Factors controlling denitrification rates of tidal mudflats and fringing salt marshes in south-west England

Estuarine, Coastal and Shelf Science (1992) 34, 471-485

Factors C o n t r o l l i n g Denitrification Rates o f Tidal Mudflats and F r i n g i n g Salt Marshes in S o u t h - w e s t E n g l a n d

M . S. K o c h a, E. M a l t b y be, G. A. O l i v e r b a n d S. A. B a k k e r d aLaboratory for Wetland Soils and Sediments, Center for Wetland Resources, Louisiana State University, Baton Rouge, LA 70803, U.S.A., bDepartment of Geography, University of Exeter, Exeter, England, EX4 4Rff , U.K. and aDepartment of Plant Ecology, University of Utrecht, Lange Nieuwstraat 106, 3512 P N Utrecht, The Netherlands Received 29 October 1990 and in revisedform 20 November 1991

Keywords:

denitrification; oxidation-reduction; marsh; mudflat; England

Denitrification rates were determined utilizing the acetylene blockage technique sites: upper mudflat, lower mudflat, and Halimione portulacoides marsh on the fringing wetlands of the Torridge River Estuary in South-west England. Denitrification rates were calculated from nitrous oxide (N20) production each month for 1 year with intact sediment cores extracted at low tide (0-5 cm). In the lower a0.d upper mud_flat sites denitrification rates were low ranging from 0"52 to 5'78 ~tmol and 1'28 to 4'36 pmol N 2 m -2 h -~, respectively. Denitrification rates in marsh sediments were consistently higher than those of the mudflat ranging from 2.51 to 59.00 Ilmol N 2 m -2 h -~. Amending river water to sediment cores stimulated lower and upper mudflat denitrification rates approximately 10-fold up to 106.39 and 96-73 ~tmol N 2m -2 h - 1 respectively. In marsh sediments, a twofold increase in denitrification was found with river water amended resulting in a maximum rate of 114.80 ~mol N 2 m -2 h -~. During the winter months, when riverine N O 3 - N levels were at a maximum (2-47 to 2.93 mg 1-~), denitrification rates were highest (75-24 to 114-99 I~mol N 2 m -2 h -1) and conversely, during the summer both N O 3 - N concentrations (1.0 to 1.70 mg 1-~) and denitrification (0"95 to 37.38 pmol N 2m - 2h - ~)rates were at a minimum. Mudflat sediment redox potentials (Eh), within the theoretical range of N O 3- instability, were limited to the upper 5 turn, thus maximum denitrifieation rates may be restricted to the sediment surface. When calculating annual denitrification rates in tidal estuaries several factors should be considered including: seasonal N O 3- concentrations in tidal water, tidal flooding duration and amplitude, and the depth of the aerobic/anaerobic zone of the sediment.

at three

Introduction R e c e n t a t t e n t i o n has b e e n g i v e n to p o l l u t i o n p r o b l e m s associated w i t h h i g h i n o r g a n i c n i t r o g e n l o a d i n g to a q u a t i c systems. I n B r i t a i n , m e a n a n n u a l r i v e r i n e N O 3- c o n c e n t r a t i o n s h a v e i n c r e a s e d m o r e t h a n 70~/o since 1955 a n d h a v e b e e n f o u n d to r e a c h c o n c e n t r a t i o n s 'Author to whom correspondence should be sent. 0272-7714/92/050471 + 15 $03.00/0

© 1992 Academic Press Limited

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M . S . Koch et al.

over I0 nag N 1-1 (Marsh, 1980). In addition to riverine nutrient sources, coastal estuaries are frequently used in Britain to discharge treated municipal sewage wastes containing high levels of N H 3 which readily oxidizes to N O 3- and NO2-. Mudflat and marsh sediments within estuarine ecosystems have been found to denitrify oxidized nitrogenous compounds to N z due to the anaerobic sediment environment of saturated soils (Kaplan et aL, 1977; Patrick & Delaune, 1977; Oremland et aL, 1984; Seitzinger & Nixon, 1985; Seitzinger, 1988). This N z may evolve from the sediment to the atmosphere by diffusion through the mud or convected via plant airspace tissue (Reddy et aL, 1989) and animal burrows, is thus lost from the system. T h e contribution of denitrification in the N budget of U.K. salt marshes may be limited (Abd. Aziz & Nedwell, 1986) either by nitrate availability or the complete reduction of N O 3- to NH4 + (Sorensen, 1978a). However, denitrification rates in coastal wetlands have been reported to reach 1095 kg-N ha- i year-1 (Kaplan et al., 1977) with estimates more frequently in the range of 5 to 60 kg-N h a - i year- 1 (Kaspar, 1982). Due to the high NO~- levels in many British rivers and estuaries, often in excess of 10 nag N 1-1 (Marsh, 1980), fringing wetlands of coastal rivers could process a significant amount of riverine N O 3- when flooding occurs in the upper reaches of the river or when tidal waters flood low lying areas. T h e objective of the present research was to investigate the potential for denitrification in riverine fringing wetlands bordering a river with high NO3- concentrations. Seasonal denitrification rates of sediments from both a mudflat and the bordering vegetated salt marsh were investigated. In addition, several factors which may influence denitrification rates were examined including nitrate availability, temperature, and sediment depth. Changes in the mudflat sediment redox profile were also identified through several tidal cycles as an indicator o f the sediment denitrification zone.

Materials and m e t h o d s Area of study A study area was chosen in north Devon on the banks of the River Torridge (Figure 1). T h e River Torridge drains an agricultural catchment and is part of an estuary formed where the Rivers Torridge and T a w converge. Tidal ranges of up to 6-0 m facilitate the transport o f fine organic clays which flocculate out of suspension by ionic interchange creating fringing mudflats and tidal marshes. T h e study area includes a fringing salt marsh which extends east to west approximately 60 m to the bare mudflat with tidal channels that drain the marsh after inundation by spring tides (Figure 1). During a neap tide, the marsh and upper mudflat remain drained, however, the lower mud_flat is inundated with each tidal cycle. T h r e e specific sites were selected for study: one marsh site located 20 m east from the edge of the marsh ( M H W M ) within a monotypic stand ofHalimione portulacoides, the dominant salt marsh macrophyte in the marsh and two mudflat sites, an upper and lower zone extending 20 and 40 metres west of the marsh-mudflat transition zone, respectively (Figure 1). Sediment characteristics Physical characteristics of the marsh, upper and lower mudflat, and the surface microlayer between the two mudflat sites (intermediate mudflat) are presented in Table 1. Clay and silt comprise the dominant size fraction in the marsh and upper mudflat sediment (Micromeritics sedigraph Model 5000D; Norcross, Georgia). A higher proportion of sand

Factors controlling denitrification

473

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Figure 1. Study area on the eastern bank of the River Torridge showing the three sampling sites: lower mudflat, upper mudflat, and Halimione portulacoides marsh. T h e location of the tower housing the tide gauge and data logger on the intermediate mudflat site is shown by + . T h e mean high water mark is illustrated ( M H W M ) and the three monthly water quality sampling stations of South West Water Authority (SWW).

size fractions characterize the upper mudflat site compared to the marsh. Sediment at the lower mudflat site is approximately 50% sand with a lower silt content, but still possess a high percentage of smaller clay size particles. Total carbon levels are similar at all three sites, but nitrogen content decreases from the marsh to the lower mudflat (Carlo Erba C:N Analyzer Model 1400; Carlo Erba, Milan). Easily oxidizable organic matter (Walkey & Black, 1934) shows a similar gradient as % N , however is greatest in the surface micro-layer of the mudflat potentially due to the presence of benthic diatoms (pers. obs.).

Monthly denitrification rates Each month for one year, beginning in October 1988, denitrification rates were determined using the acetylene inhibition technique (Sorensen, 1978b). This method measures the reduction of N O 3- and N O 2- by blocking the reduction of N 2 0 to N 2 with acetylene (C2H2) and determining the N 2 0 concentration in the head space. Twelve replicate cores to a depth of 5.0 cm (diameter 2-0 cm) were collected from each sample site every

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474

TABLE1. Physical characteristics and N and C content of sediments with depth from the marsh, upper mudflat, lower mudflat, and microlayer (0-3 ram) of the intermediate mud.flat site. Means given with_+ standard error (n = 3)

Site Marsh

UMF

LMF

IMF

Depth (on)

C (%)

N (%)

EOOM (%)

0--5 5--10 10-15 15-20 0-5 5-10 10-15 15-20 0-5 5-10 10-15 15-20 surface

5.24+0.11 4.58-+0.02 4.34-t-0-10 4"25+0'01 4.28_+0.09 4.42-+0.04 4.09-+0.12 4.36-1-0.08 4"48-+0-21 4-20-t-0-19 4.16-t-0.15 4"45_+0'10 4"75_+0.08

0-30+0.042 0-29+0.009 0.28-t-0.000 0.25-t-0.009 0-22-t-0.010 0-21-t-0.010 0.19+_0.010 0.20_+0.050 0-10-+0-009 0-11-+0-003 0-09-+0.003 0-09+0,006 0.26_+0-005

5-25-1-0.83 5-03+0.15 3-85-t-0.42 2.78+0.43 5-59_+0-77 4-60+0-10 4-30_+0.21 3.98-+0.24 3"54--+0-33 3-85-+0-33 3-34-1-0-12 3-47_+0-25 7.51_+0.10

Clay (%)

Silt (%)

Sand (%)

48"14+0-51 48-39_+0-86 3.47_+1-14 48-95-+1.25 48-47-+0.92 2.55-+0-6t 48.19___1.41 50-29___1.79 1.52-t-0.39 49.66-t-0-84 48-85+0.76 1.49+0.32 38.38_+1-07 52-56+0.95 9-05-+0.12 39-50-+1-53 52-48-1-0.91 8-03+0-78 35.53-1-0.66 51-18-t-0.89 13.28_+0-81 35.85_+0.41 47-88-+0.70 16-27+0-34 22.59-t-2.98 29-11__+2"05 48-30--+0"94 27-13--+0-895 28-30-+1"13 44-57+1-85 25'18_+1"68 25"06___0"82 49"77-1-2-42 23'27_+_0"77 20"47___0"74 56"25_+1-50 35.83-+0.78 48.93__,0.86 15-23_+1-63

EOOM = H202 removed organic matter.

m o n t h . T h e cores w e r e t a k e n at low tide w i t h a plastic c o r e r a n d e x t r u d e d into 60 m l i n c u b a t i o n jars fitted w i t h O - r i n g s a n d t r a n s p o r t e d i m m e d i a t e l y to t h e l a b o r a t o r y o n ice ( 2 - 5 °C). T w o different t r e a t m e n t s were a p p l i e d to each set o f six r e p l i c a t e cores t a k e n f r o m t h e t h r e e sites. T r e a t m e n t 1 c o n s i s t e d o f s e d i m e n t cores m a i n t a i n e d in an intact c o n d i t i o n as t a k e n f r o m t h e sites at l o w tide. T r e a t m e n t I cores w h i c h m a i n t a i n e d a vertical o r i e n t a t i o n in t h e i n c u b a t i o n jars were i n j e c t e d w i t h 2.0 m l o f C2H 2 s a t u r a t e d d i s t i l l e d H 2 0 (1-6 cc C2H2: I m l H 2 0 ) u s i n g an 8-0 c m l e n g t h s y r i n g e i n s e r t e d t h r o u g h a s e r u m s t o p p e r . S l o w l y injecting t h e w a t e r as t h e s y r i n g e was p u l l e d o u t o f t h e i n t a c t core was f o u n d to b e t h e b e s t m e t h o d for d i s t r i b u t i n g t h e C2H 2 s a t u r a t e d H 2 0 t h r o u g h the s e d i m e n t c o l u m n . T h e h e a d space was also i n j e c t e d w i t h 8 . 0 % vol:vol ratio o f C2H 2. T r e a t m e n t 2 cores w e r e a m e n d e d w i t h 30 m l o f T o r r i d g e r i v e r w a t e r (collected o n t h e s a m e day) w h i c h was s a t u r a t e d w i t h acetylene. A n a d d i t i o n a l 8.0°/0 o f C2H 2 was a d d e d to the h e a d space. T r e a t m e n t 2 vessels were s h a k e n on a m e c h a n i c a l shaker for 10 m i n . b e f o r e i n c u b a t i o n . All cores, i r r e s p e c t i v e o f t r e a t m e n t , w e r e i n c u b a t e d at a m b i e n t s e d i m e n t t e m p e r a t u r e s (seasonal r a n g e = 7-3 to 19"0 °C). A f t e r a 24 h i n c u b a t i o n , a 5 m l a l i q u o t o f h e a d space gas was i n j e c t e d into a h e a t e d e l e c t r o n c a p t u r e gas c h r o m a t o g r a p h (Pye Series 104) a n d N 2 0 c o n c e n t r a t i o n s d e t e r m i n e d . R e s u l t s w e r e s t a n d a r d i z e d to a 100 p p m (vol:vol) N 2 0 s t a n d a r d ( A l l t e c h Assoc., U . S . ) at a 5 p p m l i m i t o f detection. T r e a t m e n t 2 cores were again s h a k e n to e q u i l i b r a t e t h e N 2 0 s a t u r a t e d in the s e d i m e n t s o l u t i o n a n d t h e h e a d space b e f o r e t a k i n g t h e gas sample. T h e p e r c e n t w a t e r in each i n c u b a t i o n jar was d e t e r m i n e d to calculate the d i s s o l v e d N 2 0 u s i n g s o l u b i l i t y coefficients a n d H e n r y ' s law a c c o r d i n g to M o r a g h a n & B u r e s h (1977).

Denitrification experiments Nitrate vs. temperature. A factorial e x p e r i m e n t was c o n d u c t e d (2 x 4 x 4) w h i c h i n c l u d e d site, n i t r a t e level, a n d t e m p e r a t u r e to i n v e s t i g a t e t h e influence o f t h e s e v a r i a b l e s o n p o t e n tial denitrification. S e d i m e n t cores were e x t r a c t e d as d e s c r i b e d a b o v e f r o m one site o n t h e

Factors controlling denitrification

475

mudflat between the u p p e r and lower mudflat zone (30 m west of the marsh) and the H. portulacoides marsh site (Figure 1). Saturated C2H 2 distilled water (30 ml) plus N O 3with the following concentrations: 0.0, 5.0, 10-0 and 15.0 mg N 1- l (ranges based on values reported for British rivers and estuaries; Marsh, 1980) was added to replicated cores (n = 4). Cores were incubated for 24 h at 0, I0, 20 and 30 °C. T h e acetylene technique and gas analysis were consistent with those for the monthly sampling for T r e a t m e n t 2.

Depth profile. T o determine the differences of potential denitrification rates within the sediment profile, incubations were conducted on three integrated depths, 0-5, 5-10, and 10-15 cm. T h e extractions, preparations, and analysis were consistent with those described for the monthly sampling including treatments 1 and 2. Runs were conducted on sediment from the marsh and the intermediate mudflat zone between the upper and lower mudflat site (n = 4). T r e a t m e n t s 1 and 2, as described above, were applied to cores (n = 4) from the two sites. Tidal surveys Redox measurements. Eh (mV) measurements have been used by Patrick and Delaune (1977), Armstrong et al. (1985), and others to determine the reducing condition of wetland soils under low oxygen tensions. Although natural systems possess mixed redox couples (Bohn 1971), these couples are reduced in a sequential thermodynamic sequence (Gambrell & Patrick, 1978), thus Eh can be used as an indicator of the degree of soil reduction. In this study, redox potentials (Eh) were measured (mV) for a 72 h period during spring (21-24 April 1989) and neap (12-15 M a y 1989) tidal cycles in the upper soil profile. Eh values were corrected by subtracting the potential of the calomel reference electrode, - 244 mV, from the m V readings. N o Eh corrections for p H were made due to the mixed redox potentials measured. Also, unlike Eh determinations, p H was measured only at one depth~ 30 ram. Redoxprobe construction and installation. A waterproof housing for the redox probes and a data-logging system was constructed to monitor Eh changes (mV) over the tidal cycles. T h e waterproof housing was machined from 4" P V C pipe to approximately 30 cm in length and fitted throughout with O-rings. This casing contained holes in the bottom for thirty constructed platinum-tipped (20 gauge, 0-8 m m diameter) glass redox probes (1 cm diameter x 20 cm length), a calomel reference probe (1 cm diameter x 15 cm length), and a p H probe (1 cm diameter x 15 cm length). A 25 m multi-core cable connected the probe to a portable data-logging system (D.M.P. Electronics, U.K.). Redox probes were rinsed with dilute HC1 and distilled water and calibrated with quinhydrone in p H buffer 4-0 ( + 2 1 8 mV) before being installed in the field. T h e p H probe was also calibrated and adjusted to p H buffer 7-0 and 4.0. T h e 30 redox probes were carefully aligned with precision calipers to measure 10 depths ( n = 3) on a micro-scale: surface (0), 1, 2, 3, 4, 5, 10, 20, 30 and 60 ram. T h e p r o b e was inserted slowly into the mudflat with a welded stake and positioned such that the platimtrn tips of the 3 surface (0 m m ) redox probes were no longer visible. T h e data-logger was positioned on a scaffolding tower 5-0 m above the surface of the sediment on the mudflat between the u p p e r and lower mudflat sites (Figure 1). T h e datalogger was initiated in the field with a portable microcomputer (Amstrad, U . K . ) and p r o g r a m m e d to read all 30 probes sequentially within a 3 min. cycle every 20 rain.

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M.S. Koch et al.

Water chemistry. Changes in the interstitial soil and surface water N O 3 - + NO 2- and NH4 + concentrations along with conductivity were monitored every hour through a tidal cycle from 1500--2100h on 22 February 1989. Interstital water samples were extracted using porous ceramic cups (Debyle et al., 1988) set at 0--5 and 5-10 cm depths in the H. portulacoides marsh and the intermediate mudflat site. Surface water samples were taken from the river, the marsh creek, and mudflat creek. Conductivity measurements were made on a Whatman CDM100 temperature adjusted conductivity meter. Readings and water samples were taken every hour and samples put on ice immediately for transport back to the laboratory for analysis. All water samples were filtered (0-45 ~t) and run on a Technicon Autoanalyzer System using the continuous flow cadmium reduction method to determine NO 3- + NO2-. Ammonium was determined by spectrophotometry according to the Phenol/DIC method. Tidal water elevation. Duration and amplitude of tides on the mudflat were measured from the tower with a tide gauge and recorded with a punch chart recorder (Southwest Water Authority, SWW, U.K.). Statistical analysis Analysis of variance, regression, and correlations were performed using the SAS statistical package (SAS 1982; Cary NC, U.S.A.) and SPSS statistical package (SPSS 1985; Chicago IL, U.S.A.). Statistically significant differences in N20 production between months and depths were elucidated using Duncan's multiple range test (ANOVA). All means are statistically different at the P < 0.05 level of significance unless otherwise stated. Results

Monthly N 2 0 production Lower and upper mudflat intact sediment cores (treatment 1) had low N20 production values ranging from 0"011 to 0-122 ~tg N20 cm -3 day -1 and 0-027 to 0.092 ~tg N20 cm -3 day- l respectively (Figure 2). Monthly values of N20 production from intact marsh cores were significantly higher (P < 0.01) than either mudflat site for every month except May. No significant differences in N20 production levels were found between the lower and upper mudttat site under treatment 1 for any month of the year (Figure 2). Under treatment 2, higher sediment N20 production was found in the marsh compared to both mudflat sites every month except December, May and September (Figure 2). T h e two mudflat sites had similar N20 production rates throughout the year, with only slight but significant variations during November and January measurements (Figure 2). Monthly N20 production increased 10-fold (P < 0.01) under treatment 2 for both mudflat sites for every month except August (Figure 2). Marsh N20 values also increased with river water amended, but only two-fold. In the marsh, differences between the two treatments were significant for all months but the summer months, June through August and the month of October (Figure 2). No seasonal relationship was found for N20 production from intact sediment cores extracted under low tide conditions (treatment 1; Figure 2). However, seasonal trends were found in cores with ambient river water amended (treatment 2). Multiple range testing of N20 production means from each of the three sites under treatment 2 resulted in a grouping of months into three statistically distinct categories: (1) December, January, February, and March with the highest means ranging from 1.591 to 2.426 ~tg N20 cm -3

Factors controlling denitrification

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Figure 2. Seasonal changes in N~O production (l~gNzO cm-3 day-') of sediment cores extracted (0-5 cm) from three sites, lower mudflat, £x upper mudflat, O and marsh, • during each month October 1988--September 1989. Results from the two treatments are illustrated (a) Treatment 1, ambient low tide conditions, (b) Treatment 2, 30 mt of river water amended. Bars represent _ standard error of the mean (n = 6).

day-1 (2) April, September, October and November with intermediate means ranging from 1.079 to 1.618 ~tg N 2 0 cm -3 day - I and (3) May, June, July and August with the lowest means ranging from 0-0175 to 0.844 ~tg N 2 0 cm -3 d a y - i . T h u s for treatment 2, the highest N 2 0 production occurred during the winter and lowest during the summer months for all three sites [Figure 2(b)].

Nitrate and temperature effects Nitrate concentration had a greater influence on NzO produced in the sediments than changes in temperature over a 24 h incubation period (Figure 3). Mudflat and marsh sediment N 2 0 production was highly correlated to N O 3- concentration added (r = 0"90 and 0.85, respectively) in contrast to incubation temperature ( r = - 0 . 1 2 and - 0 . 1 7 , respectively). A logarithmic fit illustrates the relationship between N O 3- levels amended and N 2 0 production, explaining 81% of the mudflat and 85% of the marsh N 2 0 variability (Figure 3). An increase in the N O 3 - - N level above 10.0 mg 1-~ had little effect on N 2 0 produced, lending evidence that a maximum denitrification potential was attained.

478

114.S. Koch et al.

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N03-N mg I-t) Figure 3. Nitrous oxide produced (lag N20 cm 3 day-') in response to four nitrate amendments (0-0, 5.0~ 10.0 and 15-0 mg NO 3 1 *) and four incubation temperatures (II, 0; O, 10; [3, 20; + 30 °C) from sediment cores (0-5 cm depth; n = 2) extracted from the (a) intermediate mudflat and (b) marsh site.

Depth profile D e c r e a s e d N 2 0 p r o d u c t i o n w i t h d e p t h was o n l y significantly different in t h e m a r s h s e d i m e n t cores u n d e r t r e a t m e n t 1 ( T a b l e 2). T r e a t m e n t 1 m a r s h s e d i m e n t s h a d an o r d e r o f m a g n i t u d e g r e a t e r N 2 0 p r o d u c t i o n at each d e p t h t h a n the m u d f l a t s e d i m e n t s ( P < 0.01). T h e a d d i t i o n o f river w a t e r to the m u d f l a t a n d m a r s h s e d i m e n t cores r e s u l t e d in signific a n t l y h i g h e r N 2 0 levels ( P < 0 . 0 1 ) c o m p a r e d to t r e a t m e n t 1 cores, at all t h r e e d e p t h s ( T a b l e 2).

Tidal surveys Redox and pH. D u r i n g t h e s p r i n g tidal s u r v e y , at t h e s e d i m e n t surface, 0 r a m , a n d t h e 2 m m d e p t h , r e d o x p o t e n t i a l s w e r e s u s t a i n e d w i t h i n t h e t h e o r e t i c a l d e n i t r i f i c a t i o n zone, a p p r o x i m a t e l y + 2 0 0 to + 2 5 0 m V at p H 7.0 (Patrick & M a h a p a t r a , 1968; B o h n , 1971; P a t r i c k & D e L a u n e , 1977; F i g u r e 4 I(a). T h e 1 m m d e p t h was o x i d i z e d ( T u r n e r & P a t r i c k ,

Factors controlling denitrification

479

TABLE 2. Nitrous oxide production with depth in mudflat and marsh soil cores under

treatment 1= in sire low tide soil conditions and treatment 2 = 30 ml of river water amended. Means given with-+standard error (n = 4) Mud.flat

Site:

Treatment: Depth (cm) 0-5 5-10 10-15

1

Marsh 2

1

2

N20 (lagcm-3 day-~) 0~039_+0.006 0-980_+0.309 0.976_+0.071 1.469_+0.0094 0.022_+0.0091-263+0-054 0.423_+0.098 1-549+_0-048 0"039___0-0231-524_+0-144 0-377-+0-038 1-457_+0-057

1968) throughout the 72 h survey, without a clear response to the changes in tidal cycles. At the 3 m m depth, tidal water seemed to be buffering the Eh compared to the underlying anoxic sediments [Figure 4 I(b)]. Below the 3 m m depth, Eh values failed to reach the theoretical zone of denitrification. T h e effect of the tide on Eh changes at 10 to 60 m m depth was dampened. Below the 10,0 m m depth in the sediment profile, Eh values were more indicative of sulphate reducing conditions [Connell & Patrick, 1968 Figure 4 I(c)], which develop only after all the available N O 3- has been reduced. During the neap tidal survey, water did not reach the intermediate mudflat zone where the tide gauge was located [Figure 4 II(d)]. Redox potentials at the surface (0 ram) maintained oxidized Eh values for approximately the entire 72 h period [Figure 4 II(a)]. T h e Eh values for the i m m and 2 m m depths were very different from the spring tidal pattern [Figure 4 I,II(a)]. Values were negative for 24 h and then rose to oxygenated levels for the last 48 h [Figure 4 II(a)], A similar upward trend was found in Eh at 3-5 m m depths and a diurnal pattern emerged after 24 h [Figure 4 II(b)]. A noticeable rise at the 10 ram depth did occur at 55 h. No change in Eh was found below the 10 m m depth consistent with the results from the spring tidal survey [Figure 4 II(c)]. Values for p H ranged from 7.22 to 7-48 p H units through the entire 72 h period in both the spring and neap tidal measurements (data not shown). Since the p H probe was located at a depth of approximately 30 m m in the sediment profile, these readings did not represent micro-scale changes in p H on the sediment surface. Chemical changes. During the tidal survey monitoring chemical changes in the water, an d conductivity of the river followed closely with the tide height, thus denser marine waters were mixing with river water on the surface [Figure 5(c)]. Riverine N O 3- + N O 2concentrations were not lowered with incoming tidal waters indicating an estuarine source. Levels of N O 3- ÷ N O 2- in the marsh creek increased six times when tidal waters entered the channel [Figure 5(a)] and declined as the marsh drained. Interstitial N O 3- + N O 2- concentrations were low at the 0-5 cm and 5-10 cm depths in the mudflat [Figure 5(b)]. L o w N O 3- + N O 2- values were also recorded in the marsh at the 5-10 cm soil profile. In the marsh surface sediments (0-5 cm), however, N O 3- + N O 2- concentrations were consistently 10-fold higher than in the other three sites. A m m o n i u m concentrations in the river, mudflat, and marsh creek were less than 0.15 mg N 1-1 throughout the tidal cycle, although the mudflat creek had higher initial concentrations of 0-25 m g 1-1. Ten-fold higher NH4 + levels were found in the mudflat sediments compared to the marsh. Levels of ammonium in the interstitial water of the marsh and mudflat at 5-10 cm did not change through the tidal cycle. However, at 0-5 cm

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depth ammonium concentrations in the marsh decreased from 0-63 to O-15 mg 1-~ and in the mudflat sediments increased from 1.02 to 1-39 mg 1-l from 1500-2100h. Discussion

N 2 0 production Seasonal denitrification rates on the mudflat and marsh of the River Torridge estuary were highest during the winter and early spring (Figure 2) when lower temperatures would potentially suppress microbial mediated NO 3- reduction. Denitrification kinetics can be affected by incubation temperature (Denmead et al., 1979; Kaplan et al., 1979; Goodroad & Keeney, 1984), however, NO 3- availability seemed to be the limiting factor for N20 production on the Torridge mudflat and marsh in this study. Maximum denitrification

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(sww). nitrate levels from December to February and reach a minimum during the later summer and autumn (Slack, 1977). These winter peaks are related to the replenishment of soil moisture deficits by heavy rainfall in winter and the subsequent leaching and mobilization of nutrients that have accumulated in the summer (Walling & Foster, 1978). Mudflat interstitial N O 3- + N O 2- levels in the sediments at low tide were less than 0.05 mg 1-1 N O 3_ N at the upper and lower mudflat sites (0-5 cm depths) during the tidal survey [Figure 5(b)]. An increase in sediment N O 3- + N O 2- concentrations found at high tide in the mudflat [Figure 5(b)] could influence denitrification rates when nitrate is limiting. Salt marsh interstitial N O 3- + N O 2- concentrations were consistently more than an order of magnitude higher than in the mudflat. This difference in N O 3- availability may have accounted for the 10-fold higher rate of N 2 0 production found in the marsh compared to mudflat sediments (Figure 2). Intertidal mudflat sediments in the T a m a r Estuary, U.K. have also been reported to have low N O 3- concentrations, 0.14 mg 1-1 N O 3 - N (Watson et al., 1985). Denitrification rates calculated from N 2 0 production with intact sediment cores on the lower and upper Torridge mudflat ranged from 0.52 to 5.78 ~mol N 2 m -2 h - ~and 1"28 to 4-36 N 2 m -2 h -1 (0-5 cm). These rates are similar to those found in mudflats of San Francisco Bay (U.S.A.), 0.80 to 1.2 IJmol N 2 m - 2 h-~ from intact sediment cores (0-3 cm) with low N O 3- - N concentrations, 0.31 to 0.52 mg 1- i (Oremland et al., 1984). However, with the addition of 25 ml of Bay water containing 1-0 m M N a N O 3 rates increased to 101 to 135 I~mol N 2 m -2 h -l. Denitrification rates of intact sediment cores from the Torridge fringing marsh were consistently higher than those reported for mudflats ranging from 2-51 to 59-00 lamol N 2 m -2 h -~. W h e n river water was amended in treatment 2 denitrification rates were stimulated in the lower and upper mudflat sediments up to 106.39 and 96-731amol N 2 m -2 h - l , respectively. A slightly higher denitrification rate was also measured in marsh cores with a maximum of 114.87 ~mol N 2 m -2 h - 1 Denitrification rates from sediment cores at all three sites on the Torridge with river water amended are similar to those calculated from sediment slurries by Seitzinger et al. (1980) in Narragansset Bay, 50 I~mol N 2 m -2 h -1, Nishio et al. (1981) in the T a m a Estuary, Japan,

Factors controlling denitrification

483

68 lamol N 2 m -2 h -l and Oremland et al. (1984) in San Francisco Bay, 17-280 ~tmol N 2 m-2h-1. Redox potential and p H Redox potentials within the theoretical zone of denitrification (Patrick & Mahapatra, 1968; Bohn, 1971; Patrick & DeLaune, 1977) on the intermediate mudflat site were restricted to the upper 3 m m of the sediment profile during spring tide and the upper 5 mm during neap tide (Figure 4). Combining 02 and N20 microsensors, Christensen et al. (1989) measured 02 depletion at 1.0 m m depth and measured a denitrification micro-zone of approximately 4.0 mm in a wetland stream sediment. These results are consistent with the characteristic micro-aerobic surface layer underlain by a thicker anaerobic zone found on the Torridge mudflat during both spring and neap tides (Figure 4). This type of sediment profile is prevalent in saturated, highly organic sediments with high microbial respiratory rates which consume oxidized compounds as readily as they diffuse into the anaerobic layer (Gambrell & Patrick, 1978). Below the zone where Eh values fluctuated on the Torridge River mudflat, 0-10 ram, sediment redox potentials approached the range of • Eh where sulfate becomes unstable and is subsequently reduced to sulfide [Cormell & Patrick, 1968 Figure 4 I(c)]. These low redox values were sustained for 72 h during several neap and spring tidal cycles. During the spring tides, Eh values consistent with the presence of 02 and N O 3- were found in the upper 2 m m of the mudflat sediment (Turner & Patrick, 1968; Bohn 1971; Gambrell & Patrick, 1978; Figure 4 I(a)]. Just below the 2 m m depth a transition zone occurred, characterized by a sediment profile more indicative of Eh values where oxidized forms of manganese (Mn 4+ ~ M n 2+, ~200 mV) and iron (Fe 3+ ---~Fe 2+, ~ 120 mV) become unstable assuming approximately neutral pH ( T u r n e r & Patrick, 1968; Bohn, 1971; Gotoh & Patrick, 1972)..With each incoming high spring tide, N O 3- diffusion may have facilitated a change in the redox status [Figure 4 I(b)] to levels reflecting N O 3 - - o N 2 theoretical redox coupling (Bohn, 1971). Once the tide ebbed, redox potentials began to fall presumably the result of a sequential reduction of oxidized compounds (Turner & Patrick, 1968; Figure 4 I(b). No comparable Eh measurements through a tidal cycle have been attempted from a tidal mudflat. However, a similar investigation was conducted by Armstrong et al. (1985) on changes in redox potentials of a British salt marsh. Measurements were taken from June to September sampling 32 times through periods of spring and neap tides. Armstrong et al. (1985) also found tidal redox responses to be limited to the surface layers. Below 50 mm depth the sediments remained intensely reduced (~ --200 mV) in a low Spartina marsh, the site most comparable to the mudflat site on the Torridge. Even when tidal waters did not reach the mudflat during extreme neap tides, fluctuations ofredox potentials occurred over a 72 h period [Figure 4 II(a,b,c)]. At neap tide the sediment surface remained oxidized, but at 1 and 2 m m depths oxidation occurred only after 30 h, perhaps as a result of the sediments drying out and facilitating oxygen diffusion [Figure 4 II(a)]. Below 2 m m a diurnal response was measured where Eh values declined during the day and increased at night possibly in accordance with temperature changes [Figure 4 ! I(b)]. Cooler temperatures may have slowed microbial respiration and increased the solubility of oxygen, thereby increasing the redox potentials at these depths. A general upward trend in redox values over time through the neap tidal cycles suggests a progressive drying out of sediments over the period of this survey. After prolonged periods of exposure the muds on the flats were noted to fissure which may be responsible

484

M . S . Koch et al.

for the sharp changes in reduced to more oxidized conditions over time at the 1 to 10 m m depths [Figure 4 I(a,b,c)]. N o attempt was made during these two tidal surveys to distinguish between changes in redox potentials in the u p p e r vs. lower mudflat zones. Also, Eh values of the H. portulacoides marsh sediments were not monitored through the tidal cycles in this study. However, Armstrong et al. (1985) measured Eh changes in a H. portulacoides creek bank and runnel site in the Welwick Marsh, Yorkshire (U.K.) through low spring tides, high spring tides, and very high spring tides. T h e H . portulacoides creek bank and runnel sites responded similarly with low Eh values occurring at the highest spring tides. But the redox potential of the creek bank sediments, being well drained, never dropped below 0 mV, whereas, in the runnel sites fell to - 1 0 0 m V at 0-30 cm for several days. High redox potentials in the H. portulacoides sites ( + 600 mV, low tide) compared to the T o r r i d g e mudflat, substantiate the idea that N O 3- would be more available in the marsh sediment and therefore denitrification would be an important component of the N-cycling in anaerobic microsites of soil aggregates (Greenwood & G o o d m a n , 1967).

Conclusion T h e high potential for denitrification by wetland sediments is well documented (Patrick & Mahapatra, 1968; Kaplan et al., 1977; Patrick & D e L a u n e , 1977; Oremland et al., 1984; Seitzinger & Nixon, 1985), however, the factors controlling annual rates vary dramatically depending on wetland type (Kaplan et al., 1979; Kaspar et al., 1982; Smith et al., 1985). Denitrification rates in tidal estuaries can be significantly influenced by several factors including, seasonal changes in N O 3- concentration in tidal waters, tidal flooding duration and amplitude, and the depth of the aerobic/anaerobic zone where N O 3- availability would potentially limit denitrification. T h e relative influence of these variables on denitrification rates should be investigated further and taken into consideration when calculating annual denitrification rates of mudflats and marshes of tidal estuaries.

Acknowledgements T h e current study would not have been possible without the support of the Sigma Xi Research Society National Chapter, Louisiana State University Student Exchange Program and D r William Patrick Jr of the Wetland Soils and Sediment Laboratory, L S U . T h e authors would like to acknowledge the South West Water Authority for funding this research and the University of Exeter Geography D e p a r t m e n t staffwho provided support for this reasearch and manuscript. Robert Twilley, I r v Mendelssohn, Ramesh Reddy and Mike Maceina are recognized for their assistance in reviewing the manuscript.

References Abd. Aziz, S. A. & Nedwell, D. B. 1986 The nitrogen cycle of an East Coast, U.K. Saltmarsh: II. Nitrogen fixation, nitrification, denitrification, tidal exchange. EstuaHne, Coastal and Shelf Science 22, 689-704. Armstrong, W., Wright, E. J., Lythe, S. & Gaynard T. J. 1985 Plant zonation and the effects of the springneap tidal cycle on soil aeration in a Humber salt marsh. Journal of Ecology 73, 323-339. Bohn, H. L., 1971 Redox potentials. Soil Science 112, 39-45. Christensen, P. B., Nielsen, L. P., Revsbech, N. P. & Sorensen, J. 1989 Microzonation of denitrification activity in stream sediments as studied with a combined oxygen and nitrous oxide microsensor. Applied and Environmental Microbiology 55, 1234-1241.

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Connell, W. E. & Patrick, W. H. Jr 1968 Sulfate reduction in soil: Effects of redox potential and pH. Science 159, 86-87. Debyle, N. V., Henries, R. W. & Hart, G. E. 1988 Evaluation of ceramic cups for determining soil solution chemistry. Soil Science 146, 30-36. Denmead, O. T., Freney, J. R. & Simpson, J. R. 1979 Studies of nitrous oxide emission from a grass sward. Soil Science Society America Journal 43, 726-728. Focht, D. D., 1974 T h e effect of temperature, pH, and aeration on the production of nitrous oxide and gaseous nitrogen--a zero order kinetic model. Soil Science I18, 173-179. Gambrell, R. P. & Patrick, W. H. Jr 1978 Chemical and microbial properties of anaerobic soils and sediments. In Plant life in anaerobic environments (Hook, D. D. & Crawford, R. M. M., eds). Ann Arbor Science, Michigan pp. 375-423. Goodroad, L. L. & Keeney, D. R. 1984 Nitrous oxide production in aerobic soils under varying pH, temperature and water content. Soil Biology Biochemistry 16, 39-43. Gotoh, S. & Patrick, W. H. Jr. 1972 Transformation of iron in a waterlogged soil as influenced by redox potential and pH. Soil Science Society America Proceedings 38, 66-71. Greenwood, D. J. & Goodman, D. 1967 Direct measurements of the distribution of oxygen in soil aggregates and in columns of fine soil crumbs. Journal of Soil Science 18, 182-196. Kaplan, W. A., Teal, J. M. & Valiela, I. 1977 Denitrification in salt marsh sediments: evidence for seasonal temperature selection among populations of denitrifiers. Microbial Ecology 3, 193-204. Kaplan, W., Valiela, I. & Teal, J. M. 1979 Denitrification in a salt marsh ecosystem. Limnology and Oceanography 24, 726-734. Kaspar, H. F. 1982 Denitrification in marine sediment: Measurement of capacity and estimate of in situ rate. Applied and Environmental Microbiology 43, 522-527. Marsh, T. J. 1980 Towards a nitrate balance for England and Wales. Water Services 84, 601-606. Moraghan, J. T. & Buresh, R. 1977 Correction for dissolved nitrous oxide in nitrogen studies. Soil Science Society Ameriea Journa141, 1201-1202. Nedwell, D. B. 1982 Exchange of nitrate, and the products of bacterial nitrate reduction, between seawater and sediment from a U.K. saltmarsh. Estuarine, Coastal and Shelf Science 14, 557-566. Nishio, T., Koike, I. & Hattori, A. 1981. N J A r and denitrification in Tama estuary sediments. GeomicrobiologyJournal 2, 193-206. Oremland, R. S., Umberger, C., Culbertson, C. W. & Smith, R. L. 1984 Denitrification in San Francisco Bay intertidal sediments. Applied and Environmental Microbiology 47, 1106-1112. Patrick, W. H. Jr. & Delaune, R. D. 1977 Chemical and biological redox systems affecting nutrient availability in the coastal wetlands. GeoscienceandMan 18, 131-137. Patrick, W. H. Jr. & Mahapatra~ I. C. 1968 Transformation and availability to rice of nitrogen and phosphorus in waterlogged soils. In Advances in Agronomy. Academic Press, New York. Reddy, K. R., Patrick, W. H. Jr & Lindau, C. W. 1989 Nitrification-denitrification at the plant root-sediment interface in wetlands. Limnology and Oceanography 34, 1004-1013. Seitzinger, S. P., Nixon, S. W., Pilson, M. E. & Burke, S. 1980 Denitrification and N20 production in nearshore marine sediments. Geochimicaet CosmochemicaActa 44, 1853-1860. Seitzinger, S. P. & Nixon, S. W. 1985 Eutrophication and the rate of denitrification and N20 production in coastal marine sediments. Limnology and Oceanography 30, 1332-1339. Seitzinger, S. P. 1988 Denitrification in freshwater and coastal marine ecosystems: Ecological and geochemical significance. Limnology and Oceanography 33, 702-724. Slack, J. G., 1977 Nitrate levels in Essex river waters. Institute Water Engineers and ScientistsJournal 31, 43. Smith, C. J., Wright, M. F. & Patrick, W. H. Jr. 1983 T h e effect of soil redox potential and pH on the reduction and production of nitrous oxide. Journal of Environmental Quality 12, 186-188. Smith, C. J., DeLaune, R. D. & Patrick, W. H. Jr 1985 Fate of riverine nitrate entering an estuary: 1. Denitrification and nitrogen burial. Estuaries 8, 15-21. Sorensen, J., 1978a Capacity for denitrification and reduction of nitrate to ammonia in a coastal marine sediment. Applied and Environmental Microbiology 35, 301-305. Sorensen, J. 1978b Denitrification rates in a marine sediment as measured by the acetylene inhibition technique. Applied and Environmental Microbiology 36, 139-143. Turner, F. T. & Patrick, W. H. Jr. 1968 Chemical changes in waterlogged soils as a result of oxygen depletion. Transactions 9th International Congress Soil Science (Adelaide, Australia) 4, 53-63. Walkley, A. & Black, I. A. 1934 An examination of the Degtjareffmethod for determining soil organic matter and a proposed modification of the chronic acid titration method. Soil Science 37, 29-38. Walling, D. E. & Foster, I. D. L. 1978 T h e 1976 drought and nitrate levels in the River Exe Basin. Institute Water Engineers and ScientistsJourna132, 341-352. Watson, P. G., Frickers, P. E. & Goodchild, C. M. 1985 Spatial and seasonal variations in the chemistry of sediment interstitial waters in the Tamar estuary. Estuarine, Coastal and Shelf Science 21, 105--119.