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Published online 1 March 2007
Published in J Environ Qual 36:540-548 (2007)
DOI: 10.2134/jeq2006.0352
© 2007 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
677 S. Segoe Rd., Madison, WI 53711 USA
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TECHNICAL REPORTS

Surface Water Quality

Sources of Sediment and Phosphorus in Stream Flow of a Highly Productive Dairy Farmed Catchment

R. W. McDowella,* and R. J. Wilcockb

a AgResearch Ltd, Invermay Agricultural Centre, Private Bag 50034, Mosgiel, New Zealand
b National Institute of Water and Atmospheric Research, P.O. Box 11 115, Hamilton, New Zealand

* Corresponding author (richard.mcdowell{at}agresearch.co.nz)

Received for publication September 3, 2006.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Both sediment and phosphorus (P) are important contaminants for surface water quality. Knowing the main sources of sediment and P loss within agricultural catchments enables mitigation practices to be better targeted. With this in mind seasonal loads of suspended sediment (SS), dissolved reactive P (DRP), total P (TP), and bioavailable P (BAP) were measured in a low gradient stream draining an intensively farmed New Zealand dairying catchment. Integrating in situ samplers were deployed to collect samples and the results merged with continuous flow data to calculate seasonal loads during 2005 through 2006. Flow rate, SS, and TP concentrations peaked in winter-spring and were lowest in summer-autumn. Concentrations of BAP in trapped sediment were greatest in autumn, contrasting with winter and spring when greater amounts of sediment were trapped, but with lower P enrichment. Analysis of 137Cs and mixing model output showed that a major source of sediment and associated P in winter and spring was stream banks. Possible causes for this include trampling and destabilization by stock, channel straightening and sediment removal, and removal of riparian trees that stabilize banks. Modelling indicated that overland flow probably from topsoil (but could include sediment from lanes) contributed most sediment during summer and autumn. Remediation aimed at decreasing particulate P inputs to streams should focus on riparian protection measures, such as permanent stock exclusion and planting with shrubs and trees, ensuring runoff from lanes is minimized, and decreasing Olsen P to nearer optimum agronomic levels.

Abbreviations: OC, organic carbon • DRP, dissolved reactive P • BAP, bioavailable P • LSD, least significant difference • msl, meters above sea level • PP, particulate P • TP, total P • SS, suspended sediment


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
THE mobilization and bioavailability of phosphorus (P) in fluvial systems is central to the eutrophication of P-limited lowland streams. Much of this P is transported as P bound to particulate material in suspended sediment (SS) (Brunet and Brian-Astin, 1998). Due to the application of P fertilizers and manures, or dung, P is most enriched in the topsoil and concentrations decrease with depth. In contrast, stream bed or bank sediments may be relatively P deficient (McDowell and Wilcock, 2004). Consequently, if flow and mobilization of particulate P (PP) occurs, it is important to know the origin of sediment as this will dictate the best management practice (BMP) to minimize P loss to surface waters. For instance, McDowell and Wilcock (2004) were able to isolate the source of SS in an agricultural catchment in New Zealand to topsoil derived from either overland flow or tile drainage. The BMP to decrease the load of P in the stream for this system was to decrease P loss by decreasing soil Olsen P concentrations, which were already in excess of optimum concentrations for plant production. However, if SS had been found to originate from subsoil, the stream bank or the stream bed then focusing on decreasing topsoil Olsen P concentrations back to the plant optimum would have little effect on PP concentrations and load in the stream.

There are several circumstances where topsoil may not be the primary source of PP in stream flow such as when stream banks erode into flow. These circumstances are enhanced by actions such as mechanical straightening of stream channels or dredging of ditches. This is often done to optimize water flow and prevent flooding, but also causes significant sediment disturbance and P release from remaining sediments (Smith et al., 2006). Similarly, if stock is allowed stream access then this can cause significant increases of dissolved P from dung and PP via bank disturbance (McDowell, 2006).

In addition to considering the source of PP for management of total P (TP) loads in a stream, bioavailability of P to periphyton is also an issue. Total P, of which PP is often the main fraction, is a poor predictor of P available to periphyton, which take up predominantly phosphate forms. Furthermore, P bioavailability varies with sediment source. McDowell and Sharpley (2001) found that although the mean total P concentration in bank sediments was greater than bed sediments from an agricultural catchment in Pennsylvania, bed sediments had more bioavailable P.

Phosphorus in soils and sediment is rendered bioavailable via desorption and mineralization (Ekholm, 1998; Reynolds and Davies, 2001). This process and the fraction of P bioavailable to algae are mimicked by Fe oxide strips or anion exchange membranes (Sharpley et al., 1995; Uusitalo and Ekholm, 2003). Data from North American and European studies have shown the bioavailable proportion of PP to account for 7 to 50% of TP depending on surrounding land use, the landscape, management, and season (Pionke and Kunishi, 1992; Sharpley et al., 1995; Steegen et al., 2001; Uusitalo et al., 2001). However, compared with many parts of North America or Europe, the New Zealand and Australian climate is very different, and as a consequence, management practices such as grazing during winter, are very different. This leads to different patterns of P loss during the year. For example, McDowell and Wilcock (2004) showed that the greatest load of TP lost from an agricultural catchment in New Zealand came during winter, at a time when fields are frozen in much of North America and parts of Europe and little P loss occurs (Xue et al., 1998).

Among the many landscapes in which dairying is practiced, production on allophanic soils tends to be high compared with other soils due to good nutrient reserves (Morton et al., 2003). However, allophanic (Udand) soils have very high sorption capacities and strengths and also tend to erode less than other soils (Hewitt, 1998; McDowell and Condron, 2004). Therefore, P loss from allophanic soils tends to be low. Despite this, significant loads of P loss have been noted (Betteridge et al., 2005). Consequently, our objectives were to obtain a representative time-integrated sample of SS in stream flow, determine its origin and bioavailability, and suggest the most effective form of managing this loss.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Study Area
The study area was the 2100 ha Waiokura catchment in Taranaki, New Zealand (Fig. 1). Land use within the catchment is pastoral dairy farming. Elevation ranges from about 380 msl (meters above sea level) in the headwaters to sea level at the catchment outflow on the Southern coast of Taranaki. Slope ranges from 0.5 to 20%, with only areas near the headwaters >5%. Mean annual rainfall (snowfall contributes <5% of total precipitation) in the catchment ranges from 2200 mm at the headwaters to 1200 mm at the outlet, while the mean annual temperature is 13.6°C. Soil within the catchment is a Manaia silt loam (New Zealand Soil Classification, Orthic Allophanic: USDA Taxonomy, Udand) with a mean Olsen bicarbonate-extractable P concentration (within dairy farms) of 66 mg P L–1 soil (soil bulk density is about 0.7–0.8 g cm–3). Two-thirds of the 44 dairy farms within the catchment spread effluent throughout the year onto defined blocks of land (about 10% of each farm), except in winter months when cows are not milked (dried off) and grazed off farm. The other third of farms use a two-pond system (an anaerobic pond and a facultative, or oxidation, pond) (Sukias et al., 2001) with effluent from the oxidation pond discharging into the stream. Superphosphate application rates range from 17 to 106 kg P ha–1 yr–1 with an average rate of 65 kg P ha–1 yr–1. Superphosphate is generally applied in late spring and over summer. The mean stocking rate is 3.4 cows ha–1, which return in dung 10 to 23 g P d–1 cow–1 to the soil surface (Haynes and Williams, 1993).


Figure 1
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  		Fig. 1. Elevation (m above sea level) and location of sampling site within the Waiokura catchment, Taranaki, New Zealand.

 
Stream Water Collection and Analysis
The site has been monitored since 2001 as part of multi-watershed study following changes in water quality as BMPs are adopted by dairy farmers (Wilcock et al., 2006). The sampling site was located at the catchment outlet (Fig. 1), where a water level recording station gave a continuous record of stream flow. Samples of stream water and time-integrated SS using in situ samplers were taken monthly for 1 yr beginning at the start of April 2005. Water samples (1 L) were analyzed for dissolved reactive P (DRP, P measured following filtration through a 0.45-µm filter) and TP, after a persulphate digestion, by flow injection analysis and colorimetry (APHA, 1998). A measure of SS at the time of sampling was also gained by filtration of a 2-L sample through a Whatman GF/F filter paper and weighing the oven-dried residue (APHA, 1998).

Loads and specific yields for SS, TP, and DRP were calculated using flow-weighted mean concentrations and continuous flow data (Fergusson, 1987).

Suspended Sediment and Soil Sampling, and Analysis
Six time-integrated sediment samplers, based on the design of Phillips et al. (2000) were installed at 0.6 median water depth (recorded over the last 5-yr data) and attached to steel uprights by rubber ties. Briefly, samplers operate in situ and each one consists of a narrow (2-mm diam.) inlet and outlet tube, with a larger cavity (48 mm diam.) in between. This cavity (volume 2.5 L) has a cross sectional area approximately 600 times greater than the inlet and outlet tubes and drastically decreases water flow velocity relative to ambient flow, thereby enhancing sedimentation. Samplers can capture as much as 70 to 80% of the sediment that passes through them, thus obtaining a sample representative of SS over a defined time period for small streams (Phillips et al., 2000; Russell et al., 2000). They have successfully been used previously for sampling SS and associated P in a similar size stream (McDowell and Wilcock, 2004).

Each month (± 2–4 d), sediment samplers were removed, sediment retained within the cavity collected, and the sampler flushed with stream water before being tied back in place. Due to flooding in January (Fig. 2) 2006 two samplers were swept away and the resulting samples missed. A 20-mL subsample was placed in a centrifuge tube along with a Fe oxide strip designed to sequester bioavailable P (method 4; Bramley and Roe, 1993). This suspension was shaken overnight, the strip removed, any particles washed off back into the centrifuge tube, and P desorbed from the strip by shaking it with 30 mL of 0.1 M H2SO4 for 1 h. The P released is termed bioavailable P (BAP) and determined via colorimetry of a neutralized aliquot (as per APHA, 1998). Total P, in trapped SS, was determined on another subsample following an aqua regia digest (Crosland et al., 1995) and organic C (OC) via a LECO total C/N analyzer. Sediment from the remainder of the sample was isolated from the collected samples via filtration through a 0.7 µm borosilicate microfiber filter (GF75-MFS Advantec Inc., Pleasanton, Ca, USA), air-dried, and the residue weighed.


Figure 2
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  		Fig. 2. Flow and concentrations of dissolved reactive P, total P, and suspended sediment at the catchment outlet. Autumn = Mar. through May, Winter = June through Aug., Spring = Sept. through Nov., Summer = Dec. through Feb.

 
Four topsoil (plow layer, 0–20 cm depth) samples were taken in 10 paddocks (fields) abutting the stream channel in January. At the same time, subsoil samples were taken from 20- to 100-cm depth to encompass likely drainage pathways and four samples each of stream bed and bank sediment (surface 1 cm). Soil and sediment were air-dried, ground, and sieved <1 mm before analysis of BAP, OC, and TP as above. Soils (500 g), stream bed and bank sediment, and trapped sediment samples that had been bulked (up to 250 g) by season were also analyzed for 137Cs concentration via {gamma}-spectrophotometery. Counting of soils lasted for 7.5 h and sediments for 23 h (Canberra BE5030, Canberra Industries Inc., Meriden, CT). Data for 137Cs, TP, and OC in trapped sediments were input into a mixing model and compared against concentrations in topsoil, subsoil, and stream bank and bed sediments to determine their origin.

For particle size analysis, soil, sediment, and trapped sediment bulked by season were dispersed using a high speed homogenizer (25 000 rpm, Kinematica, Polytron-Aggregate, Luzern, Switzerland) for 5 min. Samples were analyzed for % fines (particles <63 µm) by a laser particle sizer (Malvern Mastersizer, Malvern Instruments Ltd, Malvern, UK).

Summary statistics and analysis of variance were generated with Genstat v8.0 (Genstat Committee, 2006).


   RESULTS AND DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
CONCLUSIONS
 REFERENCES
 
Phosphorus Export in Stream Flow
A summary of flow, sediment, and P concentrations at the catchment outlet is given in Table 1. Flow data exhibited a typical pattern for a temperate coastal catchment with greatest mean flow rates in winter and spring and lower mean flow rates during summer and autumn. This was paralleled by SS and TP concentrations (Table 1). Mean seasonal flow and concentrations of SS, DRP, and TP collected during the study period were not significantly different from means for all of the data collected during the last 5 yr. A two-way ANOVA indicated that concentrations during each season of the study period were not significantly different from the 5-yr mean—there was no significant interaction between season and when samples were taken.


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  		Table 1. Mean flow rate, suspended sediment (SS), dissolved reactive P (DRP), and total P (TP) concentrations in stream flow for each season during the study period and for the last 5 yr. LSD05 is the least significant difference at P < 0.05.

 
Concentrations of SS over the entire monitoring period (2001–present) were rarely <15 mg L–1 (Fig. 2). In an earlier study of a similarly flat (slope <5%) dairying catchment (Bog Burn, Southland, New Zealand) McDowell and Wilcock (2004) found that SS concentrations rarely exceeded 10 mg L–1. Differences between the two catchments were: (i) the Waiokura receives much more precipitation at its head waters (2200 mm) compared with the Bog Burn (mean annual rainfall 1100 mm), (ii) soils are of volcanic-ash origin in the Waiokura and of sedimentary origin in the Bog Burn, and (iii) there is less fencing of streams in the Waiokura catchment (Betteridge et al., 2005). Furthermore, the Waiokura stream has numerous tributaries (Fig. 1) and consequently, a much greater stream length and density than Bog Burn (Monaghan et al., 2007). Despite the high concentrations of SS in stream flow, the annual yield for the hydrologic year (April–March) was calculated to be 155 kg ha–1, which is low compared with a range of New Zealand catchments in pastoral agriculture (20–2000 kg ha–1; Wilcock, 1986; Wilcock et al., 1999; Vant, 2001). This is probably due to the volcanic lithology, which tends to adsorb rainfall and decrease the intensity of sediment-producing flood events. The measured sediment yield was similar to values calculated for other dairying catchments in New Zealand with similar rainfall and topology (viz. 38–231 kg ha–1yr–1) (Wilcock et al., 2006).

Yields for DRP (0.159 kg ha–1) and TP (0.715 kg ha–1) were near the middle of their respective ranges (0.010–0.300 kg DRP ha–1; 0.100–1.700 kg TP ha–1) for New Zealand catchments in pastoral agriculture (Wilcock, 1986; Wilcock et al., 1999, 2006; Vant, 2001). However, concentrations of DRP (0.021–0.053 mg L–1) and TP (0.064 to 0.255 mg L–1) during the study period were well in excess of the recommended guidelines (0.01 mg L–1; 0.033 mg L–1 for DRP and TP, respectively) for slightly disturbed lowland ecosystems in New Zealand (ANZECC, 2000). Concentrations of TP during the study period were related to the concentration of SS (TP = 0.002SS + 0.06; R2 = 0.44, P < 0.05), and it is reasonable to assume that the two were linked. In contrast, no such relationship existed between SS and DRP because dissolved P inputs are often unrelated to sediment inputs to rural waterways (viz. inputs of dissolved P via fertilizers applied during summer and autumn [Betteridge et al., 2005], direct inputs of dissolved P from dung, and inputs from effluent ponds). Furthermore, changes in runoff P concentrations from pastures, which are largely in dissolved forms (Nash et al., 2002), reflect the natural variation in soil P concentrations, which for silt loam soils tend to peak in summer (Saunders and Metson, 1971; McDowell and Trudgill, 2000). Previous work has also shown that the major form of P within dairy shed effluent is orthophosphate-detectable as DRP in flow (McDowell et al., 2005). Discharge of effluent from ponds can occur year round, but would be diluted by larger stream flows in winter and spring than in summer or autumn.

Of the DRP released into stream flow in this study, more is likely associated with effluent discharge from ponds than with overland or subsurface flow from topsoil or subsoil. Monaghan et al. (2004) estimated that about one-third of total P losses were attributed to effluent ponds. Furthermore, in a study of a nearby (30 km east) volcanic ash-derived soil, Morton et al. (2003) found that DRP concentration in overland flow from ungrazed plots was <0.01 mg L–1 even in those soils with an Olsen P concentration >60 mg L–1. In comparison, sedimentary soils of the same Olsen P concentration have far fewer Al and Fe hydrous oxides and as a result will support DRP concentrations in overland flow that are 4 to 20 times greater. The high total P of the sampled and trapped sediments and soils (Evans et al., 2004; Tables 2 and 3) illustrates the importance of metal oxyhydroxides in sorbing P within the Waiokura catchment.


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  		Table 2. Mean sediment load (per sampler) and bioavailable P (BAP), total P (TP), and organic C concentrations trapped by samplers for each season. LSD05 is the least significant difference at P < 0.05.

 

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  		Table 3. Mean bioavailable P (BAP), total P (TP), and organic C concentrations in topsoil, subsoil, and stream bed and bank sediment. LSD05 is the least significant difference at P < 0.05.

 
The seasonality of DRP in stream flow during the study period was emphasized by the ratio of DRP to TP, which was 35% in summer and autumn and about 17% in winter and spring (Table 1). Several authors have concluded that the majority of P retention occurs during periods of low flow associated with the settling of PP. For Brunet and Brian-Astin (1998) this was spring and summer, while for May et al. (2001) it was spring and early autumn, and during summer for Evans and Johns (2004). Conversely, the majority of P loss occurred during storm events, usually in winter. Patterns in our DRP and TP data concur with these findings.

Variation of Particulate Phosphorus Loss
Suspended solids concentration increases with increasing flow rate, as does the quantity of sediment collected by the samplers (Phillips et al., 2000). Trapped sediment (g) was related to cumulative flow, Q{Delta}T (m3) using the continuous flow data, thus

Formula 1[1]

To estimate the efficiency of sediment trapping, the quotient of the sum cross-sectional area of sampler inlets (n = 6) and the cross-sectional area of the stream channel (assuming that average depth was that for the annual median flow rate) was multiplied by annual SS load (363 Mg). This would have yielded 2 kg of trapped SS if 100% efficient. However, the actual trapped load was 0.49 kg giving an efficiency of 25%. Error around this value would be caused by blockage of narrow inlets and not accounting for greater cross-sectional area during high flow when most sediment transport occurs. Inlet blockage was less of a problem for McDowell and Wilcock (2004) who used samplers of twice the inlet diameter to capture an estimated 42% of SS in stream flow.

In contrast to many catchments dominated by pasture, P loss in the Waiokura appears to be dominated by PP (Wilcock et al., 2006). However, while trapped sediment load paralleled flow and was greater in winter and spring than summer or autumn (Fig. 3), there was also a significant negative power relationship (P < 0.05) between BAP or TP concentration and the mass of sediment trapped (BAP = 86SS–0.53, R2 = 0.67; TP = 3487SS–0.17, R2 = 0.52). This indicates that small amounts of trapped sediment have a greater P concentration than large amounts due to the dominance of finer P-rich material (Cooke, 1988; Sharpley and Smith, 1990). A split-line model (McDowell and Sharpley, 2001) showed that BAP and TP were significantly enriched when <2 g of sediment was trapped; this occurred in 20% of the samples taken, and usually in autumn. This probably reflected a combination of the erosion and transport during high flow of low P coarse particles that settle out and allow P-enriched fine particles to dominate sediment transported during low, steady flow, which is more likely in autumn (Table 1). Indeed, Table 2 shows that, like DRP in stream flow, BAP in trapped sediment was significantly greater in autumn than in other seasons. However, compared with other studies, the relative proportion of BAP to TP in SS was much lower; in our study it ranged from 1% in winter to 4% in autumn (Fig. 3; Table 2), whereas most others have found ranges from 7 to 50% (Pionke and Kunishi, 1992; Sharpley et al., 1995; Steegen et al., 2001; Uusitalo et al., 2001; McDowell and Wilcock, 2004). Apart from some management and climatic differences the biggest contrast is due to soil type. Our study is the only one that has been conducted on Al and Fe hydrous oxide-rich soils derived from volcanic ash. As a result, the sediment derived from these soils contains much P that is not bioavailable. The total P concentrations in unfertilized soils in Taranaki are commonly in excess of 2000 to 2500 mg kg–1 (Saunders, 1968)—but little of this P is bioavailable since it is bound so tightly. The concern may arise if PP is transported to a lake or reservoir which may stratify and be anaerobic at the bottom. In this case a proportion of the P bound to Fe oxides, about 20 to 40% of total P (McDowell and Condron, 2000), would be liberated into the water column.


Figure 3
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  		Fig. 3. Mean (± 95% confidence intervals) concentrations of bioavailable P, total P, and suspended sediment trapped within the samplers (n = 6) each month. Autumn = Mar. through May, Winter = June through Aug., Spring = Sept. through Nov., Summer = Dec. through Feb.

 
During November and December, P concentrations in SS were elevated (Fig. 3), while flow or the quantity of SS trapped was not excessively low. Indeed, no significant differences occurred between the yield of trapped SS between spring, summer, or autumn (Table 2). Hence, a sudden increase in P concentrations due to more fines would not have been the cause. More likely is that eroded sediment had received fresh superphosphate applications during late spring, which elevate soil and solution P concentrations especially in the first few weeks after broadcasting the fertilizer (Nash et al., 2004). Enrichment could then have occurred via overland or subsurface inputs of P-enriched particulates or DRP sorption by SS.

Origin of Particulate Phosphorus Loss and Management
Table 2 shows that the TP concentration of trapped sediment varied little, while BAP concentration was significantly greater in autumn. Table 3 indicates that topsoil and bank sediment had similar TP concentrations to trapped sediment but that all sources considered had BAP concentrations that greatly exceeded that in trapped sediment. Stable isotopes such as Cs137 have been used as tracers of sediment sources (e.g., He and Walling, 1996). Like P, 137Cs is more concentrated in the clay fraction (Uusitalo et al., 2001), and subject to sediment size sorting. The equations of He and Walling (1996) can be used to alter concentrations by accounting for particle size. In our samples the concentration of fines (<63 µm diam.) in trapped sediment was about 84% varying only 2% between winter and autumn, while topsoil, subsoil, stream bank, and bed sediment contained 60, 53, 15, and 55% fines, respectively. After adjustment, the resulting data (Fig. 4) indicated that the 137Cs concentration of trapped sediment did not vary significantly (P > 0.05) between seasons, but more importantly trapped sediment was similar only to stream bank sediment, and not to topsoil, subsoil, or stream bed sediment. These data were then incorporated into a mixing model (Phillips and Gregg, 2003) together with TP and OC. The output (Table 4) gives the range of possible solutions for mixing the sources together to get a sediment mixture the same as that of trapped sediment. Ranges are given and not means, as reported in other studies, since mixing models such as Isosource (Phillips and Gregg, 2003) that model many sources can have a wide range of possible solutions, none of which should have precedence over others unless there is strong supporting information to suggest otherwise. That said these data show that topsoil is probably a dominate source of sediment in summer and to a lesser degree in autumn, whereas in winter and spring, bank sediment contributes at least 20% of sediment (i.e., range never extends to zero), but in summer and autumn, bank sediment is probably a small contributor of sediment in stream flow.


Figure 4
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  		Fig. 4. Cesium-137 activity in topsoil, subsoil, bed, and bank sediment and suspended sediment captured by in-stream samplers each season. Error bars are 95% confidence intervals.

 

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  		Table 4. Range of probable solutions for the percentage of source materials in trapped sediments by season.

 
Management of Phosphorus Losses
The sediment transported in streams and rivers comes from overland and subsurface flow (including drainage), subsoil, stream banks, or stream beds. The proportion of each will depend on catchment characteristics and management. For instance, Wallbrink et al. (2003) showed that for a large forested catchment in Southeastern Australia with steep (>20%) slopes, a mean of 62% of P loss was from subsoil erosion of gullies. In contrast, Chapmann et al. (2001, 2005), Foster et al. (2003), Laubel et al. (1999), McDowell and Wilcock (2004) and Uusitalo et al. (2001) have all shown that for gently sloping (slope <5%) tile-drained land, sediment P loss via drainage or overland flow was topsoil-derived, but give no mean proportion. Slope in the Waiokura is similarly low, and while rainfall at the headwaters is high compared with the rest of the catchment, the volcanic lithology tends to adsorb rainfall and decrease the size of floods. This suggests that topsoil or subsoil subjected to subsurface flow should have been a greater contributor to PP and SS than was found in spring and winter (i.e., range should not have extended to zero).

One reason for this may be that in heavily managed catchments where stream channels have been straightened, studies have found the major source of PP loss is associated with stream bed and bank erosion (Kronvang et al., 1997a, b). Some stream straightening and piping of the stream beneath paddocks has occurred in the Waiokura catchment (Betteridge et al., 2005), but the majority of tributaries leading into the main stem are unaltered. However, banks in the Waiokura catchment tend to be steep. Laubel et al. (2003) found that for 15 first- and second-order streams in Denmark the majority of sediment and P loss was due to stream bank erosion and increased with bank steepness.

Another reason is damage by cattle trampling. For instance, Foster et al. (1988) found during a study of sediment delivery from the Seeswood Pool drainage basin in the English Midlands that the major source was channel erosion caused by trampling of stream banks. Kronvang et al. (1997b) also considered trampling of stream banks by cattle as a major mechanism of PP and SS loss in the Gelbæk catchment, Denmark, but were unable to quantify it. In contrast, Laubel et al. (2003) found that in a study of 91 stream banks, erosion was linearly related to bank angle and was decreased by buffer zones, woodland riparian areas, and fencing off stream banks from cattle grazing.

Current management in the Waiokura sees about one-third of the stream channel within dairy farms either fenced only on one side or not fenced at all (Betteridge et al., 2005). The upper catchment is characterized by numerous small tributaries (Fig. 1) many of which are accessible to grazing livestock. Furthermore, about 10% of farmers will allow grazing within the fenced-off area during winter when stream flow is greatest. This practice will clearly exacerbate sediment and PP loss. A major mitigation practice to decrease the influence of cattle grazing on sediment and PP loss is therefore to expand the length of stream channel that is fenced-off and restrict grazing of these areas, especially during winter. In addition, there are 107 culvert and bridge crossings over the stream network and often these can represent important conduits of sediment from overland flow to streams (Gruszowski et al., 2003). Diversion of feedpad, silage pit, track, and bridge runoff into a detention storage dam may alleviate some PP loss. Decreasing soil Olsen P concentrations will also be effective in the mitigation of P losses via erosion in PP overland and subsurface flow of paddocks in summer and autumn, especially since soils in the catchment are P-enriched (Olsen P concentration >70 mg P L–1 and well in excess of the agronomic optimum of 30 mg P L–1; Roberts and Morton, 1999) (Table 2).

While fencing off stream banks may mitigate the influence of cattle grazing it does not influence the effect of stream flow on stream banks. Pasture has a greater density of fine roots than woody vegetation, which provides a greater critical shear stress, but only reinforces the bank to rooting depth; commonly the top 10 to 15 cm (Thorne, 1990). In contrast, trees have fewer fine roots but a greater rooting depth. Wynn et al. (2004) hypothesized that root density at the bank toe was the most important factor influencing bank resistance to scouring since shear stress increases with stream depth. Indeed, they found that compared with pastures, trees provided greater protection against scouring and under-cutting. Presently, of the 67.9 km of stream channel in the Waiokura, 32 km is under trees, wetland, or native scrub. In areas where trees are not planted because the land would yield a better return as grazed pasture, simple in-stream structures such as bendway weirs and rock stream barbs may aid in bank stabilization (Evans and Kinney, 2000). Davies-Colley (1997) has shown that small streams in pasture catchments tend to be narrow by comparison with streams with mature riparian trees.


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
REFERENCES
 
Concentrations of DRP in stream flow and BAP in trapped sediment followed a seasonal pattern, being significantly greater in autumn than other seasons. This was attributed to a combination of sediment size sorting due to a low and stable flow regime and possible inputs from effluent ponds and soil via superphosphate application. In contrast, while much more sediment was captured during winter and spring when flow was greater it was much less P-rich. Analysis of 137Cs and mixing modeling indicated that trapped sediment and P more likely originated from stream banks in winter and spring, not topsoil, subsoil, or the stream bed. This was attributed to the scouring of relatively steeply sloped banks—especially during winter and inputs via cattle trampling and dung deposition. Phosphorus-enriched overland flow probably from topsoil or laneways where dung is routinely deposited by cattle was highlighted as a source of P in summer and autumn transported to the stream via overland and subsurface flow, while it was likely that the contribution from stream bed and bank sediments was minimal. At present, P concentrations in stream flow are in excess of current limits for good freshwater quality in New Zealand. Therefore, to decrease P concentrations, management should focus on maintaining stream bank stability by a combination of fencing and riparian planting with trees and shrubs, ensuring overland flow from crossings and lanes does not enter the stream, and decreasing Olsen P in topsoil to the agronomic optimum.


   ACKNOWLEDGMENTS
 
Funding for this work was provided by the New Zealand Foundation for Research, Science, and Technology under contracts C10X0320 and C01X0215. We are also grateful for funding from Dairy Insight, New Zealand. The assistance of staff at NIWA Wanganui and Max Gibbs of NIWA, Hamilton is also gratefully acknowledged.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 





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