b o

w la

k e

5 0

0 2

8 1

3 1

6 0

6 6

3 6

2 ,9

0 0

0 .9

N u

p h

a r

lu te

u m

C à

N u

o v

e B

e rn

in i

(O G

6 )

C O g

li o

R iv

e r,

su rf

a c e

in fl

o w

P o

n d

5 0

0 3

0 5

5 1

6 1

0 5

2 4

6 ,6

0 0

1 .2

C e ra

to p

h y ll

u m

d e m

e rs

u m

N u

p h

a r

lu te

u m

P h

ra g

m it

e s

a u

st ra

li s

L a g

o d

i M

e z z o

1 (M

I1 )

C M in

c io

R iv

e r,

su rf

a c e

in fl

o w

S h

a ll

o w

e u

tr o

p h

ic la

k e

5 0

0 3

3 1

4 1

6 4

0 8

1 4

1 2

,4 0

0 1

.5 N

u p

h a

r lu

te u

m ,

N y m

p h

a e a

a lb

a ,

T ra

p a

n a

ta n

L a g

o d

i M

e z z o

2 (M

I2 )

C M in

c io

R iv

e r,

su rf

a c e

in fl

o w

S h

a ll

o w

e u

tr o

p h

ic la

k e

5 0

0 2

9 8

0 1

6 4

1 3

7 6

1 6

,8 0

0 2

.5 –

L a g

o d

i M

e z z o

3 (M

I3 )

C M in

c io

R iv

e r,

su rf

a c e

in fl

o w

S h

a ll

o w

e u

tr o

p h

ic la

k e

5 0

0 2

7 4

2 1

6 4

1 8

8 2

2 5

,0 0

0 1

.5 T

ra p

a n

a ta

n s

G o

it o

(M I4

) C M

in c io

R iv

e r,

su rf

a c e

in fl

o w

R iv

e ri

n e

w e tl

a n

d 5

0 1

5 6

0 2

1 6

3 4

7 7

4 5

,5 0

0 0

.6 V

a ll

is n

e ri

a sp

ir a

li s

B in

e (O

G 7

) I G

ro u

n d

w a te

r d

is c h

a rg

O x

b o

w la

k e

4 9

9 9

0 7

9 1

6 1

2 8

7 6

1 2

0 ,8

0 0

1 P

h ra

g m

it e s

a u

st ra

li s,

C a

re x

sp p

338 Biogeochemistry (2011) 103:335–354

123

T a

b le

1 c o

n ti

n u

e d

L o

c a ti

o n

(c o

d e )

H y

d ro

lo g

ic a l

c o

n n

e c ti

v it

a n

d w

a te

r so

u rc

T y

p o

lo g

y G

e o

g ra

p h

ic a l

c o

o rd

in a te

(G a u

ss –

B o

a g

a )

S u

rf a c e

(m 2 )

M e a n

d e p

th (m

)

D o

m in

a n

t m

a c ro

p h

y te

L a ti

tu d

e L

o n

g it

u d

M o

si o

(O G

8 )

I G ro

u n

d w

a te

r d

is c h

a rg

M a rs

h (o

ld p

e a t

b o

g )

4 9

9 9

2 7

6 1

7 4

8 4

2 ,3

0 0

0 .3

P h

ra g

m it

e s

a u

st ra

li s,

T y p

h a

la ti

fo li

a ,

C a

re x

sp p

M a rc

a ri

a (O

G 9

) I G

ro u

n d

w a te

r d

is c h

a rg

M a rs

h (o

ld p

e a t

b o

g )

4 9

9 7

2 6

0 1

6 2

1 1

4 0

1 0

,9 0

0 0

.5 P

h ra

g m

it e s

a u

st ra

li s,

C a

re x

sp p

B e lf

o rt

e (O

G 1

0 )

I G ro

u n

d w

a te

r d

is c h

a rg

M a rs

h (o

ld p

e a t

b o

g )

4 9

9 2

7 8

6 1

6 2

1 0

2 3

3 ,9

0 0

0 .3

N u

p h

a r

lu te

u m

, N

y m

p h

a e a

a lb

a ,

P h

ra g

m it

e s

a u

st ra

li s,

T y p

h a

la ti

fo li

a ,

C a

re x

sp p

B o

g in

a (O

G 1

1 )

I G ro

u n

d w

a te

r d

is c h

a rg

R iv

e ri

n e

w e tl

a n

d 4

9 8

6 5

4 8

1 6

2 5

5 9

2 6

6 ,2

0 0

0 .5

P h

ra g

m it

e s

a u

st ra

li s,

T y p

h a

la ti

fo li

a ,

C a

re x

sp p

G e rr

e d

e ’

C a p

ri o

li (P

O 1

) I G

ro u

n d

w a te

r d

is c h

a rg

O x

b o

w la

k e

4 9

9 2

2 0

9 1

5 8

2 4

7 4

1 ,5

0 0

0 .5

P h

ra g

m it

e s

a u

st ra

li s

L a n

c o

n e

(P O

2 )

I G ro

u n

d w

a te

r d

is c h

a rg

O x

b o

w la

k e

4 9

8 6

9 1

6 1

5 8

2 4

3 7

6 1

,9 0

0 0

.4 T

ra p

a n

a ta

n s

L a z z a re

tt o

(P O

3 )

I G ro

u n

d w

a te

r d

is c h

a rg

P o

n d

4 9

9 1

4 2

4 1

5 8

4 9

5 8

1 2

,5 0

0 1

.5 S

p ir

o d

e la

p o

ly rr

h iz

a ,

L e m

n a

g ib

b a

, W

o lf

fi a

a rr

h iz

G e rr

e U

g o

la n

i (P

O 4

) I G

ro u

n d

w a te

r d

is c h

a rg

P o

n d

4 9

8 9

5 3

9 1

5 8

4 9

9 7

3 ,9

0 0

0 .8

C e ra

to p

h y ll

u m

d e m

e rs

u m

M y ri

o p

h y ll

u m

sp ic

a tu

m ,

P o

ta m

o g

e to

n c ri

sp u

C à

d e ’

G a tt

i (P

O 5

) I G

ro u

n d

w a te

r d

is c h

a rg

P o

n d

4 9

9 1

3 5

3 1

5 8

9 2

8 2

2 ,8

0 0

1 .4

S p

ir o

d e la

p o

ly rr

h iz

a ,

L e m

n a

m in

o r,

L e m

n a

g ib

b a

, W

o lf

fi a

a rr

h iz

C a sc

in a

S .

M a rg

h e ri

ta (P

O 6

) I G

ro u

n d

w a te

r d

is c h

a rg

P o

n d

4 9

9 1

2 7

4 1

5 9

0 2

9 5

2 ,0

0 0

0 .7

N u

p h

a r

lu te

u m

C e ra

to p

h y ll

u m

d e m

e rs

u m

O st

ig li

a (P

O 7

) I G

ro u

n d

w a te

r d

is c h

a rg

M a rs

h 4

9 9

6 1

5 7

1 6

6 5

8 6

8 1

4 ,8

0 0

0 .9

P h

ra g

m it

e s

a u

st ra

li s,

T y p

h a

sp p

., C

a re

x sp

p .

Biogeochemistry (2011) 103:335–354 339

123

standing water (depth [0.25 m). We categorized the 22 wetland sites by both Hydrogeomorphic Classifi-

cation (Brinson et al. 1998; Mitsch and Gosselink

2000; Lin 2006) and Functional Classification of

European Wetlands (Simpson 2002). The wetlands

exhibit different degrees of hydraulic connectivity

with the adjacent rivers (Table 1). Connectivity was

assessed from map elevation and hydrological data

(water levels) of the considered rivers (Technical

Regional Maps of Lombardy Region http://www.

cartografia.regione.lombardia.it; AdBPo 2001–2008).

All wetlands were also directly surveyed in order to

detect interventions/alterations, e.g. embankments,

levees or berms. Ten sites are permanently connected:

OG1–OG6 and MI4 have an inlet river channel while

MI1–MI3 represent a river digression with unidirec-

tional water movement. These ten wetlands are clas-

sified as ‘‘Riverine wetland with surface inflow and

unidirectional flow’’ (Brinson et al. 1998; Mitsch and

Gosselink 2000; Lin 2006) or as ‘‘Eastern Continental

River Marginal Wetlands belonging to unconfined

channel sub-type—ECRM2H3P’’ (Simpson 2002).

The remaining 12 wetlands are isolated because of

canal infilling, weirs and embankments. The primary

water sources in the isolated wetlands are ground-

water, precipitation and runoff, with the water

movement due largely to vertical fluctuation. They

are classified as ‘‘Isolated depression supplied by

groundwater discharge and characterized by vertical

fluctuation’’ (Brinson et al. 1998; Mitsch and Gosse-

link 2000; Lin 2006) or as ‘‘Eastern Continental River

Marginal Wetlands belonging to separated channel

sub-type—ECRM3H2P’’ (Simpson 2002).

Sampling program

We sampled water and sediments during winter

(January–March) and summer (June–August) 2007.

Intact sediments were collected at each site with

transparent Plexiglas cores with different dimensions

(diameter 9 length) for sediment characterization

(5 9 30 cm, n = 3) and flux measurements (8 9

40 cm, n = 4). All cores were transferred in a box

with cooled site water and brought to the laboratory

within a few hours of sampling for further processing

and incubation. Also, on each sampling date, approx-

imately 100 l of water were collected and brought to

the laboratory for cores maintenance and incubations.

During sampling, site water was characterized for

temperature, pH, dissolved oxygen with a YSI Multi-

Probe (model 556, Yellow Springs, OH, USA) and

dissolved inorganic carbon and nitrogen (see later for

analytical methods).

Sediment characteristics

Benthic microalgal biomass was measured as chlo-

rophyll-a (Chl-a) concentration in the top 0.5 cm of

sediment and determined spectrophotometrically

after extraction with 90% acetone (Lorenzen 1967).

Bulk density was determined as the ratio between wet

weight and volume (typically 5 ml) of sediment.

Organic matter content (OM) was measured as

percentage of weight loss by ignition (450�C, 2 h) from dried sediment. Total C and N content were

determined from dried sediments with a Carlo Erba

elemental analyzer (CHNS-O EA 1108).

Dissolved oxygen, inorganic carbon and nutrient

flux measurements

Intact cores preincubation and incubation proce-

dures followed the standardized protocol described

by Dalsgaard et al. (2000). In the laboratory, all

cores were immediately submersed with the top

open in a tank containing in situ aerated and well

mixed water at ambient temperature. During prein-

cubation, water was stirred inside the tubes and

headwater exchange with the tank water was

ensured by Teflon-coated magnetic bars suspended

5 cm above the sediment–water interface to mini-

mize particle resuspension and driven by an external

motor at 40 rpm.

The day after the sampling, the water in the tank

was exchanged and the cores were incubated for

flux measurements of O2 (or sediment oxygen

demand, SOD), dissolved inorganic carbon (DIC),

and dissolved inorganic nitrogen (DIN = NH4 ?

NO2 -

? NO3 -

). Incubations were started by lower-

ing the water in the tank just below the top of the

cores and by sealing the cores with floating lids

provided with a sampling port. The cores were

incubated in the dark at field temperature for 2–5 h

with continuous stirring. Incubation time was set to

keep oxygen concentration at the end to within 20%

of the initial value.

340 Biogeochemistry (2011) 103:335–354

123

http://www.cartografia.regione.lombardia.it

Water samples (60 ml) were collected at regular

time intervals using plastic syringes from the water

phase (750 ml). Samples for O2 determinations were

transferred to glass vials (Exetainer, Labco, High

Wycombe, UK) and Winkler reagents were added

immediately (Strickland and Parsons 1972). Samples

for DIC were also transferred in glass vials and

immediately titrated with 0.1 N HCl (Anderson et al.

1986). Samples for NH4 ?

, NO3 -

and NO2 -

determi-

nations were filtered through Whatman GF/F glass

fiber filters, transferred to plastic vials and frozen.

NH4 ?

was determined spectrophotometrically using

salicylate and hypochlorite in the presence of sodium

nitroprussiate (Bower and Holm-Hansen 1980).

NO3 -

was determined after reduction to NO2 -

the presence of cadmium, and NO2 -

was determined

spectrophotometrically using sulphanilamide and

N-(1-naphtyl)ethylendiamine (Golterman et al. 1978).

Fluxes of O2, DIC, NH4 ?

, NO2 -

and NO3 -

were

calculated from the changes in concentrations in the

cores with time by linear regression and expressed as

rate per square meter. Negative fluxes indicate flux

from the water column to the sediment while positive

fluxes indicate effluxes from sediment to the water

column.

From dark fluxes of inorganic carbon and oxygen

we calculated the respiratory quotient RQ (Dilly

2003; Hargrave et al. 2008), according to the

following equation:

RQ ¼ CO2 flux

O2 fluxj j :

Denitrification measurements

The isotope pairing technique (Nielsen 1992) was

used to measure denitrification (dark rates) on the

same set of cores used for solute fluxes. Incubations

for solute fluxes and denitrification were sequential

and both performed the day after the sampling. The

water in the tank was renewed between the two

incubations, and the open cores were submersed for a

couple of hours. The IPT allows for differentiation of

total denitrification (Dtot), denitrification of nitrate

diffusing to the anoxic sediment from the water

column (DW) and denitrification of nitrate produced

within the sediment due to nitrification (DN).

Methodological concerns have been raised about

the IPT, mainly due to the concurrence of anammox

which cannot be discriminated from denitrification as

a source of N2 and makes invalid the assumptions on

which IPT calculations are based. Recent studies

showed that anammox can contribute half or more of

the N2 production in coastal shelves and deep seas

(Dalsgaard et al. 2005; Trimmer et al. 2006), while it

represents a minor fraction of N2 production in

eutrophic and organic-rich freshwater ecosystems

(Burgin and Hamilton 2007; Koop-Jakobsen and

Giblin 2009). Therefore, we adopted the IPT as we

considered this method as accurate for freshwater

eutrophic sites, assuming that the anammox contri-

bution was negligible to N2 fluxes (Risgaard-Petersen

et al. 2003; Trimmer et al. 2003; Schubert et al. 2006;

Koop-Jakobsen and Giblin 2009).

At the beginning of the experiment, the water in

the tank was lowered just below the top of the cores;

and a water subsample (5 ml) was taken from each

core for NO3 -

concentration measurement. Different

amounts of 15

NO3 -

from a 15 mM Na 15

NO3 solution

were then added to the water column of each of the 4

replicate cores to perform a concentration series

experiment (see Dalsgaard et al. 2000). Labelled

nitrate was added to the cores to have a final 15

atom% of at least 30%. Within 5 min of the addition

of 15

NO3 -

another water sample was collected from

each core to calculate the 14

N/ 15

N ratio in the NO3 -

pool; the cores were then closed with floating lids and

the incubation started (3–5 h). At the end of the

incubation, 5–10 ml of ZnCl2 (7 M) was added to the

water phase, and then sediment and water were

mixed. An aliquot of the slurry was transferred to a

12.5 ml gas-tight vial; 14

N 15

N and 15

N 15

N abundance

in N2 were analyzed by mass spectrometry at the

National Environmental Research Agency, Silkeborg,

Denmark (Risgaard-Petersen and Rysgaard 1995).

The denitrification rates were calculated according to

the equations and assumptions of Nielsen (1992).

Denitrification rates were also estimated from

sediment oxygen demand and concentrations of

nitrate and dissolved oxygen with the model by

Christensen et al. (1990):

DW ¼ FO2 � a �

ffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi

1 þ DNO�

DO2 � CNO�

CO2 � 1

� �

� 1 " #

where FO2 is the dark sediment oxygen demand, a is the ratio between depth-specific denitrification and

O2 consumption activity (*0.8, Christensen et al. 1990), DNO�

3 and DO2 are the diffusion coefficients of

Biogeochemistry (2011) 103:335–354 341

123

nitrate and oxygen, and CNO� 3

and CO2 are the water

column nitrate and oxygen concentrations, respec-

tively. Modeled rates represent denitrification of

nitrate diffusing into anoxic sediments from the

water column.

Statistical analyses

Statistical analyses were performed with SPSS (ver.

16.0) to test for significant differences between

seasons and between hydrological connectivity of

wetlands. Inferential testing was done for every

measured variable (temperature, sedimentary oxygen

demand, dissolved inorganic carbon fluxes, respira-

tory quotient, ammonium fluxes, oxidized forms of

inorganic nitrogen fluxes, denitrification rates,

organic matter content, benthic microalgal biomass,

and bulk density).

We did not use parametric statistics because of

violations of assumptions of normality (Kolmogorov–

Smirnov test with Lilliefors significance correction)

and homogeneity of variance (Levene test). Even after

the application of the Cox–Box method, for log

transformations, normality was not verified, largely as

a result of frequent outliers.

Therefore the non-parametric Wilcoxon paired-

sample T test was used to test the differences between

winter and summer distributions (n1 = n2 = 22), and

the Mann–Whitney U test was used to test for

differences between isolated and connected wetlands

(n1 = 24 and n2 = 20). When the Wilcoxon test was

significant, we analyzed the hydrological connectiv-

ity after separating the winter and the summer data

(n1 = 12 and n2 = 10). We identified the correlative

relationship among variables with the Spearman’s

rank correlation coefficient (q). Differences were considered non-significant if P [ 0.05.

Results

General features of water column and surface

sediments

Environmental conditions in both the water column

and sediments were measured in association with rate

measurements (Tables 2, 3). Water temperature

medians were 10�C (range 5–13�C) in winter and 25�C (range 22–27�C) in summer. Water column pH

was alkaline in both seasons and ranged between 7.21

and 8.57. DIC concentrations ranged from 1.90 to

7.07 mM in winter and from 2.07 to 6.63 mM in

summer; no significant differences were found

between seasons and between connected and isolated

sites. Dissolved CO2 saturation, calculated from

temperature, pH and DIC data, indicated a general

supersaturation in the water column with significant

differences between winter and summer periods

(median 217%, range 147–464% and median 581%,

range 155–4,173%, respectively; Wilcoxon T test,

P \ 0.001). In particular, for summer, we found higher percentages of CO2 saturation values in

isolated than in connected sites (Mann–Whitney U

test, P \ 0.05). Dissolved O2 concentrations were significantly

higher in winter than in summer (Wilcoxon T test,

P \ 0.01), while percentages of dissolved O2 satura- tion were similar between seasons. In winter, median

values of % O2 saturation were close to 100% for

both wetland types (97% in isolated and 85% in

connected sites). In summer, % O2 saturation tended

to be lower at isolated sites (median value 63%) and

hypoxia established in wetlands colonized by Lemn-

aceae (Table 2).

Nitrate concentrations were extremely variable

and were not significantly different between seasons

(Wilcoxon T test, P = 0.3) (Table 2). In wetlands

permanently connected with the river, median values

of nitrate concentrations were not different between

seasons (360 lM in winter and 367 lM in summer), but were significantly higher than those measured at

isolated sites (Mann–Whitney U tests, P \ 0.001). Here, nitrate contents were significantly greater in

winter (median 10 lM) than in summer (median 2 lM). According to the predictions of the model proposed by Christensen et al. (1990), the NO3

- :O2

ratios of connected sites were theoretically favorable

to denitrification, with similar or higher availability

of nitrate compared to oxygen in bottom waters

(Table 2).

Sediment OM content, Chl-a concentrations and

bulk density were not statistically different between

seasons (Table 3). OM content and bulk density

ranges were typical of environments characterized by

elevated sedimentation rates, soft and reducing

sediment, and anaerobic metabolism. Sediment OM

content was significantly higher and bulk density was

significantly lower in isolated compared to connected

342 Biogeochemistry (2011) 103:335–354

123

T a

b le

2 H

y d

ro c h

e m

ic a l

v a ri

a b

le s

m e a su

re d

in w

a te

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m p

le s

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ll e c te

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fr o

m th

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m p

li n

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te s

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o rt

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; in

th e

se c o

n d

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lu m

n ,

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d I

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c o

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d is

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te d

w e tl

a n

d s

S it

e s

H y

d ro

lo g

ic a l

c o

n n

e c ti

v it

y p

H [D

IC ]

(m M

) C

O 2

sa t

(% )

[O 2 ]

(l M

) O

2 sa

t (%

) [N

O 3 -

] (l

M )

[N O

3 -

]: [O

2 ]

W in

te r

S u

m m

e r

W in

te r

S u

m m

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W in

te r

S u

m m

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te r

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m m

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O G

1 C

8 .3

6 8

.1 2

6 .0

7 6

.1 0

3 2

6 7

9 1

3 7

7 1

1 9

1 0

7 4

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3 1

.3 4

3 .0

O G

2 C

8 .2

9 8

.3 8

6 .1

4 7

.0 3

3 8

8 5

2 0

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8 1

4 7

1 1

6 5

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9 1

.2 9

3 .2

O G

3 C

8 .3

5 8

.0 1

6 .5

2 6

.3 9

3 2

8 1

0 1

1 1

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8 7

4 6

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8 .2

O G

4 C

8 .2

0 8

.2 5

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1 5

.2 9

3 5

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1 0

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8 .1

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6 5

.0 8

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3 0

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.1 4

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6 C

8 .3

5 8

.0 4

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9 4

.5 4

3 2

5 7

5 6

2 8

9 2

5 2

8 2

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8 4

.2 1

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C 8

.2 5

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.0 3

3 .3

2 1

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.1 7

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.2 6

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4 3

.2 5

3 .3

1 1

8 7

1 6

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0 0

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C 8

.2 3

8 .5

1 3

.0 3

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1 7

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C 8

.1 9

8 .1

2 2

.2 4

2 .3

1 2

6 3

3 3

0 3

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9 9

1 8

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1 1

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2 0

.5 1

O G

7 I

8 .5

7 8

.2 6

7 .0

3 6

.2 1

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9 I

8 .3

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7 4

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0 .0

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1 0

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.4 3

8 .0

9 5

.0 8

4 .3

9 2

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0 .0

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.0 0

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1 1

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.4 6

7 .3

1 6

.0 6

3 .1

6 2

2 1

2 2

9 9

2 2

4 5

6 4

2 0

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0 9

0 .3

6 0

.0 2

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1 I

8 .2

0 8

.2 9

3 .1

5 3

.0 5

2 5

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0 7

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.0 8

0 .0

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2 I

8 .1

3 8

.1 3

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0 3

.3 3

1 6

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.0 1

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8 .4

4 8

.0 1

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8 4

.0 1

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3 7

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2 7

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.0 1

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4 I

8 .3

7 8

.2 6

3 .3

7 2

.0 7

1 4

7 1

9 7

4 8

2 2

9 1

1 3

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1 0

1 3

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.0 3

0 .0

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5 I

8 .3

6 8

.0 8

3 .0

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.0 1

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P O

6 I

8 .1

3 8

.0 6

3 .0

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0 .0

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.0 2

P O

7 I

8 .2

0 8

.2 6

3 .0

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sa m

p le

s w

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c o

ll e c te

d c lo

se to

th e

se d

im e n

t su

rf a c e

Biogeochemistry (2011) 103:335–354 343

123

wetlands (Mann–Whitney U tests, P \ 0.01 and P \ 0.001, respectively), likely due to their pensile and stagnant condition. Higher values of OM were

observed in marshes (old peat bogs) and in ponds

characterized by pleustonic communities. Chl-a con-

centrations were high and similar among sites and

seasons. Sediment C:N ratio displayed a large

variability among sites and no statistical differences

were found between wetland types in both winter and

summer.

Benthic processes

As expected, season had effects on benthic processes,

and these are likely related to temperature differ-

ences. SOD was significantly correlated with water

temperature (Spearman q = -0.379, P \ 0.05) and with bulk sediment density (Spearman q = 0.332, P \ 0.05) but not with OM content in sediments (Spearman q = -0.162, P = 0.3); SOD increased from winter to summer (median values -1.14 and

-2.54 mmol O2 m -2

h -1

, respectively) (Wilcoxon T

test, P \ 0.05) (Fig. 2). Significant differences of SOD rates were found between isolated and con-

nected wetlands only in winter (Mann–Whitney U

tests, P \ 0.01) whilst in summer no statistical differences resulted (Mann–Whitney U tests,

P = 0.1) (Fig. 2). In a number of sites, DIC effluxes

tended to increase from winter to summer, but they

were not significantly correlated with water temper-

ature (Spearman q = 0.193, P = 0.3). Differences between seasons were also not significant (Wilcoxon

T test, P = 0.2) with median values of 1.95 (range

0.64–16.70) and 3.61 (range 0.41–11.06) mmol DIC

m -2

h -1

in winter and in summer, respectively.

Inorganic carbon effluxes corresponding to SOD rates

in place and time resulted in variable respiratory

quotients in both seasons (RQ) (Fig. 2). RQ medians

were slightly above 1 in winter and in summer, with

values of 1.21 (range 0.52–11.33) and 1.29 (range

0.11–6.95), respectively (Wilcoxon T test, P = 0.5).

The inorganic nitrogen fluxes were evaluated for

patterns associated with season and connectivity

(Fig. 3). Ammonium fluxes underwent a great spatial

and seasonal variability. Differences between seasons

were statistically significant (Wilcoxon T test,

P \ 0.05) with lower values in winter (median -5, range -462 to 864 lmol N m-2 h-1) than in sum- mer (median 345, -676 to 2,670 lmol N m-2 h-1) when fluxes were mainly directed from sediment to

the water column (Fig. 3). No statistical differences

were identified among wetland types, although during

winter in connected wetlands ammonium flux tended

to be directed from the water to the sediment (Mann–

Whitney U test, P = 0.1). A significant correlation

was found between ammonium fluxes and SOD

(Spearman q = 0.480, P \ 0.01), temperature (Spearman q = 0.452, P \ 0.01) and bulk sediment density (Spearman q = -0.386, P \ 0.05). As a general tendency, the fluxes of the oxidized forms of

inorganic nitrogen were directed from the water

column to the sediment surface. Nitrite fluxes

(from -131 to 188 lmol N m-2 h-1, median -1 lmol N m-2 h-1) comprised between 0 and 42% (average of the whole dataset 5%) of NOx

benthic exchanges (=NO2 -

? NO3 -

). Median values

of NOx -

fluxes in winter and in summer were

respectively -44 lmol N m-2 h-1, range -1,616 to 608 lmol N m-2 h-1, and -68 lmol N m-2 h-1, range -4,062 to 2,014 lmol N m-2 h-1 (Fig. 3). NOx

- sedimentary demand did not demonstrate a

statistically significant seasonal difference and

increased in summer relative to winter in only 14

out of 22 sampling sites (Mann–Whitney U test,

Table 3 Median and range (in parenthesis) are reported for organic matter content (OM), bulk density (d), chlorophyll-a concen- trations (Chl-a) and C:N molar ratios in surface sediments (0–1 cm)

Variable Riverine connected wetlands Isolated wetlands

Winter Summer Winter Summer

OM (%) 10.4 (2.3–14.1) 10.6 (1.5–16.9) 15.9 (5.2–34.5) 17.8 (5.8–34.8)

d (g cm -3

) 1.2 (1.0–1.5) 1.1 (1–1.7) 1.0 (0.9–1.4) 0.9 (0.9–1.2)

Chl-a (lg cm-3) 24.7 (22.9–48.9) 25.6 (14.4–32.5) 23.75 (11.8–37.2) 23.15 (13.1–35.9)

C:N (mol:mol) 22.4 (10.7–27) 20.8 (10.9–25.7) 14.9 (10.9–21) 15.4 (11–24.8)

n = 10 for riverine connected wetlands and n = 12 for isolated wetlands in each season

344 Biogeochemistry (2011) 103:335–354

123

P = 0.1). Wetland type affected NOx -

fluxes (Mann–

Whitney U test, P \ 0.05) (Fig. 3). Influx rates were higher in environments connected to rivers in summer

(Mann–Whitney U test, P \ 0.05), but no effects of connectivity were found in winter (Mann–Whitney U

test, P = 0.5). A positive correlation was found

between NOx -

sedimentary demand and NO3 -

concentration (Spearman q = 0.575, P \ 0.001) and between NOx

- sedimentary demand and per-

centage of dissolved CO2 saturation in the water

column (Spearman q = 0.323, P \ 0.05). Results of ammonium and NOx

- flux measurements were

combined. No significant differences were found

between seasons or between wetland types for DIN

(NH4 ?

? NOx -

) fluxes (Fig. 3).

Denitrification rates

Nitrogen removal via denitrification was responsive to

season. Total denitrification rates (Dtot = DW ? DN)

were extremely variable and ranged between \5 and *2,000 lmol N m-2 h-1; N2 effluxes were signifi- cantly higher in summer than in winter (Wilcoxon T

test, P \ 0.01) (Fig. 4). Dtot was mostly supported by water column nitrate (DW) and exhibited the highest

values in connected wetlands in both seasons (Mann–

Whitney U tests, P \ 0.001 and P \ 0.001 in summer and winter, respectively). Both DW and DN were lower

in winter than in summer (Wilcoxon T tests, P \ 0.05). DW ranged between 2 and 300 lmol N m

-2 h

-1 in

winter and between 3 and 1,888 lmol N m-2 h-1 in summer supporting about 90 ± 18 and 71 ± 26% of

Dtot in the two periods, respectively. DN ranged from 0

to 67 and from 0 to 205 lmol N m-2 h-1 in winter and summer, respectively.

Overall, denitrification activity was affected by

wetland type and correlated with several environ-

mental factors. Comparing rates among wetland

types, we found lower DW rates in isolated sites both

in winter and in summer (Mann–Whitney U tests,

P \ 0.01 and P \ 0.001, respectively). DN rates were only significantly lower in winter (Mann–Whitney U

test, P \ 0.05). Dtot rates correlated positively with NO3

- concentrations (Spearman q = 0.748,

P \ 0.001), sedimentary NOx -

demand (Spearman

q = 0.691, P \ 0.001), dissolved inorganic carbon concentration (Spearman q = 0.339, P \ 0.05) and sediment bulk density (Spearman q = 0.341, P \ 0.05). Dtot was not correlated with sediment OM content. Similar relationships resulted for DW,

reflecting the higher importance of this fraction on

total denitrification. DN was correlated only with

temperature (Spearman q = 0.368, P \ 0.05). Increasing amounts of added

15 NO3

- in the cores’

m m

o l m

-2 h

-1

-7

-6

-5

-4

-3

-2

-1

DIC fluxes

m m

o l m

-2 h

-1

WI SI WC SC

** **

O2 fluxes +

Fig. 2 Dark sediment oxygen demand (SOD), dissolved inorganic carbon fluxes (DIC) and respiratory quotient (RQ)

measured at the 22 sampling sites and split up considering

winter (W) and summer (S) seasons and connected (C) and

isolated (I) wetland types. Seasonal differences ( ?

P \ 0.05, ?? P \ 0.01, ??? P \ 0.001) were tested with the Wilcoxon T test; hydrogeomorphic differences (* P \ 0.05, ** P \ 0.01, *** P \ 0.001) were tested with the Mann–Whitney U test

Biogeochemistry (2011) 103:335–354 345

123

water phase resulted in an immediate stimulation of

denitrification rates as D15 (i.e. denitrification of 15

NO3 -

added to the water column). The increases in

rate were linear with the nitrate enrichment, both in

some isolated wetlands with low nitrate concentration

and in a nitrate-rich, connected wetland (Fig. 5).

Finally, we applied the model of Christensen

et al. (1990) using as data input NO3 -

and O2 concentration in the water (Table 2) and dark O2 fluxes measured via cores incubation. The model is

unidimensional, works in sediments that are neither

bioturbated nor colonized by phanerogams and

predicts only rates of denitrification supported by

water column nitrate (DW). Theoretical DW values

calculated for the 22 wetlands (pooled winter and

summer data) showed a very good fitting with DW measured via the IPT (Spearman q = 0.847, P \ 0.01) only in a limited range of rates (0–140 lmol N m-2 h-1) (Fig. 6). Outputs of the model agree well with experimental results from

wetlands characterized by low NO3 -

in the water; on

the contrary predicted rates are much higher

than those measured at sites with elevated nitrate

concentrations.

Discussion

Denitrification in shallow eutrophic wetlands

compared to other ecosystems

and methodological considerations

The removal of nitrogen in freshwater wetlands is

widely reported in the literature (Piña-Ochoa and

Álvarez-Cobelas 2006). However, compared to other

aquatic environments, few direct measurements of

denitrification rates have been performed. A summary

m o

l m -2

h -1

-1000

-500

500

1000

1500

2000

2500

(a)

(b)

(c)

(d)

3000

WI SI WC SC

NH4 + fluxes +

DIN fluxes

m o l m

-2 h

-1

-4000

-3000

-2000

-1000

1000

2000

SCWI SI WC

NOX - fluxes

m o l m

-2 h

-1

-400

-300

-200

-100

100

200

300

WI SI

* *

NOx - fluxes

m o

l m -2

h -1

-5000

-4000

-3000

-2000

-1000

1000

2000

3000

Fig. 3 NH4 ?

(a), NOx -

(=NO2 -

? NO3 -

) (b, c) and DIN (d) dark fluxes measured via intact core incubations at the 22

sampling sites and split up considering winter (W) and summer

(S) seasons and connected (C) and isolated (I) wetland types.

NOx -

fluxes are shown in two graphs with different scales.

Seasonal differences ( ?

P \ 0.05, ?? P \ 0.01, ??? P \ 0.001) were tested with the Wilcoxon T test; hydrogeomorphic differences (* P \ 0.05, ** P \ 0.01, *** P \ 0.001) were tested with the Mann–Whitney U test

346 Biogeochemistry (2011) 103:335–354

123

of recent reports of denitrification measurements is

presented in Table 4. A direct comparison with our

data is somewhat limited by the fact that experi-

mental designs and methods are different among

the studies. Reported rates range between 1 and

3,188 lmol N m-2 h-1 and bracket our measured averages both in isolated, 49 lmol N m-2 h-1, and in connected, 340 lmol N m-2 h-1, wetlands (medians 30 and 286 lmol N m-2 h-1, respectively). The range of isolated wetlands within the Po watershed

compares well with other riparian wetlands and

with some constructed wetlands (Table 4). River-

connected sites have higher rates and are within

the reported ranges for constructed wetlands in

California, Sweden and Arizona. These all receive

pulsing floods by river and hydraulic connec-

tivity may explain the higher rates in literature

(12–3,188 lmol N m-2 h-1). The connected riverine wetlands that we studied

are among the ecosystems with highest denitrification

rates. Reported values are higher than those found in

marine environments and comparable to freshwater

ecosystems, in particular with rivers. Much lower

denitrification rates measured in isolated wetlands

are similar to those reported for marine areas

(Piña-Ochoa and Álvarez-Cobelas 2006).

In the recent literature a novel microbial process,

the anaerobic oxidation of ammonium, has been

found to be responsible for a major fraction of N2

m o

l N m

-2 h

-1

100

150

200

250

300

m o

l N m

-2 h

-1

200

400

600 1700

1800

1900

2000

Dtot ++

m o

l N m

-2 h

-1

200

400

600 1700

1800

1900

2000

DW +

WI SI WC SC

D N +

** **

***

(a)

(b)

(c)

*** ***

***

Fig. 4 Rates of denitrification measured with the isotope pairing technique at the 22 sites. Data are grouped in winter

(W) and summer (S) values of connected (C) and isolated (I)

wetland types and show total denitrification rates (Dtot, a), denitrification of nitrate diffusing to anoxic sediments from the

water column (DW, b) and denitrification of nitrate produced within sediments by nitrification (DN, c). Seasonal differences ( ? P \ 0.05, ?? P \ 0.01, ??? P \ 0.001) were tested with

the Wilcoxon T test; hydrogeomorphic differences (* P \ 0.05, ** P \ 0.01, *** P \ 0.001) were tested with the Mann– Whitney U test

D W

to t

(1 5 D

W +

1 4 D

W , µ

m o

l N m

-2 h

-1 )

100

200

300

400

500

600

NO3 - tot (15NO3

-+14NO3 -, µM)

0 10 20 30 40 50 60 70

670 680 690 700 710 720 730 740 750

PO6 PO4 PO2

OG10 OG8

OG2

low NO3 -

high NO3 -

Fig. 5 Results from a concentration series experiment in which increasing amounts of labelled nitrate were added to the

water phase of intact sediment cores. Sites where increasing 15

NO3 -

concentrations resulted in higher denitrification rates

are reported. Denitrification of nitrate diffusing to anoxic

sediments from the water column (DW) was generally

stimulated meaning that, at sites with low nitrate concentration

in the water column, the denitrification potential was elevated

but only partially expressed

Biogeochemistry (2011) 103:335–354 347

123

flux, in particular in deep marine environments

(Thamdrup and Dalsgaard 2002). Apart from the

relevance of anammox for the nitrogen cycle, this

finding poses serious methodological concerns on the

assumptions on which the IPT is built (Risgaard-

Petersen et al. 2003; Trimmer et al. 2006). In

particular, as the relevance of anammox to N2 total

flux (ra) increases, the risk of overestimating true

denitrification rates by IPT increases (Risgaard-

Petersen et al. 2003). There are only a few papers,

mostly dealing with deep marine ecosystems, where

anammox has been measured and rates compared

with those of denitrification. The available measure-

ments performed in eutrophic brackish or freshwater

environments indicate a minor relevance of anammox

to the overall N2 fluxes, but more investigation is

needed (Trimmer et al. 2003; Schubert et al. 2006;

Koop-Jakobsen and Giblin 2009). Using the equa-

tions reported in Risgaard-Petersen et al. (2003) and

Trimmer et al. (2006), we calculated from our

experiments the risk of overestimating denitrification

rates by combining ra and r14 (the ratio between 14

NOx -

and 15

NOx -

in the nitrate reduction zone).

Even if not properly designed for such calculations,

our denitrification experiment, based on increasing

additions of 15

NO3 -

to the water phase, allows such

estimates. Our results indicate that ra is low at both

connected and isolated sites and during both sampling

seasons (0–16%, median value 6%), while r14 is high

at connected sites ([3) and variable (0.1 \ r14 \ 3) at isolated sites. The combination of these percent-

ages gives a theoretical overestimation of N2 pro-

duction via denitrification between 0 and 26%

(median values of pooled data 7%). We recognize

that further experimental work should be addressed to

evaluate more accurately anammox rates and the

regulation of this process in freshwater wetlands.

Regulation of denitrification

In this study, denitrification was positively correlated

with nitrate concentration and was favored in con-

nected wetlands where we found higher concentra-

tions and availability of nitrate in both seasons (Fig. 4;

Table 2). NO3 -

concentrations in the water column of

connected sites were comparable to those in the

adjoining rivers (Mincio River, about 150 lM, and in the Oglio River, about 400 lM), while isolated wetlands had much lower concentrations. Rates

measured in environments hydraulically connected

to rivers were up to 1 or 2 orders of magnitude higher

than rates measured in isolated wetlands (Fig. 4).

As widely reported in the literature, water column

nitrate concentration is the main factor controlling

the kinetics of denitrification in many ecosystems

(Piña-Ochoa and Álvarez-Cobelas 2006; Seitzinger

et al. 2006). As a general rule we found that DW contributed mostly to Dtot. Nitrification in organically

500

1000

1500

2000

2500

3000 D

W t

h e

o re

tic a

l ( µ

m o

l N m

-2 h

-1 )

DW measured (µmol N m -2 h-1) DW measured (µmol N m

-2 h-1)

0 500 1000 1500 2000 2500 3000 0 20 40 60 80 100 120 140

D W

t h

e o

re tic

a l ( µ

m o

l N m

-2 h

-1 )

100

120

140(b)(a)

Fig. 6 Denitrification rates of nitrate diffusing to anoxic sediments from the water column (DW), calculated according to the model of Christensen et al. (1990), are plotted versus rates measured at all sites (a) and at sites with low water column nitrate content (b)

348 Biogeochemistry (2011) 103:335–354

123

loaded wetlands was severely limited by oxygen

availability and was a minor source of NO3 -

for

coupled denitrification. The exceptions were a few

isolated environments with little to no nitrate in the

water column during summer months and with a

sufficient amount of oxygen for nitrification to

proceed. Similar trends, reported by Piña-Ochoa and

Álvarez-Cobelas (2006), confirm that water column

nitrate regulates the ratio between DW and DN.

Our experiments using a concentration series

(increasing amounts of 15

NO3 -

added in cores) dem-

onstrated that most of the sites exhibited high denitri-

fication potential (D15), and thus the process was not

saturated. This was especially true for almost all

isolated wetlands in which the in situ process was

probably nitrogen-limited by stagnant conditions and

by limited nutrient recharge from flooding by river

pulses. In connected environments with extremely high

nitrate concentration ([500 lM), we did not measure a

corresponding increase of D15, which is indicative of

substrate saturation, with only one exception (Fig. 5).

The kinetics of denitrification are more uncertain when

nitrate concentrations are saturating and other factors

can limit the reaction, e.g. availability of labile organic

carbon (Piña-Ochoa and Álvarez-Cobelas 2006). Also,

in those environments where nitrate concentrations did

not change significantly in the two sampling periods,

denitrification rates increased from the winter to the

summer, indicating that water temperature was an

important regulating factor. Many studies reported a

breakpoint of temperature response of denitrification

that reflects the fact that rates fall non-linearly at lower

temperatures (Focht and Verstraete 1977; Hènault and

Germon 2000).

There are several potential links between denitri-

fication and other metabolisms in sediments, includ-

ing SOD. SOD in this study was in the higher range

of rates reported in the literature for wetlands,

Table 4 Denitrification rates measured in different types of shallow eutrophic freshwater environments, including results from the present study

Type of wetland Location Method for measuring

denitrification rates

Denitrification rate

(lmol N m-2 h-1) References

Riparian wetlands New Jersey

(USA)

N2 fluxes \20–260 Seitzinger (1994)

Constructed wetlands receiving

river water

California (USA) Mass balance 12–3,188 Reilly et al. (2000)

Wetland system treating effluent

from a sewage treatment plant

Netherlands Acetylene inhibition 1–360 Toet et al. (2003)

Wastewater treatment wetland Sweden Acetylene inhibition 714–1,786 Bastviken et al. (2005)

Riparian buffer zone Estonia N2 fluxes 14–571 Teiter and Mander (2005)

Constructed wetland Switzerland Isotope mass balance 3–661 Reinhardt et al. (2006)

Marsh receiving water from river Louisiana (USA) Isotope ratio mass

spectrometry (IRMS)

179–679 Yu et al. (2006)

Constructed wetlands receiving

river water

Ohio (USA) Acetylene inhibition 14–129 Hernandez and Mitsch

(2007)

Constructed marsh receiving water

pumped from river

Texas (USA) N2/Ar (MIMS) 54–278 Scott et al. (2008)

Constructed wetland receiving

water from drainage ditch

Indiana (USA) Isotope pairing 40–175 Herrman and White (2008)

(in press)

Constructed wetlands Arizona (USA) Mass balance 677–1,248 Kadlec (2008)

Constructed wetlands receiving

river water

Japan Acetylene inhibition 149–185 Zhou and Hosomi (2008)

Isolated wetlands Italy Isotope pairing 2–231 This study

Riverine connected wetlands Italy Isotope pairing 35–1,888 This study

Apart from Seitzinger (1994), reported values were calculated from original data and expressed in lmol N m-2 h-1

Biogeochemistry (2011) 103:335–354 349

123

eutrophic lakes, and rivers (Christensen et al. 1990;

Nielsen et al. 1990; Seitzinger 1994; Scott et al.

2008). Seasonal temperature was a key factor for

SOD, but unlike other reports we did not find a

significant correlation between SOD and Dtot (Sei-

tzinger 1994; Eyre and Ferguson 2002). According to

the simple but robust model proposed by Christensen

et al. (1990), denitrification can be predicted by a

combination of SOD and NO3 -

:O2 ratio. Shallow

water environments with high SOD, high nitrate

concentrations and anoxia should have a great

potential for nitrogen removal via denitrification.

The model was originally developed for Northern

European sites characterized by lower SOD and

lower NO3 -

:O2, but appears to be more generally

applicable to our systems. However, in some con-

nected wetlands where nitrate concentrations were

high, we found much higher predicted than measured

rates. This could result from saturation of denitrifi-

cation, nonlinearity of the model at extremes SOD

rates or NO3 -

:O2 ratios, occurrence of other more

favorable microbial transformation of NO3 -

DNRA or underestimation of real rates due to

insufficient 15

NO3 -

labeling (Brunet and Garcia-Gil

1996; Gardner and McCarthy 2009).

Denitrification rates were not correlated with the

large pool of sedimentary OM content (*10% or more), even though this variable is widely recognized

as an important regulator of the process (Ingersoll and

Baker 1998; Piña-Ochoa and Álvarez-Cobelas 2006;

Seitzinger et al. 2006). Nitrogen removal was also not

correlated with Chl-a content in sediments, even if

this represented a large pool of labile, easily degrad-

able organic carbon. Excess organic matter availabil-

ity at the study sites can mask the effect of these

factors.

Sensitivity of denitrification to nitrate concentra-

tion has significant implications to water quality and

management of wetlands. N removal appears to be

highly sensitive to both the availability of nitrate and

to temperature. The connected wetlands showed

highest denitrification rates due to regular nitrate

recharge by water bodies. The potential for increased

denitrification in isolated wetlands was demonstrated

by our series of increasing nitrate additions. If these

environments were connected and received pulsing

floods of river water, they should be able to remove

high amounts of nitrate from the water column and

serve an important role in nutrient removal.

Denitrification efficiency in shallow eutrophic

wetland sediments: N sink or sources?

Hydrological connectivity of riverine wetlands influ-

ences nitrogen removal efficiency in more ways than

simply through increasing nitrate concentration.

Denitrification efficiency is evaluated as the ratio

between denitrification rates and inorganic nitrogen

effluxes across the sediment–water interface (Dtot/

(Dtot ? DIN)). It represents the percentage of the

total processed inorganic nitrogen during organic

matter decomposition released as N2 (Eyre and

Ferguson 2002). According to Eyre and Ferguson

(2002), efficiencies close to 1 mean that net loss of

nitrogen is in gaseous form from the system, while

values close to 0 mean that inorganic nitrogen ions

are recycled to the water column and may sustain

new primary production. Calculated denitrification

efficiencies for the 10 connected and 12 isolated

wetlands studied in this work are reported in Fig. 7.

In both seasons, denitrification efficiency in isolated

wetlands was below 0.2, suggesting that these

environments tended to be sources of inorganic

nitrogen with ammonification (or dissimilative reduc-

tion of nitrate to ammonium) prevailing over deni-

trification (Gardner and McCarthy 2009). Low

denitrification efficiency at isolated wetlands means

that these sites tended to maintain eutrophic–hype-

reutrophic conditions with cycling between plank-

tonic or macrophytic primary production, organic

matter settling, fast regeneration of the labile fraction

and burial of the recalcitrant litter, resulting in rapid

infilling. Benthic nitrogen cycling in these environ-

ments appears to be simplified due to negligible

nitrification/denitrification processes and limited

availability of oxidized inorganic nitrogen forms.

Connected wetlands tended to be sources of DIN to

the water column, with ratios below 0.5 in winter,

while in summer they tended to be nitrogen sinks

(denitrification efficiency median 0.65, range 0.4–

0.8). Hydraulic connectivity likely maintains oxi-

dized conditions in the water and in the upper

sediment layers and low ratios of sedimentary organic

carbon to nitrate availability, enhancing permanent N

removal.

The fraction of total NO3 -

uptake undergoing

denitrification was calculated dividing N2 production

(from DW) by NO3 -

fluxes (only negative values). At

connected sites the ratio was 0.5 and 1.0 in winter and

350 Biogeochemistry (2011) 103:335–354

123

summer, respectively (median values). These num-

bers suggest that in winter only half of the nitrate

uptake was due to denitrification while in summer

most of the nitrate was removed via this process. At

isolated sites the calculated Dw to NO3 -

flux ratio

was 0.6 and 0.5 in winter and summer, respectively

(median values), suggesting that at both sampling

seasons other processes besides denitrification were

responsible for an important fraction of NO3 -

uptake.

Dissimilative nitrate reduction to ammonium, dark

incorporation in microphytobenthos and bacteria or

other microbially mediated or chemical processes

could be relevant processes in the studied wetlands.

Denitrification rates in freshwater wetlands could

be addressed in a mass balance context to assess the

relevance of the process at a basin scale. We

calculated the N budget according to the ‘‘Soil

System Budget’’ (Oenema et al. 2003) for the lower

Oglio River basin, where most of the studied

wetlands are located. For this 3,700 km 2

watershed

we estimated an annual nitrogen surplus of 32,633

metric ton (t) N and we calculated, assuming highest

measured denitrification rates (1,260 kg N ha -1

year -1

) and hydraulic connection for all wetlands in

the basin, a potential N removal of only

250 t N year -1

. In fact even if maximum denitrifica-

tion rates on areal basis are about nine-folds higher

than average N surplus (143 kg N ha -1

AL year -1

;

AL = Agricultural Land), total riverine wetlands

area (*200 ha) is three orders of magnitude smaller than agricultural lands (*230,000 ha) within the catchment. According to Verhoeven et al. (2006),

wetland environments contribute significantly to

water quality improvement if they account for at

least 10% of the watershed area, while in the Oglio

basin and likely in the Po River Plain the potential of

wetlands to remove excess N is at present irrelevant

due to their overall negligible surface (�1%).

Benthic respiration and denitrification in shallow

eutrophic wetlands

This last section intends to (1) clarify the role of

denitrification in organic matter mineralization in

light of measured benthic metabolism and (2) quan-

tify the links among metabolic rates, connectivity and

N removal. The elevated rates of oxygen consump-

tion and the low oxygen reserve in the water column

measured in the shallow wetlands of the Po River

Plain are typical of environments that risk anoxia. We

calculated that in summer isolated wetlands can turn

hypoxic to anoxic in less than 10 h, mainly in the

night, without the offset of photosynthesis. The ratio

between SOD and DIC fluxes enabled us to evaluate

the importance of anaerobic metabolism. SOD,

corrected for oxygen demand due to nitrification,

was stoichiometrically comparable or significantly

below DIC fluxes. On average, about 77 and 59% of

measured DIC fluxes were uncoupled from corre-

sponding O2 uptake rates in summer and winter,

respectively. Imbalances in SOD and DIC effluxes

are typical of eutrophic to dystrophic aquatic envi-

ronments and are associated with accumulation of

reduced compounds in pore water (Ingvorsen and

Brock 1982; Capone and Kiene 1988). Denitrification

in isolated wetlands represented molar equivalents on

average 9 and 12% of SOD in winter and summer,

respectively. In connected environments, these per-

centages increased to 19 and 25%, and at some sites

the moles of N2 produced via denitrification were

winter

D to

t/ (D

to t+

D IN

)

0,0

0,2

0,4

0,6

0,8

1,0

connected

summer

D to

t/ (D

to t+

D IN

)

0,0

0,2

0,4

0,6

0,8

1,0

connectedisolated all sites isolated all sites

Fig. 7 Denitrification efficiencies calculated according to Eyre and Ferguson (2002) (see the text for major details)

Biogeochemistry (2011) 103:335–354 351

123

comparable to the moles of O2 respired. Oxygen

consumed via nitrification, calculated as the sum of

2DN ? 2NOx -

efflux, was a minor fraction of SOD.

It averaged 11 and 3% of SOD in winter and summer,

respectively, and represented only 2 and 14% of SOD

for isolated and connected sites. Oxygen consump-

tion was thus mainly due to aerobic carbon miner-

alization and to the reoxidation of anaerobic

metabolism end products.

We estimated carbon oxidation by denitrification

assuming 1.25 C:1 N (mol:mol) (Richards 1965) to

evaluate the contribution of denitrification to organic

matter mineralization. Total denitrification rates were

responsible for an extremely variable production of

inorganic carbon, varying between \5 and 450 lmol C m

-2 h

-1 in winter (average of pooled data

98 ± 23 lmol C m-2 h-1) and between 11 and 2,369 lmol C m-2 h-1 in summer (average of pooled data 355 ± 108 lmol C m-2 h-1). Assuming that DIC fluxes were a proxy for organic carbon mineral-

ization, we estimated that the fraction of carbon

oxidation due to denitrification ranged from \1 to 83%, with winter and summer averages of 7.7 and

13.0%, respectively. In connected sites, this fraction

was generally ecologically significant (10–83%),

while in isolated sites it was \5%. We thus demon- strated that in riverine connected wetlands denitrifica-

tion was significant (1) for the removal of nitrogen

from NO3 -

-polluted waters and (2) as a respiratory

process responsible for a significant fraction of carbon

mineralization. Similar results were reported for

estuarine and shallow marine environments by Yoon

and Benner (1992) and Laursen and Seitzinger (2002)

and for riparian freshwater wetlands (Seitzinger 1994).

Concluding remarks

This study demonstrates that connectivity of eutrophic

wetlands with rivers is a primary factor in controlling

denitrification and nitrogen cycling. At connected

sites, denitrification can be an important pathway for

nitrogen cycling and a significant component of

benthic respiration. Here, denitrification was mostly

sustained by nitrate diffusing to anoxic sediments

from the water column, which was primarily regu-

lated by nitrate availability and temperature. Mea-

sured values were within those reported in the

literature and, for a wide range of rates (from \5 to

140 lmol N m-2 h-1), can be generally predicted by a simple diffusion–reaction model. Denitrification

coupled to nitrification was generally low, probably

due to limited oxygen availability within sediments.

At hydrologically isolated sites, denitrification was

a negligible process with ammonium regeneration

largely prevailing over nitrate removal. This can

enhance primary productivity, infilling and, perhaps

ultimately, the shift of these temporary environments

towards more terrestrial conditions, as export of

newly formed organic matter is very low. Therefore,

the imbalance of nitrogen pathways could also

generate feedbacks with a detrimental loss of riparian

aquatic habitats which act as regulators of the whole

river ecosystem.

Pulses of nitrate immediately stimulated denitrifi-

cation meaning that all the studied environments had

a potential capacity for greater nitrogen removal. This

feature has to be taken into account as a key element

in the management of the hydrographic network of

the Po River Plain, and similar impacted rivers,

where nitrate contamination of surface and ground

waters is a critical issue. There is evidence from this

study, that actions for rehabilitating the lateral

connectivity between rivers and riverine aquatic

habitats could greatly improve nitrogen removal via

denitrification with beneficial effects on water quality

and persistence of habitats themselves.

Acknowledgements This study was supported by the Fondazione Lombardia per l’Ambiente, by the Fili d’Acqua

Project, funded by Oglio Sud Natural Park, by the US National

Science Foundation grant DEB-0621014, through the Virginia

Coast Reserve Long-term Ecological Research program and by

the East Carolina University, Biology Department.

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